United States Department of Agriculture

Riparian Research and Management: 

Past, Present, Future: Volume 1

Forest Rocky Mountain General Technical Report

Service Research Station RMRS-GTR-377 November 2018
Front cover: A backwater with beaver dam along the Gunnison River at Neversink, Curecanti National Recreation Area, Gunnison County, 

Colorado (photo by Kenneth J. Kingsley).

All Rocky Mountain Research Station publications are published by U.S. Forest Service employees and are in the 

public domain and available at no cost. Even though U.S. Forest Service publications are not copyrighted, they 

are formatted according to U.S. Department of Agriculture standards and research findings and formatting cannot 

be altered in reprints. Altering content or formatting, including the cover and title page, is strictly prohibited.

Johnson, R. Roy; Carothers, Steven W.; Finch, Deborah M.; Kingsley, Kenneth J.; Stanley, John T., tech. eds. 2018. 

Riparian research and management: Past, present, future: Volume 1. Gen. Tech. Rep. RMRS-GTR-377. Fort Collins, 

CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 226 p.

doi: https://doi.org/10.2737/RMRS-GTR-377 

Abstract

Fifty years ago, riparian habitats were not recognized for their extensive and critical contributions to wildlife and the 

ecosystem function of watersheds. This changed as riparian values were identified and documented, and the science 

of riparian ecology developed steadily. Papers in this volume range from the more mesic northwestern United States to the 

arid Southwest and Mexico. More than two dozen authors—most with decades of experience—review the origins of 

riparian science in the western United States, document what is currently known about riparian ecosystems, and project 

future needs. Topics are widespread and include: interactions with fire, climate change, and declining water; impacts from 

exotic species; unintended consequences of biological control; the role of small mammals; watershed response to 

beavers; watershed and riparian changes; changes below large dams; water birds of the Colorado River Delta; and 

terrestrial vertebrates of mesquite bosques. Appendices and references chronicle the field’s literature, authors, “riparian 

pioneers,” and conferences.

Keywords: riparian, ecosystem, ecology, riparian processes, restoration, aquatic, arid, semi-arid, upland, freshwater, 

groundwater, hydrology
Technical Editors

R. Roy Johnson, Research Associate, Museum of Northern Arizona, Emeritus Professor, University of Arizona, and retired 

Research Scientist, National Park Service, Tucson, Arizona.

Steven W. Carothers, Founder, Senior Scientist, and Member of the Board of Directors of SWCA Environmental 

Consultants, Phoenix, Arizona.

Deborah M. Finch, Biological Scientist and Program Manager, USDA Forest Service, Rocky Mountain Research Station, 

Albuquerque, New Mexico.

Kenneth J. Kingsley, Retired Senior Scientist, SWCA Environmental Consultants, Tucson, Arizona, and Volunteer 

Scientist, Black Canyon of the Gunnison National Park, Montrose, Colorado.

John T. Stanley, Restoration Ecologist, WWWRESTORATION; formerly Principal Ecologist at Harvey & Stanley 

Associates, Inc.; Principal Restoration Ecologist, The Habitat Restoration Group; and Program Manager, Tahoe Regional 

Planning Agency.

Contributing Authors

Steven W. Carothers, Founder, Senior Scientist, and Member of the Board of Directors of SWCA Environmental 

Consultants, Phoenix, Arizona.

Hisham El Waer received his Ph.D. at the University of Denver in Biological Sciences in 2013, after which he returned 

to Libya as faculty at the University of Al-Fateh, Faculty of Agriculture, Department of Range and Forestry. He has been 

unreachable since his return to his home country. The research presented here was work for his dissertation.

Helen C. Fairley, Sociocultural Program Manager, Grand Canyon Monitoring and Research Center, U.S. Geological 

Survey, Flagstaff, Arizona.

Deborah M. Finch, Biological Scientist and Program Manager, USDA Forest Service, Rocky Mountain Research Station, 

Albuquerque, New Mexico.

Suzanne C. Fouty, Hydrologist and Soils Specialist, USDA Forest Service, Wallowa-Whitman National Forest, Baker City, 

Oregon.

Jonathan M. Friedman, Research Scientist, the Fort Science Center, U.S. Geological Survey, Fort Collins, Colorado.

Jennifer K. Frey, College Professor and Member of the Graduate Faculty, Department of Fish, Wildlife, and Conservation 

Ecology, New Mexico State University in Las Cruces, New Mexico.

Stanley V. Gregory, Distinguished Professor of Fisheries—Emeritus, Department of Fisheries and Wildlife, Oregon State 

University, Corvallis, Oregon.

Robert H. Hamre, retired Leader, Research Information Group at the USDA Forest Service Rocky Mountain Research 

Station, Fort Collins, Colorado.

Annie Henry, Ph.D. student, University of Denver, Colorado.

Osvel Hinojosa-Huerta, Director of the Water and Wetlands Conservation Program for Pronatura Noroeste, a non-profit 

organization, Ensenada, B.C., Mexico.

Elaine E. Johnson, Project Leader, Southwest Arizona National Wildlife Refuge Complex, Yuma, Arizona.

R. Roy Johnson, Research Associate, Museum of Northern Arizona, Emeritus Professor, University of Arizona, and retired 

Research Scientist, National Park Service, Tucson, Arizona.

Kenneth J. Kingsley, Retired Senior Scientist, SWCA Environmental Consultants, Tucson, Arizona, and Volunteer 

Scientist, Black Canyon of the Gunnison National Park.

Mary Anne McLeod, Senior Scientist, SWCA Environmental Consultants, Flagstaff, Arizona.
ii

Eric Mellink, Senior Researcher, Centro de Investigación Científica y de Educación Superior de Ensenada, B.C., Mexico.

Katie Merewether, Avian Field Biologist, Bird Conservancy of the Rockies, Brighton, Colorado.

Duncan T. Patten, Emeritus Professor of Life Sciences, Arizona State University, Founding President, Arizona Riparian 

Council, and Emeritus Research Professor and past Director of the State Water Center at Montana State University, 

Bozeman, Montana.

John S. Richardson, Department Head, Forest and Conservation Sciences, University of British Columbia, Vancouver, 

B.C.

Anne Sands, Aquatic Biologist, editor of first riparian conference proceedings published in the United States; Woodville 

Ranch, Dogtown, California.

Michael L. Scott, Adjunct Faculty, Watershed Sciences Department, Utah State University, Logan, Utah.

Bo Shelby, Professor, Department of Forest Ecosystems and Society, Oregon State University, Corvallis, Oregon.

Anna A. Sher, Professor, Department of Biological Sciences, University of Denver, Denver, Colorado.

D. Max Smith, Research Associate, USDA Forest Service, Rocky Mountain Research Station, Albuquerque, New Mexico; 

lives in Portland, Oregon.

John T. Stanley, Restoration Ecologist, WWWRESTORATION; formerly Principal Ecologist of Harvey & Stanley 

Associates, Inc.; Principal Restoration Ecologist, The Habitat Restoration Group; and Program Manager, Tahoe Regional 

Planning Agency, Paradise, California.

Frederick J. Swanson, Pacific Northwest Research Station, USDA Forest Service, Corvallis, Oregon.

Raymond M. Turner, retired from the U.S. Geological Survey, Tucson, Arizona.

Robert H. Webb, Adjunct Faculty, University of Arizona, Tucson, Arizona.

William E. Werner, Wildlife Biologist, Bureau of Land Management, Phoenix, Arizona; formerly Environmental Program 

Manager, Arizona Department of Water Resources; retired Habitat Specialist, Arizona Game and Fish Department.
iii

Foreword

The availability of drinking water is the paramount environmental issue of the 21st century in the western United States and 

especially in the Southwest. Growing populations in Albuquerque, Phoenix, Tucson, Las Vegas, and Southern California are 

already testing water supply and delivery infrastructure in the face of just a moderate drought. Conservation in major cities 

and innovations in agriculture have greatly reduced per capita and per acre water use over the last 15 years, providing a 

margin of security to water supplies. But as populations continue to grow, it’s important for the people of the Southwest 

to reflect upon and confront the stresses caused by modest drought. Now is the time to collaborate to develop responses 

and put in place coordinated systems that can handle deeper drought, reduced water availability, and greater variability in 

precipitation predicted in upcoming decades.

Water delivery in the arid Southwest relies on the connection of forests, particularly the National Forests, to water storage 

facilities and eventually the faucets in our homes. The National Forests are a small part of the watershed, but they are 

disproportionally where rain and snowfall can become the water supply the system relies upon. Forests, streams, and 

riparian areas are the green infrastructure that captures, stores, and releases water, which is then delivered to our cities 

through the dams, canals, and pipelines we’ve built in support of our daily lives. Without the water supply provided by the 

green infrastructure, the built infrastructure is useless.

Just as the built infrastructure has been recently tested by drought, the green infrastructure is now being tested by fire 

and land use practices. Can the green infrastructure reliably provide human communities drinking water, wildlife habitat, 

and other critical ecosystem services? Can riparian ecosystems in particular produce the expected level of services as 

populations increase, and the climate changes, with the Southwest becoming both warmer and drier? Healthy riparian 

systems are linchpins connecting land and water, and they are integral to future water availability. This compendium of 

research on riparian areas could not be timelier or address a more essential need in the West.

In the past couple of decades, researchers, land managers, and regulators have focused on the holistic management of 

ecosystems, watersheds, and meeting the needs of human communities, while still addressing individual species, habitat 

components, and services. Community leaders and land managers have also embraced collaborative planning and 

cooperative solutions across broader landscapes. They have worked in a more inclusive manner to look past previous 

polarized thinking to see shrinking streams and riparian systems and value them beyond managing for individual interests. 

This publication, guided by the inseparable nature of streams and riparian ecosystems, emphasizes the interrelationships 

and continuity of riparian areas along with dependent wildlife and human services.

The scientific papers in this General Technical Report continue the long, demanding, and now urgent task of conveying 

scientific information on riparian systems, organisms, and their human interactions to give the reader a better sense of 

the history, conditions, and working of these resources, along with potential solutions for today’s challenges. This report 

represents the current state of knowledge and points to some essential steps for connecting science, management, 

and politics for the restoration and sustenance of riparian ecosystems in the West and the sustainability of the human 

communities that depend on them.

Calvin Joyner

Regional Forester

Southwestern Region

USDA Forest Service
iv

Preface—The Western Riparian Project

“You can’t talk about streams without talking about riparian ecosystems.” 

(David L. Rosgen, Leader in Stream Classification and Stream Restoration; Western Riparian Project Science Advisor)

“Most impacts are insidious and cumulative.” 

(Christopher C. Estes, At-Large Director, Instream Flow Council; Western Riparian Project Science Advisor)

The Western Riparian Project (WRP) is a riverine project with emphasis on riparian ecosystems in western North America. 

Our objectives are:

• To review the historic needs for and beginnings of the science of riparian ecology;

• To assess the current state of riparian research and management; and

• To lay a foundation for the planning of conservation and research activities.

To attain these objectives, the WRP consists of two basic parts: (1) publishing a General Technical Report of selected 

papers that review the past 50 years of research and management activities to be distributed electronically and in print, and 

(2) convening a conference with invited speakers consisting of many of the report chapter authors. A conference is planned 

near the time of the release of this General Technical Report. Speakers will be selected by the Steering Committee. This 

General Technical Report is to be issued by the Rocky Mountain Research Station of the Forest Service, U.S. Department 

of Agriculture, in two volumes. More than 50 scientists from Mexico to Alaska contributed to this effort, which covers a broad 

spectrum of issues from “phreatophyte control” to current and future impacts of the newly introduced tamarisk beetles 

(Diorhabda spp.) on naturalized and native riparian ecosystems. The effort also covers implications of global climate 

change for riparian habitat restoration and conservation.

This discussion of riparian ecosystems uses a watershed approach that recognizes connectedness of the entire waterdriven continuum, from atmospheric moisture to instream flows and groundwater. The term “riparian” is used to describe a 

range of conditions:

Wet riparian ecosystems (hydroriparian and mesoriparian) associated with lotic waters—on the banks of flowing and 

intermittent streams. These range from the Rio Grande, Colorado, Gila, and Sacramento rivers, some of the nation’s largest 

rivers, to the Salt, Bill Williams, and San Pedro rivers, southwestern streams that are perennial in the upper portions of their 

watersheds but become intermittent downstream.

• Ecosystems associated with lentic waters—on the shores of lakes.

• Xeroriparian ecosystems—along banks of usually dry ephemeral streams such as desert washes.

In short, we are concerned with biotic and abiotic factors related to biologically available water. Our watershed approach 

emphasizes both the instream flow model of connectivity as well as the connectedness of the hydrologic cycle. Our motto, 

“You can’t talk about streams without talking about riparian ecosystems,” acknowledges the importance of healthy riparian 

ecosystems to the health of a stream. This motto is also reversible, i.e., “You can’t talk about riparian ecosystems without 

talking about streams,” thereby acknowledging the interconnectedness of water and instream flows to riparian ecosystems. 

Thus, aquatic ecosystems and instream flow characteristics are considered an integral part of the riparian story. This 

approach emphasizes connectivity in time and space as proposed by the Instream Flow Council—longitudinal, lateral, and 

vertical space—and the natural hydrograph, addressing the temporal factors of seasonality.

The connectedness of water is conceptualized as the hydrologic cycle, often referred to as “the water cycle,” detailing 

the movement of water from the earth’s surface into the atmosphere, through evaporation, and finally back to the earth’s 

surface as precipitation, much of it then percolating underground. In riverine systems, biotic and abiotic factors associated 

with aquatic and riparian ecosystems, and even interconnected upland ecosystems, are inseparably linked in a feedback 

loop related to water and instream flows. A thorough analysis of these systems requires information from ecology, 

hydrology, geomorphology, biochemistry, and a multitude of related disciplines.

We use the term “regeneration” to include both natural reestablishment and humanly assisted restoration of riparian 

habitats, while “restoration” is generally used only to indicate intentional human-induced changes. Our project includes 
v

practical field applications for riparian ecosystem restoration and riparian zone management as well as observations of 

naturally occurring habitat reestablishment as evidenced by natural revegetation and wildlife repopulation of recovering 

riparian areas. We include these findings with the goal of encouraging integration of riparian ecology and wildlife 

conservation with the routine operations of land and water resource management agencies at all levels of government as 

well as non-governmental groups.

The geographic area to be covered in this phase of the project extends from California eastward to Trans-Pecos Texas and 

adjacent northwestern Mexico and from there, northward, to the Pacific Northwest but emphasizing the more arid and semiarid regions of the West. Our reason for emphasis on this region is the suite of biophysical differences compared to regions 

of the eastern United States. We particularly emphasize the Rio Grande, Colorado, Gila, and Sacramento-San Joaquin 

watersheds because of the extensive change effected by anthropogenic activities in their riparian and aquatic ecosystems. 

In addition, a great deal of research, related to regeneration and restoration, has been and continues to be conducted along 

these rivers.

This General Technical Report is the book part of the Project and covers three primary topics:

• Review of the development of the science of riparian ecology, the history of riparian research, and the riparian 

movement in the region from their beginnings in the 1960s until the present;

• Determination of the current status of research and of riparian ecosystem conditions, in relation to ongoing 

management actions (e.g., biological control of Tamarix); and

• Evaluation of needed research studies and management actions in response to climate change and future water 

resource development.

Chapters in the book are designed for widespread applicability to particular riparian issues, examining each topic as it 

pertains especially to the region. A philosophical and practical approach to these issues is presented with examples of 

specific projects used to illustrate wider scale concepts. Chapter authors for the book have been selected because of 

first-hand field experience as well as a robust publication record and/or reports of successful accomplishments. Most of the 

participants have a long-time record of involvement in research and/or conservation of riparian ecosystems.

We consider this the most comprehensive publication yet produced on riparian ecosystems. Unlike most previous efforts, 

our focus is on examining watersheds as a whole to facilitate regeneration of sustainable riparian ecosystems now and into 

the future. Unlike most previous riparian literature that primarily presents a series of reports on specific projects, here we 

use project reports only as illustrations of larger scale riparian issues. We hope that this endeavor will serve to inspire and 

support continuing and future efforts in the study, protection, and restoration of our western riparian resources.

R. Roy Johnson

Steven W. Carothers

Co-Directors

Acknowledgments

This is the first of a two-volume technical report on riparian habitats in the western United States. The idea for this series 

of technical reports was an outgrowth of what is now recognized as The Western Riparian Project, a project initiated by a 

group of riparian ecologists and resources managers, all of whom have decades of experience in the biological aspects 

of riparian habitat productivity, management, and conservation. The Western Riparian Project is in its sixth year and has 

benefitted from participation by more than 50 scientists who have collectively spent thousands of uncompensated hours 

to ensure the success of the Project. Several of us have been involved with riparian ecology since the earliest days of its 

establishment as a science beginning in the late-1960s. The idea for the project originated Thanksgiving week, 2012, during 

a three-way conversation in Yuma, Arizona, between Karen Reichhardt, Bureau of Land Management, Elaine Johnson, 

U.S. Fish and Wildlife Service, and R. Roy Johnson. Reichhardt and E. Johnson had been involved in our Arizona Verde 

River riparian habitat avian field studies almost 50 years earlier (Carothers and Johnson 1970; Carothers et al. 1974; 

Johnson 1971). Now, they suggested it was time to examine the progress made in western aridland riparian ecology and 

conservation since those early studies.
vi

As the idea for a book on the status of southwestern riparian habitat began to fully develop, and we began to focus on what 

we know, what we don’t know, and what we still need to know about the remarkably productive riparian habitats in western 

North America, we naturally revisited the primary literature on the subject and began searching for potential chapter 

authors. Most of the authors of the chapters in this General Technical Report (GTR) have spent decades attempting to 

document the importance of conserving and managing riparian habitats due to their disproportionate value to wildlife and 

people compared to upland habitats.

The Project began in earnest in June of 2014 when 19 Project participants attended a 2-day workshop at the Museum 

of Northern Arizona, Flagstaff. Workshop attendees were from as far as Alaska—Christopher Estes, At-Large Director of 

the Instream Flow Council; New York—Jon Kusler, Associate Director, Association of State Wetland Managers; Oregon—

Suzanne Fouty, USDA Forest Service; Colorado—Robert Hamre, retired, former Leader, Research Information Group, 

USDA Forest Service Rocky Mountain Research Station, Fort Collins, Colorado, who had previously issued proceedings 

of several riparian conferences; Utah—Dale A. Jones, co-editor of the nation’s second riparian conference (Johnson and 

Jones 1977) and retired Director of Wildlife, Fisheries, and Endangered Species, USDA Forest Service, Washington DC; 

and California—F. Thomas Griggs, River Partners.

Travel to the workshop and additional funding was provided by SWCA Inc. Environmental Consultants, Phoenix, Arizona, 

and local arrangements were made by Larry Stevens and Jeri Ledbetter of the Museum of Northern Arizona. In addition to 

many of these participants being involved in ongoing riparian studies, a sense of the history of the development of riparian 

ecology prevailed at that meeting. Six of the 19 workshop attendees had been directly involved in the aforementioned Verde 

Valley riparian studies of the early 1970s: Steve Carothers, Christopher Estes, Elaine Johnson, Kenneth J. Kingsley, Larry 

Stevens, and R. Roy Johnson. Two of the workshop attendees had helped to provide funding for those early studies: Bud 

Bristow, then with Arizona Game and Fish Department, and Dale Jones of the USDA Forest Service. Besides those already 

recognized, we have been fortunate to have scientists and managers of many agencies, universities, and NGOs associated 

with riparian and related issues involved in the Project. Without their input, this work would not be complete. This includes 

Duncan T. Patten, founding President of The Arizona Riparian Council, the nation’s first State riparian council; and three of 

the five founding board members of the Society for Ecological Restoration (SER): John Reiger, Founding President, John T. 

Stanley, and Anne Sands, SER Board Members. Anne also served as a co-director of our project during 2013 and 2014.

Critical assistance has been rendered on an assortment of issues throughout the project by members of our Science 

Advisory Board, including Exequiel Ezcurra, Director, Institute for Mexico and the United States, U.C. Riverside, CA; 

Stefan Lorenzato, Coordinator of the California Riparian Habitat Joint Venture, an association of 18 government and 

NGOs involved in protecting and enhancing riparian habitat; Daniel and Ellen Plunkett, founders of the 7,200 Acre Doris 

Wilderness sanctuary, Belize; Ann Riley, member of the California State Water Quality Control Board; David Rosgen, 

leading authority and practitioner in stream classification and restoration; Bo Shelby, leading researcher on issues related 

to recreation in riparian habitats; and Robert Webb, hydrologist, and one of the leading experts on the Colorado River in 

Grand Canyon.

In addition to appreciation for our co-editors (Deborah Finch, Kenneth Kingsley, and John Stanley), Bob Hamre deserves 

special mention for untold hours spent editing most of the GTR chapters, including several renditions of some chapters. 

Additional reviewers included Dale A. Jones, Duncan T. Patten, Harley G. Shaw, and Patricia M. Woodruff. Deborah Finch 

assisted throughout the project and with the review process by “loaning” us several USDA Forest Service employees 

working for the Forest Service Southwest Regional Office and Rocky Mountain Research Station, including Katelyn P. 

Driscoll, Roy Jemison, Ernest Taylor, and F. Jack Triepke. We also thank David L. Hawksworth, of the Rocky Mountain 

Research Station, for help in reviewing the overall manuscript for formatting errors. Space prohibits us from thanking all 

of the authors, individually, of the various chapters but their affiliations can be found in this GTR’s front matter. Finally, we 

thank Cal Joyner for writing the insightful Foreword to the report.

R. Roy Johnson and Steven W. Carothers

Co-directors

The Western Riparian Project
Contents

Chapter 1. Development of the Science of Riparian Ecology in the Semi-Arid 

Western United States ...................................................................................................................1

Duncan T. Patten, Steven W. Carothers, R. Roy Johnson, and Robert H. Hamre

Introduction...............................................................................................................................................................1

The River Continuum Concept and Riparian Habitat Values Recognized....................................................................3

Anthropogenic Changes to Western Rivers.................................................................................................................4

Early Ecological Studies .............................................................................................................................................4

Focus on Riparian Habitats ........................................................................................................................................6

Hydrology, Instream Flows, and the Natural Hydrograph ...........................................................................................7

Rise and Fall of Tamarix..............................................................................................................................................8

Restoration of Riparian Habitat: Putting the Genie Back in the Bottle.......................................................................12

Summary .................................................................................................................................................................14

References...............................................................................................................................................................16

Chapter 2. Development of Riparian Perspectives in the Wet Pacific Northwest 

Since the 1970s.............................................................................................................................25

Frederick J. Swanson and Stanley V. Gregory

Introduction.............................................................................................................................................................25

Acknowledgments....................................................................................................................................................29

References...............................................................................................................................................................30

Chapter 3. Impacts of Interacting Fire, Climate, and Hydrologic Changes on 

Riparian Forest Ecosystems in the Southwest .........................................................................32

D. Max Smith and Deborah M. Finch

Introduction.............................................................................................................................................................32

Disturbance and Woody Riparian Vegetation........................................................................................................................... 35

Effects of Flooding and Flood Reduction.................................................................................................................................. 35

Drought Effects........................................................................................................................................................................ 36

Fire Effects............................................................................................................................................................................... 37

Middle Rio Grande Case Study ................................................................................................................................37

Natural Hydrological Regime .................................................................................................................................................. 38

Changes to the Regime............................................................................................................................................................ 38

Effects on Riparian Trees.......................................................................................................................................................... 39

Historical/Current Wildfire Effects............................................................................................................................................ 39

Response of Woody Plants to Wildfire ..................................................................................................................................... 40

Ecosystem Implications of the Current Disturbance Regime..................................................................................................... 40

Implications of Climate Change............................................................................................................................................... 41

Conclusions: Disturbance Change and the Future of Aridland Riparian Forests ........................................................41

Acknowledgments....................................................................................................................................................43

References...............................................................................................................................................................43

vii
Chapter 4. Invasion and Restoration of Western Rivers Dominated by Tamarix spp................47

Hisham N. El Waer, Annie Henry, Katie Merewether, and Anna A. Sher

Introduction.............................................................................................................................................................47

Tamarix Introduction and Impact ............................................................................................................................................. 47

Vegetation Response to Tamarix Removal................................................................................................................................ 48

Sampling Methodologies and Invasive Removal Techniques .................................................................................................... 50

Methods and Site Locations .....................................................................................................................................50

Research Findings and Implications .........................................................................................................................52

Regional Differences............................................................................................................................................................... 52

The Response of Tamarix to Removal Efforts ............................................................................................................................ 53

The Response of Plant Communities to Tamarix Removal Efforts.............................................................................................. 55

Conclusion and Implications....................................................................................................................................57

Acknowledgments....................................................................................................................................................58

References...............................................................................................................................................................58

Chapter 5. Unintended Consequences: Tamarisk Control and Increasing Threats 

to the Southwestern Willow Flycatcher......................................................................................62

Mary Anne McLeod

Introduction.............................................................................................................................................................62

The Search for Biocontrol for Tamarisk .....................................................................................................................63

Conflict With the Southwestern Willow Flycatcher...................................................................................................65

Beetle Releases........................................................................................................................................................68

Northern Beetles ..................................................................................................................................................................... 68

Other Beetle Species............................................................................................................................................................... 68

Implementation Phase of Biocontrol Program.......................................................................................................................... 69

Beetles Spread Into Flycatcher Habitat .....................................................................................................................70

Virgin, Muddy, and Lower Colorado Rivers.............................................................................................................................. 70

Rio Grande.............................................................................................................................................................................. 72

Cessation of Biocontrol Releases............................................................................................................................................. 72

Response by Flycatchers to Tamarisk Defoliation......................................................................................................73

St. George, Utah...................................................................................................................................................................... 73

Mormon Mesa, Nevada........................................................................................................................................................... 75

St. George vs. Mormon Mesa .................................................................................................................................................. 77

Short-Term Consequences........................................................................................................................................78

Long-Term Prospects................................................................................................................................................79

References...............................................................................................................................................................81

Chapter 6. Beavers, Livestock, and Riparian Synergies: Bringing Small Mammals 

Into the Picture .............................................................................................................................85

Jennifer K. Frey

Introduction.............................................................................................................................................................85

Taxonomic Bias in Riparian Restoration ...................................................................................................................87

Terrestrial Small Mammal Riparian Communities.....................................................................................................88

Ecological Roles of Riparian Small Mammals...........................................................................................................90

Animal Production and Biomass.............................................................................................................................................. 90

Fuel for Predator Communities................................................................................................................................................ 91

viii
Soil Condition ......................................................................................................................................................................... 91

Plant Community Composition and Succession....................................................................................................................... 92

Terrestrial Small Mammal Community Habitat Relationships ...................................................................................93

Conclusions.............................................................................................................................................................96

Acknowledgments....................................................................................................................................................97

References...............................................................................................................................................................97

Chapter 7. Euro-American Beaver Trapping and Its Long-Term Impact on Drainage 

Network Form and Function, Water Abundance, Delivery, and System Stability.................102

Suzanne C. Fouty

Introduction...........................................................................................................................................................102

Conceptual Model Part 1: Watershed Response to Long-Term Beaver Presence......................................................103

Reductions in Downstream Flood Magnitudes....................................................................................................................... 106

Conceptual Model Part 2: Watershed Response to Beaver Trapping After a Long-Term Presence.............................107

The Geographies of Euro-American Beaver Trapping, General Land Office Surveys, and Early Expeditions ............113

Arroyo Formation in the Southwest and Intermountain West ..................................................................................115

The Role of Beaver Trapping .................................................................................................................................................. 115

The Role of Native Peoples.................................................................................................................................................... 118

Reexamining Changes on the Gila River in Light of Beaver Trapping ......................................................................119

Explaining the Absence of Arroyo Formation From 1750 to 1825 Despite Large Herds of Spanish and Mexican 

Livestock ...........................................................................................................................................................122

The Question of Stream Sensitivity to Climate Variability....................................................................................................... 123

From Past to Present: Placing Hydraulic Geometry Relationships in Their Historic Disturbance Context.................123

From Present to Future ...........................................................................................................................................127

Conclusions...........................................................................................................................................................128

Acknowledgments..................................................................................................................................................129

References.............................................................................................................................................................129

Chapter 8. Arizona as a Watershed—Then and Now: Case Studies of Changed 

Management of Rivers and Habitat in the Lower Colorado River System............................134

William E. Werner

Introduction...........................................................................................................................................................134

Colorado River Below Glen Canyon Dam and the Role of Tributaries ....................................................................137

Alamo Dam and the Bill Williams River.................................................................................................................138

Salt River in Salt River Valley..................................................................................................................................140

Gila River Below Painted Rock Dam ......................................................................................................................142

Colorado River Below Hoover Dam.......................................................................................................................143

Discussion .............................................................................................................................................................144

References .................................................................................................................................................... 144

Chapter 9. Evaluating Riparian Vegetation Change in Canyon-Bound Reaches 

of the Colorado River Using Spatially Extensive Matched Photo Sets.................................148

Michael L. Scott, Robert H. Webb, R. Roy Johnson, Raymond M. Turner, Jonathan M. Friedman, and Helen C. Fairley

Introduction...........................................................................................................................................................148

Hydrologic and Climatic Context...........................................................................................................................149

Riparian Vegetation in the Grand Canyon...............................................................................................................150

ix
First Predictions of Change.....................................................................................................................................153

Methods.................................................................................................................................................................154

Results and Discussion...........................................................................................................................................155

New High-Water Zone.......................................................................................................................................................... 157

Old High-Water Zone ........................................................................................................................................................... 165

Management Implications......................................................................................................................................169

Acknowledgments..................................................................................................................................................170

References.............................................................................................................................................................170

Chapter 10. Breeding Waterbirds of the Mexican Portion of the Colorado River Delta...........174

Eric Mellink and Osvel Hinojosa-Huerta

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

Fresh Water, Riparian, and Marsh Habitat ..............................................................................................................175

Upper Estuary ........................................................................................................................................................178

Montague Island ....................................................................................................................................................179

Conservation Problems ..........................................................................................................................................182

Restoration and Conservation Actions....................................................................................................................183

Concluding Remarks..............................................................................................................................................183

Acknowledgments..................................................................................................................................................184

References.............................................................................................................................................................184

Chapter 11. Terrestrial Vertebrates of Mesquite Bosques in Southwestern North America ...188

R. Roy Johnson, Elaine E. Johnson, and Steven W. Carothers

Introduction...........................................................................................................................................................188

Mesquite Bottomlands as Wildlife Habitat..............................................................................................................188

The Mesquite-Invertebrate-Vertebrate Foodweb......................................................................................................189

Mesquite Bosques and Cottonwood-Willow Riparian Gallery Forests.....................................................................205

Summary ...............................................................................................................................................................206

References.............................................................................................................................................................206

Appendix A: Western Pioneers of Riparian Study Through the 1980s .....................................212

Compiled by: Kenneth J. Kingsley, John T. Stanley, Steven W. Carothers, R. Roy Johnson, Deborah M. Finch, 

Duncan T. Patten, Anne Sands, Bo Shelby, and Frederick J. Swanson

Appendix B: Conferences, Symposia, and Other Gatherings Pertaining To Riparian 

Ecosystems, Riparian Ecology, Riparian Habitat Restoration, and Riparian Area 

Conservation...............................................................................................................................220

Compiled by: John T. Stanley, R. Roy Johnson, Steven W. Carothers, Duncan T. Patten, and John S. Richardson

x
USDA Forest Service RMRS-GTR-377. 2018 1

Chapter 1. Development of the Science of Riparian 

Ecology in the Semi-Arid Western United States

Duncan T. Patten, Steven W. Carothers, R. Roy Johnson, and Robert H. Hamre

Introduction

The science of riparian ecology in the West developed over several decades, 

especially in the Southwest and California, as the importance of this ecosystem, its 

components, productivity, functions, and relationship to system hydrology became better understood. While it seems incredible today, it was only 50 years ago that streamside 

vegetation in the arid Southwest and portions of California were the subject of much 

concern as most flood control and water agencies wanted the permanent removal of 

riparian habitat. The early research within riparian habitats in the 1950s and 1960s 

focused primarily on methods of removal and reasons for its destruction (Gatewood 

et al. 1950; Fox 1977; Horton et al. 1964; Robinson 1952). Most studies of western 

rivers in the 1960s addressed the “fact” that riparian vegetation, referred to mostly as 

phreatophytes (i.e., plants tapping groundwater), was utilizing a large portion of the 

shallow groundwater moving through watersheds. At that time, most water managers 

believed that this water could be better allocated to human activities and needs, primarily agricultural, municipal, and industrial uses. In the semi-arid West, water lost to 

evapotranspiration by watersheds and phreatophytic plants was considered to be water 

unavailable for human use, and thus “wasted.” Consequently, several studies beginning 

in the early 1950s were designed to demonstrate water consumption by watersheds 

(Gottfried et al. 1999) as well as riparian/phreatophytic plant communities with little or 

no concern or focus for other values such as the unique riparian habitat (Decker et al. 

1962; Robinson 1952). In fact, at the time, a U.S. Department of Agriculture scientist 

and leading water expert wrote: “Phreatophyte vegetation seriously affects water supplies in arid and semi-arid regions. Knowledge of the extent and nature of the vegetation 

cover is needed as a basis for planning treatments of the vegetation and estimating the 

potential water savings and other effects” (Horton et al. 1964: 34).

The Horton et al. 1964 paper was typical of phreatophyte-focused papers and 

reports of the time. During these early days, most riparian research focused on water use 

by individual plants, while other research projects attempted to determine water saving 

through modification of large tracts of land by removing riparian vegetation. Some of 

these habitat modification projects began as early as 1950 when woody riparian plants 

in general (native and nonnative) were targeted for removal (Gatewood et al. 1950) and 

continued well into the 1960s when the focus had mostly shifted to removing the exotic 

or invasive shrub/tree tamarisk (saltcedar; Tamarix spp.) (Chew 2013; Culler et al. 

1982). Anticipated results of these studies prompted governmental and private agencies 

to invest significant financial resources in attempting to reduce or rid drainage ways of 

riparian vegetation. Studies of water use by riparian plant communities, especially nonnative communities, continued from the 1950s and 1960s into the 1980s and 1990s 
2 USDA Forest Service RMRS-GTR-377. 2018

(e.g., Sala et al. 1996; Smith et al. 1998). The practice of removing phreatophytes, 

especially large tracts of tamarisk, continues to today. Indeed, controversy over the 

relative costs/values of tamarisk vs. native riparian species is greatly debated today. For 

example, efforts to control tamarisk by introducing several species of beetle (Diorhabda

spp.) from Asia continue to remind us that threats to riparian ecosystems still persist (see 

below and chapters 4 and 5).

During the late 1960s in Arizona, phreatophyte issues were mostly focused on 

native streamside vegetation. At that time, the Salt River Project and Army Corps of 

Engineers either removed or planned removal of extensive stands of cottonwoods 

(Populus fremontii) and other riparian plants along watercourses, including the Verde 

and Gila Rivers, in Arizona, to “salvage” water and “enhance” river flow. The Verde 

project resulted not only in loss of terrestrial wildlife habitat but also aquatic impacts 

including “siltation of the river gravels and spawning beds” (Tellman et al. 1997). Under 

pressure from avian experts (see below) and in consideration of alteration of spawning 

gravels, the practice, at least within native stands of riparian habitat, was “found to have 

little long-term impact on water supplies downstream, so it was discontinued after more 

than a decade of effort” (Tellman et al. 1997). However, as detailed below, efforts at 

Tamarix removal have persisted into the early 21st century.

U.S. Department of Agriculture scientists were joined by scientists from other 

agencies in developing comprehensive literature on water salvage. From 1958 

(Robinson 1958) into the 2000s, and even as early as 1927 (Meinzer 1927), the U.S. 

Geological Survey (USGS) issued a series of Professional Papers that addressed both 

surface water and groundwater issues (e.g., Culler et al. 1982). In addition, USGS issued 

numerous publications as Water Supply Papers suggesting, as the title implies, that the 

importance of water was primarily viewed in the context of human usage. Earlier in that 

era Horton compiled and abstracted 691 publications on the general subject of phreatophyte control (Horton 1973). In general, the publications Horton cited assigned little to 

no value to the ecological importance of streamside vegetation. As Water Supply project 

results accumulated, however, it became increasingly clear that the projected water savings from most phreatophyte control projects were ephemeral to non-existent (Barz et 

al. 2009; Graf et al. 1984).

In the early 1970s, the scientific community was awakening to the ecological 

uniqueness and high biotic productivity of streamside habitats (see Appendix A for 

early riparian scientists), and it was after several of these ecological studies, primarily 

on birds, that the disproportionate productivity of riparian habitats compared to upland 

habitats was being documented (Carothers et al. 1974; Hubbard 1971, 1977a,b; Hynes 

1975; Johnson 1971; Johnson et al. 1977, 1987; Wauer 1977; Zimmerman 1970). In addition, between the late 1960s and early 1980s, a series of events occurred that ushered 

out the era of phreatophyte control targeted on native riparian species and initiated the 

establishment of the science of riparian ecology and the riparian habitat conservation movement. Academic and management conferences in Arizona, California, and 

Colorado were held documenting the importance of the relationship of the riparian ecosystems to the natural riverine community (see Appendix B for a list of conferences).

One of the early studies, a 5-year riparian breeding bird study, documented the 

highest population densities of nesting birds ever recorded in any habitat in North 

America (Carothers and Johnson 1970; Carothers et al. 1974). The study was initially 
USDA Forest Service RMRS-GTR-377. 2018 3

funded by the Forest Service, U.S. Department of Agriculture and later by the Arizona 

Game and Fish Department in direct response to the request by the Salt River Project in 

Phoenix, Arizona, to the Forest Service to remove the cottonwood and willow habitat 

on public land along the Verde River in central Arizona. One of the more pragmatic 

results of this study was that the ecological damage of indiscriminate phreatophyte 

control was exposed, and the practice, especially with native riparian habitats, began to 

decline significantly. During this same time, the series of riparian conferences continued 

to highlight and strengthen the data on the importance of the riparian corridor and its 

significance to wildlife and general riverine ecology.

The River Continuum Concept and Riparian Habitat Values Recognized

The recognition of both high species diversity and high density of obligate wildlife 

and plant species within riparian ecosystems was coincident with the introduction of 

another important ecological concept, the River Continuum Concept (RCC) (Vannote 

et al. 1980). The RCC emphasized the interrelatedness of riparian and aquatic riverine 

ecosystems. In addition, it recognized and fully valued the synergism of the interdependency of biological and physical interactions in a river system, both longitudinally 

and laterally, which included the stream side community. Before then, riverine studies 

tended to focus on individual elements of aquatic systems, e.g., fisheries or limnology, 

or else components of riparian system, e.g., birds, mammals and plants. Other important 

riparian-related activities during the 1980s included classification of riparian plant 

communities (e. g., Brown 1982; Szaro 1989), a systematic evaluation of the state of 

riparian ecology, and an assessment of needs for further, especially quantified, information about riparian ecosystems (Johnson et al. 1985; Knopf et al. 1988).

About this time, some investigators began to recognize the value of the invasive 

Tamarix as functional wildlife habitat and considered hundreds of miles of streamside 

habitats on some rivers dominated by Tamarix as productive and “naturalized” components of riparian ecosystems (Johnson 1977; Johnson and Carothers 1982, 1987; see 

also Chew 2009, 2013). Studies thereafter and into the 1990s focused on the need for 

maintaining a natural hydrograph for the recruitment of native vegetation and proper 

functioning of the riverine system (e.g., Junk et al. 1989; Poff et al. 1997). Finally, in the 

late 1990s and continuing to the present, widespread concern developed for the protection of existing native riparian habitats as well as restoration of previously degraded 

habitats where possible (Rieger et al. 2014; Zeedyk and Clothier 2012).

Development of riparian ecology and the riparian environmental movement were 

mostly western undertakings with much of the research located in West Coast States of 

California, Oregon, and Washington, and semi-arid States of Utah, Nevada, Arizona, 

New Mexico, and Texas. Other studies were going on in Colorado and the northern 

Rockies as well. Johnson and Lowe (1985) described development of riparian ecology 

in the Southwest, attributing its late appearance to a lack of interest in the transition 

zone from water to upland as well as “the complexity of riparian ecosystems [forming] the ecotone [and] a textbook example of the edge effect.” The term riparian was 

found mostly in “western” publications while terms such as “bottomland hardwoods,” 

“floodplain forests,” and similar terms were used mostly in the East. The U.S. Fish 

and Wildlife Service (USFWS) “wetlands manual,” written by three eastern and one 
4 USDA Forest Service RMRS-GTR-377. 2018

midwestern scientists, did not even use the term “riparian” (Cowardin et al. 1979). A 

search of the Library of Congress catalog by a committee of the National Research 

Council of the National Academy of Science (NRC 2002) using the term “riparian” disclosed an increasing number of entries from 1960 to 1990, mostly from western United 

States. A search using Google Scholar for the first 20 riparian publications also showed, 

for the United States, very few when the terms “riparian” and “eastern U.S.” were used 

together.

Anthropogenic Changes to Western Rivers

By the 1960s, both surface water and groundwater resources in the vicinity of 

large metropolitan and high-intensity agricultural areas had been severely impacted by 

human activity. Instream flow levels in stream reaches above ground, as well as levels 

of water tables below ground, had been greatly reduced throughout much of the region 

(Zektser et al. 2005). Large rivers in which flow levels were measurably reduced or 

modified by large dams, diversions for urban and agricultural use, and other water projects included most lower reaches and tributaries of the Colorado, Gila, Salt, and Verde 

Rivers in Arizona; Rio Grande and Pecos Rivers in New Mexico and western Texas; 

and Sacramento, San Joaquin, and Owens Rivers in California. In most of these river 

systems the natural hydrograph had been irrevocably modified in all but relatively short 

reaches of their headwaters.

In addition to large dams and water diversion projects, many smaller dams along 

major tributaries further reduced normal flows interrupting the natural hydrograph along 

thousands of miles of rivers, drowning riparian habitat in reservoirs above dams and 

desiccating downstream riparian ecosystems through dewatering. Even rivers in lands 

set aside as “natural areas,” such as Big Bend and Grand Canyon National Parks, have 

been impacted by large dams, upstream and downstream of the Park’s boundaries, such 

as Glen Canyon Dam and Hoover Dam on the Colorado River in Arizona (Carothers 

and Brown 1991; Johnson 1991; Johnson and Carothers 1987;) and Elephant Butte, 

Caballo, and Amistad Dams on the Rio Grande in New Mexico and Texas (Chew 2006, 

2013; Wikipedia 2016).

Rivers such as the Santa Cruz, Rillito, and Gila Rivers in southern Arizona were 

perennial streams at time of settlement by Euro-Americans but had been largely reduced 

to intermittent or ephemeral streams by excessive groundwater and irrigation withdrawals by the mid-1950s (Webb and Leake 2006; Webb et al. 2014). Indeed, by the first few 

decades of the 20th century it was the rare western river system where any semblance of 

the natural hydrograph was preserved for any distance below its headwaters. The exception to this is the Yellowstone River, which is the longest undammed, naturally flowing 

river in the West although 2016 plans for a dam on its lower reaches may change this.

Early Ecological Studies

By the early- and mid-1970s, numerous ecological studies, especially on birds, 

were underway along rivers of the Southwest and California (Carothers and Johnson 

1975; Carothers et al. 1974; Gaines 1974, 1980; Gavin and Sowls 1975; Hubbard 
USDA Forest Service RMRS-GTR-377. 2018 5

1971, 1977a,b; Johnson and Simpson 1971; Ohmart and Anderson 1977; Wauer 1977; 

Zimmerman 1970). We found only one much earlier study comparable to these studies 

conducted in 1949 along Dry Creek in the San Joaquin Valley, California (Ingles 1950). 

In addition to data on avian population densities and species richness, most of the early 

studies also contained information on vegetation associated with birds.

The studies conducted throughout the Southwest and California in the 1970s were 

undertaken, in part, due to increasing losses of riparian and related riverine habitat, 

losses that had started with the settlement of the West by Euro-Americans (Carothers 

1977; Dobyns 1981; Knopf et al. 1988). Conservation and loss of these ecosystems were 

not of much general environmental concern until the 1970s and 1980s when several 

riparian conferences addressed these issues. To counter loss or removal of riparian plant 

communities, studies were designed to demonstrate that these communities were, in 

reality, complex ecosystems supporting a broad range of other species. Some of these 

studies were cooperative ventures among several individuals and agencies.

For example, the project to demonstrate the avian impacts of cutting down cottonwood trees along the Verde River was a joint effort of the Forest Service, U.S. 

Department of Agriculture, Arizona Game and Fish Department, U.S. Fish and Wildlife 

Service, Museum of Northern Arizona, and Prescott College. These earliest avian studies were of great scientific interest, demonstrating the critical habitat characteristics 

of the riparian plant community. Simultaneously, studies of riparian forests in the Gila 

River Valley of New Mexico documented the remarkable density and diversity of 

riparian birds (Hubbard 1971, 1977a,b; Zimmerman 1970). Shortly thereafter similar 

information was developed for the birds of the lower Rio Grande (Wauer 1977) and in 

the Lower Colorado River Valley (Anderson and Ohmart 1977; Rosenberg et al. 1991).

To our knowledge, the aforementioned studies provide the earliest quantified documentation of the importance of riparian vegetation as avian habitat in the Southwest. 

This also triggered a demand to better understand the importance of riparian ecosystems 

not only as habitat for birds and other wildlife and invertebrates (e.g., Patton 1977), but 

also as an ecological system that offers other important functions and services, including water quality, flood control, groundwater recharge, nutrient cycling, seed dispersal, 

migratory corridors, and recreation.

Even today, studies focused on other riparian vertebrates, i.e., amphibians, 

reptiles, and mammals, as well as invertebrates, have not received the same amount 

of attention as birds. Birds are generally more visible and more easily studied than 

other vertebrates because they are mostly diurnal, arboreal, and call attention through 

song. To determine factors such as species richness and population densities of the 

other three vertebrate groups, trapping and other special equipment and techniques 

are generally required. We have shown that riparian avian studies were widespread in 

the West by the mid to late 1970s. However, a 1988 publication by the Forest Service 

(Szaro et al. 1988) entitled Management of Amphibians, Reptiles and Small Mammals 

in North America, containing 54 papers, had only one paper on riparian amphibians 

and reptiles and three on riparian mammals. Not surprisingly, only one of these papers 

compares species richness in upland habitat types to riparian types in Arizona, finding 

much higher species diversity of amphibians and reptiles in riparian habitat compared 

to upland habitat (Jones 1988).
6 USDA Forest Service RMRS-GTR-377. 2018

Focus on Riparian Habitats

During the 1970s and continuing through the 1990s, with increasing interest in 

riparian systems, mapping and classification of vegetation in the Southwest became 

the focus of some researchers (Brown 1982; Lowe and Brown 1973). Different riparian communities were recognized based on elevation differences with lower elevation 

communities along river systems ranging from ephemeral to perennial, while higher 

elevation riparian systems where available water was more certain tended to include a 

mix of riverine and upland species (Szaro 1990). Lowe (1961, 1964) illustrated these 

communities in his “The Vertebrates of Arizona”, and more recently, Patten (1998) 

recognized the importance of the interaction of elevation and water availability gradients. Brown et al. (1977) published a drainage map for Arizona and in 1993–1994 the 

Arizona Game and Fish Department developed a dataset that identified and mapped 

riparian vegetation associated with perennial waters in response to the requirements of 

the 1992 Riparian Area Act (this act amended Arizona Revised Statute CARS 45-101; 

Kubly et al. 1997). Maps were created using two major sources of imagery: Landsat 

Thematic Mapper digital satellite data and Multiple Resolution Aerial Videography. 

The data were created to serve as base information for use in GIS systems for a variety 

of planning and analysis purposes. These data also demonstrated how much perennial 

vegetation had been lost along streams and rivers since the early 1900s, with nearly all 

losses associated with dams, groundwater pumping, agricultural clearing, and phreatophyte control.

Other impacts of human land use such as ranching (i.e., livestock grazing) and 

other forms of agriculture, especially when linked to irrigation, also were documented 

as causing major changes and loss of riparian areas (Belsky et al. 1999; Ohmart 1996). 

William Platts was a pioneer in many studies of livestock impacts on riparian systems 

(Platts 1979, 1981). Also, many of his studies dealt with grazing effects on fisheries and 

thus indirectly on riparian systems (Platts 1981). Platts’ articles, although not all necessarily from the Southwest, greatly influenced thinking about impacts of grazing in the 

Southwest as did articles by Wayne Elmore who headed Bureau of Land Management’s 

riparian program for many years (e.g., Elmore 1992; Elmore and Beschta 1987). In 

the late 1960s, the USFS Southwest Region realized that many riparian forests did not 

include multiple age classes but only mature trees. This apparent lack of recruitment 

led to greater control of grazing along rivers and reduction of removal of riparian plant 

communities.

From the 1970s well into the 21st century, riparian research continued to expand 

from descriptive studies of riparian plant and animal communities, to ecophysiological 

studies, to ecosystem studies, with emergence of new journals such as Ecohydrology 

in 2008. Many of these studies addressed functional relationships, such as between 

environmental drivers, e.g., water table and streamflow changes and vegetation response 

(Busch et al. 1992; Fenner et al. 1985; Shafroth et al. 1998, 2000; Stromberg et al. 2005, 

2010), or canopy changes including avian community responses (e.g., Anderson and 

Ohmart 1977; Anderson et al. 1983; Johnson and Haight 1985; Stromberg et al. 1993). 

As riparian communities declined or were altered, ecologists attempted to explain those 

factors that were causing this decline, many being associated with water management 

(Busch and Smith 1995). As more water was removed from rivers of the West, studies 
USDA Forest Service RMRS-GTR-377. 2018 7

were developed to demonstrate the range of instream flow requirements of riparian 

ecosystems (Arthington et al. 2006; Shafroth et al. 2010). These eventually were called 

environmental flows as they were flows that supported all components of the riverine 

system and often mimicked natural flow requirements, flows necessary for a healthy 

riverine system (Poff et al. 1997; Poff and Zimmerman 2010).

Riparian research in semi-arid regions during the 1980s and into the 1990s and 

later was being conducted not only in the Southwest but also along the eastern Sierras of 

California. Unlike the more mesic west slopes of the Sierras, the eastern slopes support 

riparian ecosystems similar to those along rivers in the Southwest. Most of the research 

in this region related to water relations of the riparian system, or components of it, since 

many of the eastern Sierra streams were being controlled and/or diverted to move water 

into the Los Angeles aqueduct or for hydropower production. Examples of this research 

included studies by Stromberg and Patten (1990, 1992) and Smith et al. (1991). Also, 

in the late 1970s and 1980s, a group of students from California studied the flows into 

and dropping levels of Mono Lake, a closed basin lake in the eastern Sierra. This group, 

led by David Gaines (1974), an environmental activist and ornithologist, helped form 

the Mono Lake Committee which today works toward protection of the lake. In 1987 

a committee of the National Research Council/National Academy of Sciences studied 

dropping lake levels of Mono Lake, its causes, reduced stream inflows, and consequences (NRC 1987).

Hydrology, Instream Flows, and the Natural Hydrograph

Further concern for appropriate flows for riparian and riverine processes was based 

on, and stimulated research on, the hydrological processes that triggered responses of 

the riparian ecosystem. Flooding was shown to be an essential disturbance process for 

the long-term maintenance of riparian systems because most characteristic riparian 

woody species were successional species recruited during and after flood (Brady et al. 

1985; Stromberg 2001). For example, the importance of uncontrolled flood flows and 

their timing and magnitude was found to be critical to establishment and maintenance 

of riparian systems (Brown et al. 1977). The “recruitment box” model developed by 

Mahoney and Rood (1998) documented how these processes interacted to enhance 

seedling establishment of riparian plant species. Other disturbance processes that were 

found to be important to riparian dynamics included fire that results in a change in population size structure with more mature cottonwoods, smaller willows, and fewer Tamarix

(Stromberg and Rychener 2010).

A significant change in the long-term pattern of annual flows can substantially 

reduce the vigor and productivity of riparian ecosystems or the basic characteristics of 

a stream’s natural hydrograph (Annear et al. 2004; Merritt and Poff 2010; Poff et al. 

1997). Those elements of the “natural hydrograph” currently recognized as necessary 

to protect and nurture aquatic and terrestrial ecosystems consist of seasonally variable 

flows that include relatively stable flows as well as flood flows from winter storm events 

and summer monsoon rains (Annear et al. 2004). The annual pattern, and the multi-year 

variability, in these flows set the stage for the aquatic and terrestrial biological resources 

that can survive and evolve in the system. Each stream, unless it is dam regulated or 

otherwise significantly altered by anthropogenic influences, or natural causes such as 
8 USDA Forest Service RMRS-GTR-377. 2018

earthquakes, has a pattern of flow that is generally dictated by climate, topography, 

geology, depth to groundwater, and other factors (Annear et al. 2004).

A fundamental understanding of what is needed to maintain aquatic and terrestrial 

riverine ecosystems includes four elements of the important instream flow principle of 

connectivity: longitudinal flows (flows that extend over the length of a stream course); 

lateral flows (flows that cover the width of the channel and floodplain); vertical flows 

(flows that provide sufficient stream depth and connect to the hyporheic corridor); 

and flow duration (the amount of time a flow is at a particular level) (see Annear et al. 

2004). Moderate to high floods reshape the riparian habitats by scouring and depositing 

sediments onto substrates above the active stream channel and redistributing organic 

material that can benefit healthy riparian vegetation (Annear et al. 2004; Bunn and 

Arthington 2002; Leenhouts et al. 2006; Poff et al. 1997; Shafroth et al. 2000; Smith 

et al. 1991). These same flows also can benefit aquatic habitats by redistributing gravel 

and fine substrates that are important for fish spawning and rearing. Extreme flooding 

events can uproot riparian vegetation and scour instream aquatic food upon which fish 

species rely (Meffe and Minckley 1987). Large floods preferentially impair nonnative 

fishes over flood-adapted natives (Gido et al. 2013; Propst et al. 2008). This “re-setting” 

of the fish community can give the native fishes at least a temporary window in time to 

rebound from the negative (competitive and predatory) interactions with nonnatives.

During the later decades of the 1900s, several river systems were studied as 

integrated ecosystems along elevation and river volume or flow gradients. Some were 

small, such as Aravaipa Creek (e.g., Barber and Minckley 1966; Lowe 1964; Minckley 

1981), some intermediate in size, such as San Pedro River (e.g., Stromberg and Tellman 

2009), and some large, such as the Colorado River (e.g., Carothers and Brown 1991; 

Johnson and Carothers 1987) and studies by the Bureau of Reclamation’s Glen Canyon 

Environmental Studies and other Federal and State agencies in the 1980s and early 

1990s (NRC 1991).

Rise and Fall of Tamarix

One of the more interesting issues in telling the history of our understanding 

aridland riparian ecosystem concerns the role of introduced species, especially Tamarix

species. Although Russian olive (Elaeagnus angustifolia) has also been introduced and 

spread throughout western North America, it does not occupy as dominant a role as 

tamarisk in the riparian ecosystems of the arid Southwest where understanding the role 

of tamarisk and methods for its eradication dominate discussions about nonnative riparian species (Nagler et al. 2011).

Since Tamarix was introduced to the United States from Asia in the early 1800s, 

the complex of several species that we know today as saltcedar or tamarisk have 

garnered a significant amount of attention from erosion control specialists, horticulturalists, wildlife ecologists, and, of course, water managers whose responsibilities include 

providing water for the human environment. As discussed earlier, it was those water 

managers and the rapidly proliferating complex of Tamarix species that led to phreatophyte control in the mid-1900s. In the past several decades it has become apparent 

that Tamarix as wildlife habitat, while not as productive as native riparian woodlands, 

can have significant value (Bateman et al. 2013; Davis et al. 2011; Hunter et al. 1988; 
USDA Forest Service RMRS-GTR-377. 2018 9

Johnson and Carothers 1987; Paxton et al. 2011; Stromberg et al. 2009). High wildlife 

use and the “naturalization” of Tamarix is well documented for hundreds of miles of 

river downstream of dams where either pre-dam conditions precluded the establishment 

of extensive riparian woodlands, or alternatively where regulated floodplains were no 

longer suitable for native riparian species (Anderson et al. 1977; Carothers and Brown 

1991; Carothers and Dolan 1982; Johnson and Carothers 1987).

Now, with the introduction of biological control of Tamarix and its apparent success in defoliating the plant throughout most of its range in western riparian habitats, 

another important riparian resource is well on its way to being lost (see Chapter 5). 

Interestingly, when the U.S. Department of Agriculture biologists of Animal and Plant 

Health Inspection Service (APHIS) developed their effective Diorhabda spp. beetle release program, they made several fundamental errors in judgment including inadequate 

experimental evidence concerning the effect of the beetles (APHIS 2005).

For example, while acknowledging that the post-beetle revegetation with native 

species would not be possible in some areas, they generally approached this biological 

control project with the expectation that removal of Tamarix alone was the goal. APHIS 

generally assumed that once Tamarix was disadvantaged, native species, especially 

cottonwoods and willows, would return to riparian systems. It is now apparent that 

without “strategic restoration approaches” the native riparian community is unlikely to 

recover (Shafroth et al. 2013). Observations of the Colorado River above, within, and 

below Grand Canyon include the unfortunate reality that once the beetles have done 

their work, nonnative perennials and annual non-woody weeds and shrubs sprout and 

spread under the dying Tamarix. Unfortunately, recolonization of native riparian species 

in the shadow of the declining Tamarix has proven to be unlikely without significant restoration of soils and other biotic and abiotic environmental elements (see references in 

Chapter 5; Paxton et al. 2011; Shafroth et al. 2008).

APHIS also significantly underestimated the rate at which the beetles would 

spread. They predicted decades for the beetle to eventually occupy much of the western 

range of Tamarix (DeLoach et al. 2000); however, the reality is that within less than 

a decade the beetles swept like a plague through the western range of the species and 

occupied and defoliated a large part of the Tamarix range (see Tamarisk Coalition 

information online, http://www.tamariskcoalition.org). The beetles have neither invaded 

central or eastern Arizona below the Mogollon Rim nor occupied the Lower Gila and 

Salt Rivers or the Lower Colorado River below Bill Williams River, but it is only a 

matter of time before it will be difficult to find stands of Tamarix that have not been 

infested with the beetles. Perhaps the greatest oversight made by APHIS was their apparent assumption that Tamarix was of little wildlife value. They ignored a large body 

of knowledge regarding the value of Tamarix to wildlife, including studies documenting 

the significant use and carrying capacity for birds, mammals, reptiles, amphibians, and 

invertebrates (see references in Sher and Quigley 2013).

Attitudes toward Tamarix have changed since it was first introduced. Chew 

(2013) has recognized three phases of scientific and management attitudes toward the 

species. Clearly, we are now entering the fourth stage of the focus on Tamarix as a 

significant riparian species (Stromberg et al. 2009). In this fourth phase we are entering the unknown. What replaces Tamarix and the rate at which recolonization occurs 

are not predictable at this time. In Chew’s first phase from introduction to the early 
10 USDA Forest Service RMRS-GTR-377. 2018

1900s, the species were celebrated and widely propagated for their role in erosion 

control, wind breaks, ornamental landscaping, and shade, especially in dryland situations where available water was limited and the species could thrive. By the second 

phase in the mid-1900s, when Tamarix was beginning to be reviled for its heavy water 

consumption and rapid proliferation throughout aridland watersheds, the word phreatophyte or “well plant” was first used (Meinzer 1923). This is when the phreatophyte 

control concept was born.

Unfortunately, while the focus was on Tamarix, other native riparian woodland 

species were also targeted (Carothers et al. 1974). It was firmly believed during the mid1900s to the late 1960s that if Tamarix and, secondarily, native riparian species were 

eliminated or controlled, the water used in evapotranspiration by these “water wasters” 

would be more available for beneficial uses (see above and Gatewood et al. 1950). It 

was during Chew’s second phase when limited local and regional attempts were made, 

including mechanical removal, burning, poisoning and drowning, to remove Tamarix, 

albeit ineffectively. This focus on Tamarix and other riparian woodland species occurred 

decades before the disproportionately high value to wildlife of riparian habitats compared to upland habitats was fully recognized (see Carothers 1977).

Additionally, later studies demonstrated that projected savings in “water salvage” 

had been greatly exaggerated and, in some cases, seemed to be non-existent. This was 

the case for a USGS Tamarix removal study area of about 5,500 acres (2,225 ha) upstream of the Coolidge Dam and its reservoir, San Carlos Reservoir, on the Gila River 

in Arizona (Culler et al. 1982, Graf et al. 1984). Not only did removal of Tamarix not 

improve water savings, Tamarix rapidly recovered and the floodplain is now densely 

covered with the plant.

It was not until Chew’s third phase (1960s to early 2000) that “Tamarix suppression” began in earnest (Chew 2013). Notwithstanding organized attempts at eradication 

in the 1960s and 1970s, little success on Tamarix control was ever realized. We know 

now that part of the reason for the success of the species is the fact that Tamarix easily 

proliferates, dispersing seed over many months. Compared to seed dispersal patterns of 

native species that disperse seed aligned with winter/spring high-water, Tamarix are especially successful and replaces the native species in areas where the natural hydrograph 

has been modified by dams, irrigation withdrawal, flood control structures, declining 

water tables, livestock overgrazing and other anthropogenic watershed and stream 

alterations (Culler et al. 1982; Johnson 2013; Merritt and Poff 2010; Stromberg et al. 

2007; Turner 1974).

Indeed, along river systems within the Southwest where the natural hydrograph 

has not been significantly modified, Tamarix is often relatively rare and unable to gain 

a foothold. Examples can be found along the upper portions of the Verde River and 

tributaries (Beauchamp and Stromberg 2007) and the San Pedro River and tributaries 

(Stromberg and Tellman 2009) in Arizona, and portions of the Upper Gila and tributaries (Hubbard 1971, 1977b) in New Mexico where the native riparian community 

remains largely intact.

For sections of those rivers where anthropogenic perturbations to the stream occur, Tamarix has been found to dominate. Merritt and Poff (2010) have convincingly 

demonstrated that the decline of western riparian systems can be primarily traced to 

flow alterations and that, with or without the introduction of Tamarix, the native riparian 
USDA Forest Service RMRS-GTR-377. 2018 11

gallery forests would have declined. However, Tamarix is sufficiently adaptive such that 

it may establish and sometimes proliferate even in areas where the natural hydrograph 

has not been modified (Merritt and Poff 2010; see also Johnson 2013).

Following Chew’s chronological sequence, we now find ourselves in the 

fourth and what may prove to be the final phase of the rise and fall of Tamarix in the 

Southwest. This fourth phase began in earnest in about 2000 with the introduction of 

biological control. The release of exotic Asian beetles has to be considered the most 

aggressive and finally successful attempt at removing Tamarix. As documented in 

McLeod’s excellent account of the history of the beetle release program (Chapter 5), the 

biological control of the Tamarix was far more effective than its designers could have 

ever hoped or imagined. The Tamarix-defoliating beetle is now invading the nesting 

areas of the endangered southwestern willow flycatcher (and 90-plus other avian species 

and countless other vertebrates and invertebrates) in Arizona, California, New Mexico, 

Nevada, Texas and Utah http://www.tamariskcoalition.org/events/tamarisk-beetlemaps). If the beetle spreads farther without mitigation, the flycatcher’s survival is likely 

to be threatened—and perhaps that of several other species—and it may significantly 

change the wildlife carrying capacity of hundreds of thousands of acres of naturalized 

second-growth riparian habitat comprised of non-native tamarisk.

APHIS indicated that mitigation could be needed if the release of beetles resulted in defoliation of occupied habitat of the endangered flycatcher (APHIS 2005). 

Mitigation could include planting native willows and cottonwoods or appropriate 

riparian species to replace dying Tamarix—to help endangered birds. The Center for 

Biological Diversity and the Maricopa Audubon Society are currently suing APHIS in 

U.S. District Court for Arizona [Case 4:09-cv-00172-FRZ]. The Plaintiffs allege that 

APHIS violated its affirmative obligations under the Endangered Species Act section 

7(a)(1) by failing to adopt a program to offset the harm caused by the beetle release, 

and by taking no action to mitigate the adverse effects of the beetle prior to and after 

suspending the program. (Note: As of June 2010 APHIS’s beetle release program was 

formally suspended in 13 States: Colorado, Idaho, Iowa, Kansas, Missouri, Nebraska, 

Nevada, North Dakota, Oregon, South Dakota, Montana, Washington, and Wyoming. 

The Lower Colorado River States were not included because the southwestern willow 

flycatcher continues to use Tamarix for nesting in those States.)

APHIS is taking the position that the Plaintiffs are mistaken about the scope 

of an agency’s duties under ESA section 7 by improperly assuming the agency has a 

duty to offset all adverse effects of their action. APHIS also argues that it did the best 

it could and took a variety of affirmative conservation actions to benefit numerous 

listed species, including the flycatcher, and that the mitigation the Plaintiffs seek is 

beyond APHIS’ authority. While the legal case1 is far from being resolved, one thing 

is certain: The introduced riparian species complex Tamarix is predicted to soon be a 

shadow of its former value to western riparian wildlife. With some luck, once Tamarix

is further reduced (nobody really expects complete eradication), the beetles’ appetite 

will not switch to native riparian species. Additionally, if appropriate hydrological 

conditions can be maintained in areas with Tamarix removal, cottonwood and willow 

can compete in the early growth stages and may prevent further invasion and enhance 

natural recovery (Stromberg 1997).
12 USDA Forest Service RMRS-GTR-377. 2018

Restoration of Riparian Habitat: Putting the Genie Back in the Bottle

During the developing period of riparian studies, the young science of restoration 

ecology, including efforts to restore denuded riparian habitat, was also gaining significant traction and attention (e.g., Goodwin et al. 1997; and Stanley et al., in prep.). The 

rapid expansion of this interest in riparian habitat restoration is documented by Koehler 

and Thomas’ (2000) annotated bibliography Managing for Enhancement of Riparian 

and Wetland Areas of the Western United States, which lists approximately 250 documents focused primarily on riparian restoration along living streams. In a 2007 study 

of riparian and river restoration projects in the Southwest, approximately 33 percent 

were riparian management while others included, for example, flow modification and 

instream habitat recovery (Shah et al. 2007). Riparian management included intensive 

management of livestock allotments to protect or improve the conditions of riparian 

areas.

In the past several years, riparian habitat studies have expanded beyond hydroriparian ecosystems (associated with perennial streams) and mesoriparian ecosystems 

(associated with intermittent streams) (Johnson et al. 1984). Now workers ranging 

from ecologists to city planners (Krausman et al. 1985; Steiner et al. 1999) conduct 

investigations on conservation, management, and restoration of xeroriparian ecosystems 

associated with the long-overlooked but spatially abundant and wildlife-rich ephemeral 

river systems along washes and arroyos (Cooke and Reeves 1976; Levick et al. 2008).

Most published studies available today indicate that while some restoration successes are impressive, complete restoration of riverine systems is often an elusive goal 

once they are damaged. This includes both the aquatic and terrestrial components. 

However, in some situations, riparian ecosystem resilience is remarkable in its ability to recover from direct destructive impacts like flooding, wildfire, and livestock 

overgrazing. In fact, ample evidence exists that once floods and fires have passed, it is 

not unusual for natural processes to return the damaged areas to their original vitality. 

For example, a massive 850 cms (30,000 cfs) flood in Arizona’s Aravaipa Canyon in 

October 2008 scoured over 25 miles of the canyon bottom and destroyed approximately 

half of the cottonwood-willow gallery forest; however, within a decade the area was 

largely recovered to pre-flood conditions (Burtell 2013). Similarly, a human-caused 

wildfire in Grand Canyon’s Deer Creek in 1972 completely denuded a mile of floodplain of mature cottonwoods; yet again, within a decade or so, the area was naturally 

revegetated to a closed gallery forest (Carothers and Brown 1991). The most remarkable 

example of the impact of overgrazing on riparian habitat and the resilience of the riparian system to recover is the Bureau of Land Management’s San Pedro Riparian National 

Conservation Area in southern Arizona. Once livestock grazing was prohibited within 

the conservation area (approximately 50 miles), cottonwood-willow gallery forest 

tripled in size in a 30-year period (NRST 2012; Stromberg and Tellman 2009).

While many and varied techniques are available to reestablish riparian vegetation 

once it is damaged or removed, the constant onslaught of human activities that change 

fluvial processes and natural channel ecological functions, combined with the long history of exotic species introductions and our general inability to counteract these aliens, 

provides a constant frustration to river restoration attempts. Moreover, as Stanley et 

al. (in prep.) have documented for the State of California, most restoration projects 
USDA Forest Service RMRS-GTR-377. 2018 13

are extremely small and of 276 separate restoration projects analyzed in the 1990s, 83 

percent were less than 20 acres in size. Shah et al. (2007), when analyzing data from 

576 riparian restoration projects in the Southwest, indicated that few were over 3 km in 

length thus allowing for a better chance of success.

What is important to emphasize in the above flooding, fire, and grazing examples 

of natural restoration of the riparian habitat is that in all three areas, the natural hydrograph and the ground-water recharge system were largely intact. In areas where the 

hydrologic regime can be restored to some semblance of the “natural hydrograph” and 

the deviation from normal surface and subsurface flow and storage cycle conditions 

can be understood, large-scale restoration is possible. An excellent example of riparian 

restoration success is the effort of The Nature Conservancy (TNC) in the Lower San 

Pedro River Basin to acquire and protect large tracts of land adjacent to riparian areas. 

Once the land is acquired and it is possible to remove ecosystem stresses, such as excessive groundwater pumping, removal of diversion structures, restoration of base flows, 

control of grazing, and off-road vehicle use, riparian habitat eventually comes back as 

watersheds restore in the uplands and groundwater recharge returns to some semblance 

of pre-depletion levels (Haney 2005, Stromberg and Tellman 2009).

However, there are locations where riparian restoration is almost impossible on a 

large scale and will rarely occur without the intervention of intense management. These 

include situations where impoundments, diversions, depletion of surface and groundwater supplies, and proliferation of exotic species are the reasons for native riparian habitat 

declines. While we can celebrate the recent efforts of TNC in the Lower San Pedro 

River and elsewhere throughout the arid Southwest, we have clearly reached a new level 

of crisis with the release of the tamarisk beetle and the destruction of thousands of acres 

of “second class” riparian habitat. This has clearly reduced wildlife-supporting riparian 

habitat throughout the arid Southwest.

The ravages of the tamarisk beetle have set this crisis stage and there does not 

seem to be a reasonable solution in the near future. Major dams and impoundments 

altered the natural hydrograph on most large streams and rivers and ushered in the first 

profoundly destructive phase of riparian habitat demise. Many of those impoundments, 

while clearly controlling most flood flows, precluded conditions necessary for the 

recruitment of cottonwoods and willows (Mahoney and Rood 1998) and allowed for 

the proliferation of the nonnative Tamarix. As detailed in earlier sections of this paper, 

Tamarix was largely reviled by water managers and conservationists alike. Nevertheless, 

the nonnative woody plant did proliferate over thousands of acres of hydrologically 

altered stream channels. Though a nonnative, Tamarix was better than no woody vegetation along the native-plant-denuded stream channels, and it did form suitable habitat for 

hundreds of native wildlife species as well as provide stream cover and nutrient input in 

the form of allochthonous organic material. Now, as the tamarisk beetle has proliferated 

virtually unchecked throughout the arid Southwest, there are hundreds of stream channels where dead and dying Tamarix are no longer even second-class riparian habitat. 

Areas under the dying Tamarix are mostly revegetating with non-woody shrubs and 

herbaceous species, many of which are nonnatives (Chapter 5).

It is important to point out that the Tamarisk Coalition, founded in 1999 in advance 

of the onslaught of the tamarisk beetle, is a non-profit dedicated to the advancement 

of the restoration of riparian lands through collaboration, education, and technical 
14 USDA Forest Service RMRS-GTR-377. 2018

assistance (http://tamariskcoalition.org/about-us/accomlishments). They are actively 

pursuing restoration of habitat where the Tamarix has been defoliated, but their efforts 

to date are relatively small and include removing dead and dying vegetation and reestablishment of native grasses combined with planting some woody cottonwood and willow 

species. Large-scale efforts to restore more than a few acres of habitat are extremely 

costly and non-existent at this time.

Additionally, restoring non-salt-tolerant trees such as willows and cottonwood 

is initially difficult because Tamarix alters soil chemistry enough to reduce restoration 

success (Siemion and Stevens 2015). On the Rio Grande in the vicinity of Elephant 

Butte Reservoir, the Bureau of Reclamation has been initiating riparian restoration 

efforts as mitigation for listed species, but like most attempts to restore native species 

where Tamarix has dominated for years, the restoration areas are quite small (<20 acres) 

(BOR 2016). Reintroduction of beaver to restore local hydrology and stimulate recovery 

of a woody riparian community has been used along some smaller rivers in the semiarid West with some success (Carrillo et al. 2009; Castro et al. 2015; Welch 1997). 

Unfortunately, beavers’ preference for cottonwood and willow over the nonnative 

Tamarix slows recovery of a native riparian forest (see chapters 6 and 7 [beavers] and 4, 

and restoration chapters in Johnson et al., in prep.).

Climate change also poses significant problems for riparian habitat restoration. The 

conventional wisdom for the trajectory of climate change seems to tend toward understanding that the arid Southwest will become warmer and drier, thus piling on additional 

threats to existing riparian habitat. Some researchers are now advocating that the use 

of “local stocks” of riparian vegetation be replaced with more southerly stocks where 

the plants are already adapted to warmer and drier conditions. The belief is that climate 

change is happening so quickly that the local seed and cuttings that are currently sources 

for revegetation are maladapted.

Thus, if the objective is to revegetate cottonwood and willow stands along the 

Verde River in central Arizona, seed sources and cuttings should come from farther to 

the south, perhaps from the San Pedro River drainage (Carroll et al. 2014; T. Whitham, 

Northern Arizona University, personal communication; Whitham et al. 2006). 

Additionally, a warmer climate will change winter and spring runoffs from higher elevations where snow collects. This is expected to affect the phenology of riparian species 

and thus results of restoration activities. Genotypic elasticity of woody riparian species 

may be sufficient to overcome this, but use of more adapted genotypes may be required 

for successful restoration.

Summary

A review of the history of riparian ecological science shows an evolution from 

attempting to solve anthropogenic issues relating to water supply to guiding broad 

programs of riverine management. In the early years of riparian ecology, river systems 

were primarily considered sources of water for urban systems, agriculture, and industry. 

Consequently, the more water made available for these uses, that is, the social ecological system, the more the economy and human welfare would be advanced. To improve 

delivery of water by riverine systems, science was expected to reduce competing 

uses of water along the delivery system. This included lining canals and removing, or 
USDA Forest Service RMRS-GTR-377. 2018 15

challenging, the use of water by vegetation growing along the riverine floodplain, that 

is, riparian vegetation.

Scientists began to evaluate the amount of water used by riparian vegetation and 

suggested that removing this vegetation could salvage enough water to balance any negative aspects of removal. Consequently, “experiments” were untaken to remove riparian 

trees and shrubs from the edge of the rivers, or total removal of that plant community 

where it had overrun the extended floodplain. Success in improving water yield by these 

procedures was never proven as significant.

Concurrent studies in the 1970s and 1980s shed light on the ecological values of 

the riparian vegetation, not just as an aggregation of plants but as a complex ecosystem 

important to riverine functions and one that formed important habitat for many species. 

Other studies documented the ecological values and disproportionate diversity and density of some vertebrates using these systems. Most early studies along these lines were 

avian studies. Birds were found to use many strata of the riparian vegetation and thus 

these habitats were critical in survival and maintenance of most of the avian species of 

the Southwest lowlands. This information, along with evidence of little or no water savings with riparian vegetation removal, essentially put a stop to programs that removed 

or highly modified riparian vegetation.

Unfortunately, discovery of the importance of riparian vegetation followed on a 

lengthy period of dam building in the West. Some dams were constructed many years 

earlier but, again to improve water delivery to urban/industrial/agricultural centers, 

water needed to be “stored” and these social ecological systems also needed electrical 

power that the dams could supply. Dams were managed to control downstream flows 

along with inundating upstream riverine systems. Many important riparian systems went 

under water and the negative consequences of a highly modified downstream river flow 

was demonstrated by riparian research. The importance of the whole riverine system 

took on a new meaning with the development of the River Continuum Concept, and 

importance of natural flows (including the concept of environmental flows) was emphasized and supported through extensive riparian/riverine research.

Into the 1980s, more riparian research was initiated to address the response of 

different components of riparian ecosystems to alterations of the many drivers that influenced these systems. Changes in flows, alteration of groundwater levels, introduction 

of exotic species, and diversion of channel systems are all examples of modifications 

studied. Exotic species became of great importance with the expansion of Tamarix

throughout the West. Tamarix impacted many functions of both modified and unmodified riparian ecosystems. Consequently, removing or destroying some riparian systems, 

those dominated by Tamarix, was back in vogue. The rise and fall of the Tamarix spp. 

in the western United States is a classic story of the unintended consequences of not just 

the introduction of an exotic species, but of ecosystem destruction. Tamarix changed the 

natural hydrograph of streams and rivers, followed later by largely unadvisable attempts 

to fight it, but it was found to be beneficial as wildlife habitat, with yet another introduced species, the tamarisk leaf beetle.

Today, basic research on riparian functions is still ongoing but more emphasis is 

being placed on understanding and restoring riparian systems that have been altered. 

This effort will reach new heights in the shadow of the Tamarix beetle’s impacts as hundreds of thousands of acres of second-growth riparian habitats are becoming available 
16 USDA Forest Service RMRS-GTR-377. 2018

for restoration efforts. Although costs for this restoration will be substantial, benefits to 

wildlife and stream ecology should more than adequately justify the expense and effort.

Riparian research in the West will continue to develop more refined data on 

function and how this information can guide management. These efforts must include 

continuing dialog among riparian specialists, users, and the public if we expect riparian 

systems to continue to be part of the western mosaic of ecosystems and habitat for those 

species for which this system is critical. The future of riparian ecology is not only better 

understanding of functional components of these important systems but better ways in 

which scientists, managers, decisionmakers, and the public can work together to use all 

forms of data to create a successful future for these critical western riverine systems.

Note 

This case was legally resolved in June 2018 (see McLeod, volume 1) as this publication was in press.

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USDA Forest Service RMRS-GTR-377. 2018 25

Chapter 2. Development of Riparian Perspectives in the 

Wet Pacific Northwest Since the 1970s

Frederick J. Swanson and Stanley V. Gregory

Introduction

Streams and riparian zones have been fertile ground for ecosystem science and 

a battleground for forest policy and management in the wet Pacific Northwest west of 

the crest of the Cascade Range for many decades. Competing, high-value resources 

of salmon and big Douglas-fir timber and their iconic places in cultures of the region 

sharpened the clash of values. Landslides from forestry operations and roads and elevated water temperature in streams where forest cover had been removed were points 

of physical connection between steep slope forestry and cold-water fishes. Logging 

slash from harvest operations had dammed streams and depleted dissolved oxygen, 

leading fisheries agencies and advocates to call for removal of wood from streams in 

the 1950s and 1960s.

In the decades since, science has played important roles in characterizing ecosystem components and dynamics and in identifying issues and management options. 

Social conflicts have propelled the science forward. In this essay, we offer a brief 

historical overview of steps in the development of concepts about riparian zones in this 

region and societal context from the perspective of the large, interdisciplinary science 

team—the Stream Team—based on the Oregon State University campus in Corvallis 

and at the H.J. Andrews Experimental Forest in the Willamette National Forest east of 

Eugene. Team members come from the University, Forest Service research and land 

management branches, and other institutions; and the participants have roots in stream 

ecology, fisheries and forest science, geomorphology, and other fields. The nucleus 

of the Stream Team has been large research programs—the International Biological 

Program in the 1970s and the Long-Term Ecological Research program since 1980, 

both supported by the National Science Foundation and the Forest Service, and based in 

Oregon State University. Important work occurred elsewhere in the region, most notably 

based in Seattle in fisheries and forestry research, and outreach programs based at the 

University of Washington (e.g., Naiman et al. 2005), but we do not attempt to cover that 

work in this chapter.

An apparent contrast in the perceptions of riparian zones between the dry interior 

west and the wet Pacific Northwest west of the crest of the Cascades may stem in part 

from the contrasts in the stature of vegetation. Riparian zones in many arid lands can be 

conspicuous as lush stands of shrubs and scattered trees in a sea of knee-high sagebrush. 

Westside conifer forests, on the other hand, can be 70+ m tall, dwarfing streamside 

willow (Salix spp.) and red alder (Alnus rubra) stands and creating continuous forest 

canopy from stream banks to ridge. Westside forest ecologists naturally focused their 

science on the interactions among forests, streams, and riparian zones. In the case of the 
26 USDA Forest Service RMRS-GTR-377. 2018

Andrews Forest team, stream ecologists pressured the forest scientists to extend their 

work down into the riparian zone.

The evolution of thinking about riparian zones in the Andrews Forest team was 

the gradual awakening to the many interactions between forests and streams. Driven 

by both science and policy questions, big wood in streams became a pivotal issue in 

the mid-1970s. The policy question: Should loggers be required to remove wood from 

streams to provide for fish passage and limit biological oxygen demand? The science 

question: What does big wood contribute to the geomorphic structure, organic matter 

and nutrient budgets, and overall functioning of stream ecosystems? The history of logging slash and stream management often is characterized as, “First they told us to take 

wood out of streams and now they are telling us to put it back. When will those darn 

scientists change their minds again and tell us to take it out again?”

But, the management history was more nuanced; early rules called for leaving the 

pre-existing wood in streams and removing only logging debris. The pendulum swung 

from having too much wood in streams—especially readily mobilized logging slash—to 

removing too much wood from streams, especially wood of a natural size distribution, including big, stable pieces. The arguments about how much wood is appropriate 

continues, but the general thread of the story is that wood is a natural part of stream 

systems so policy now sustains that function through direct intervention in streams and 

management of riparian forests for future wood supply. The big wood connection between forest and stream turned out to be vital to the interdisciplinary spirit of the science 

team—the work required integrating the perspectives of stream and forest ecologists and 

geomorphologists.

By the latter half of the 1970s, the Andrews group had made substantial progress 

on their studies of nitrogen and organic matter budgeting (Triska et al. 1984), addressing both processes and standing stocks of materials all with attention to how forests 

affect streams. A succession of papers (in chronological order: Cummins 1975; Meehan 

et al. 1977; Swanson et al. 1982; Gregory et al. 1991) developed the thinking about 

forest-stream interactions in terms of regulation of light levels influencing primary 

productivity, water temperature, fish foraging efficiency, and other processes. Also summarized in figure 1, the roles of forests in supplying organic matter range from fine litter 

to whole old-growth trees that shape stream channels, provide cover, and provide substrates for biological activity. This evolution led to the notion of defining riparian zones 

in ecosystem terms as the zone of interaction rather than on the basis of hydrologic, botanical, or soil considerations (Gregory et al. 1991). Unlike lowland fluvial systems with 

well-defined floodplains, hydrologic criteria do not work well in steep mountain streams 

with their high levels of topographic complexity imposed by boulders, big wood, and 

narrow valley floors.

Botanical criteria have limitations in part because the great stature of vegetation 

means that trees distant from the stream can have important influences. Studies of riparian vegetation have been rather limited in these mountain environments. In the Andrews 

Forest, for example, in 1979 and 1990 forest ecologists established several large 

(2–2.4 ha) stem maps straddling streams of different size, but it took some years and a 

major flood to develop a record of sufficient length to reveal the disturbance dynamics 

of the riparian system (Acker et al. 2003). Hydric soils criteria for defining riparian 
USDA Forest Service RMRS-GTR-377. 2018 27

systems in the forest setting are far too narrow and fail to encompass the terrestrialaquatic interactions that strongly influence stream ecosystems.

Also, during the late 1970s, the highly influential River Continuum Concept 

(RCC) project was in full swing (Vannote et al. 1980, which had been cited more than 

8,300 times as reported in Google Scholar as of 22 March 2017). A national program 

led by Robin Vannote of the Stroud Water Research Laboratory in southeastern 

Pennsylvania explored concepts of how forest influences on stream ecosystems varied 

from small headwater streams to large rivers. As the stream widened downstream, more 

light reaches the channel, so food resources for aquatic organisms shift from productivity driven dominantly by forest litter to in-stream primary production. A cascade of 

ecological consequences follows, including shifts in composition of the aquatic invertebrate community from one that processes organic inputs from the surrounding forest 

to grazers that process algae and diatoms produced in the channel itself. Many other 

aspects of the stream ecosystem, including big wood, were examined at four stream 

sizes from first- to seventh-order channels. The Andrews Forest was one of four study 

areas in the RCC project scattered across the country. This work helped place stream 

ecosystems in a landscape context by explaining variation along the longitudinal profile 

of the stream system. The RCC also argued for connectivity through the stream network 

via the influence of upstream areas as sources of nutritional resources for downstream 

rivers and floodplains.

A key feature of the Andrews Forest program has been its close partnership with 

land managers of the Willamette National Forest. Preparation of the forest management plan culminating in 1990 included a prime example of that partnering in the form 

of a 65-page supplement for stream and riparian management guide authored by two 

researchers working in collaboration with National Forest personnel (Gregory and 

Ashkenas 1990). This guide affirms the many important ecological functions of riparian 

zones, the policy direction to sustain them, and the necessary management standards 

and guidelines in the context of a full watershed perspective. The collaborative approach 

has been mutually beneficial; researchers bring the most up-to-date science and the land 

managers bring a great deal of real-world experience plus exposure to the competing 

Figure 1—Foreststream interactions 

as a basis for 

defining the 

riparian zone, as 

viewed in 1978 

(source: Meehan 

et al. 1977, public 

domain).
28 USDA Forest Service RMRS-GTR-377. 2018

values within society. Therefore, the ultimate plan has the best chance to be credible on 

science, societal, and operational fronts.

This sense for integrating science and land management practices was valuable a 

few years later during the Forest Ecosystem Management Assessment Team (FEMAT) 

process convened by President Clinton in April 1993, to move beyond the injunction 

that Judge Dwyer had leveled on logging on Federal lands in the range of the northern 

spotted owl—10 million hectares along the Pacific Coast (FEMAT 1993). The FEMAT 

process set the path to the Northwest Forest Plan (NWFP) signed by the secretaries of 

Agriculture and Interior in 1994. A synthesis of some of the findings from Andrews 

Forest science and other sources was encapsulated in a figure that helped shape thinking 

about width of riparian reserves to maintain many functions of streamside forests within 

cutting units (fig. 2).

This conceptual framework was a dramatic departure from the policy debates 

about riparian buffer widths over the previous 20 years. One of the first questions that 

emerges in riparian zone management is: How wide should buffers be? Most discussions focused on uniform distances from the streams edge based on the operational 

willingness to forego some or all timber harvest within that distance. The Riparian 

Reserves of the NWFP designed riparian widths based on site-potential tree heights (the 

height of an average tree in late succession stage of stand development). As a result, 

riparian zone widths were conceived as varying among areas with different forest composition and site productivity. Riparian reserve widths could also be variable and shaped 

to local topography and potential interactions with the stream. As a result, riparian 

management areas were ecologically defined and based on the overall landscape rather 

than the tape measure.

The fixed-width riparian reserves prescribed in the Matrix land allocation of 

the NWFP, where some logging was to be permitted, were expected to be modified 

after “watershed analysis” provided a comprehensive, watershed-wide view of biotic 

and geomorphic conditions that might motivate widening or narrowing of the reserve 

widths. However, these modifications did not occur in most areas for a variety of 

reasons.

Figure 2—Effectiveness of streamside forest in providing 

litter fall, root strength in streambanks, shading, and large 

wood to the channel as a function of distance from the 

channel as measured in proportion of tree height (source: 

FEMAT 1993, public domain).
USDA Forest Service RMRS-GTR-377. 2018 29

The NWFP also charged the research-management partnership team based at the 

Andrews Forest with developing a landscape management plan based on the historic 

wildfire regime. This ecosystem-dynamics approach to landscape planning, developed 

as the Blue River Landscape Plan, contrasts with the species-specific conservation approach that dominates the NWFP (Cissel et al. 1999). A key part of the Blue River plan 

is to consider the frequency and severity of disturbance in the uplands as well as the 

streamside forest when managing for aquatic conservation objectives consistent with 

the historic disturbance regime. Several timber sales in native, mature-age-class (ca. 

150 years) forests that are part of implementing and testing the plan are completed, but 

further implementation has been stopped with the region-wide cessation of logging accomplished by environmentalists over the past 20-plus years. Despite the management 

outcomes, the multi-discipline, multi-scale research carried out at the Andrews Forest in 

past decades proved to be an important part of the foundation for both the local plan and 

the regional conservation strategy of which stream and riparian networks are a vital part.

Even as input to policy and management proceeded, research efforts were gaining new insights to the complexity of riparian systems. Several decades of study have 

revealed secrets of the hyporheic system—the down-valley, subsurface flow of water 

beneath the streambed and within the valley-floor alluvium that experiences periodic 

exchange with surface waters (Wondzell and Swanson 1996). The hyporheic system facilitates interaction of surface water senwith root systems of riparian vegetation, which 

in the westside Pacific Northwest often includes red alder, a nitrogen-fixing species. An 

isotopic nitrogen tracer study revealed that some nitrogen in streamwater can actually 

flow upward into the terrestrial system via a pathway beginning as hyporheic flow is 

taken up by riparian plants, which incorporate the dissolved nitrogen in streamwater 

into foliage, which is then consumed by herbivorous invertebrates fed upon by birds that 

integrate with the terrestrial food web (Ashkenas et al. 2003).

Given the ever-changing perceptions of riparian systems prompted by new science, new tools, and biophysical and social disturbance events, we are confident that 

the next generation of students of riparian zones will make many interesting discoveries 

of forest-stream interactions at micro-site to large watershed scales. We often wonder: 

What is it that is right in front of us now that we cannot see—just like the fallen logs we 

tripped over in the early 1970s until recognizing their importance and building much of 

our science careers around them? Young scientists and land managers beginning their 

careers have exciting opportunities for discovery in riparian systems.

Acknowledgments

We write this as a shared, personal essay, with a sense of great gratitude to many 

wild, wonderful, and too-numerous-to-name colleagues of the Stream Team based on 

the Oregon State University campus over the decades. Decades of funding and other 

forms of institutional support of NSF, OSU, and USFS Pacific Northwest Research 

Station are deeply appreciated. We dedicate this essay to the memory of Jim Sedell, 

our inspiring friend and colleague whose ashes are now part of biogeochemical system 

associated with Mack Creek in the Andrews Forest; his ideas live on as well. Jim loved 

rivers and worked for their betterment in many capacities while working in academia, 

government research, corporate, and NGO sectors.
30 USDA Forest Service RMRS-GTR-377. 2018

References

Acker, S.A.; Gregory, S.V.; Lienkaemper, G.; [et al.]. 2003. Composition, complexity, and tree 

mortality in riparian forests in the central western Cascades of Oregon. Forest Ecology and 

Management. 173: 293–308. http://andrewsforest.oregonstate.edu/pubs/pdf/pub2740.pdf 

[Accessed December 6, 2016].

Ashkenas, L.R.; Johnson, S.L.; Gregory, S.V.; [et al.]. 2004. A stable isotope tracer study 

of nitrogen uptake and transformation in an old-growth forest stream. Ecology. 85(6): 

1725–1739. http://andrewsforest.oregonstate.edu/pubs/pdf/pub3605.pdf [Accessed December 

6, 2016].

Cissel, J.H.; Swanson, F.J.; Weisberg, P.J. 1999. Landscape management using historical fire 

regimes: Blue River, Oregon. Ecological Applications. 9(4): 1217–1231. http://andrewsforest.

oregonstate.edu/pubs/pdf/pub2502.pdf [Accessed December 6, 2016].

Cummins, K.W. 1976. The ecology of running water: Theory and practice. In: Baker, D.B.; 

Jackson, W.B.; Prater, B.L., eds. Proceedings of the Sandusky River Basin symposium; 1975 

May 2; Tiffin, OH. Washington, DC: U.S. Government Printing Office: 277–393.

Forest Ecosystem Management Assessment Team (FEMAT). 1993. Forest ecosystem 

management: An ecological, economic, and social assessment. Washington, DC: U.S. 

Department of Agriculture, Forest Service; U.S. Department of Commerce, National Oceanic 

and Atmospheric Administration, National Marine Fisheries Service; U.S. Department 

of the Interior, Bureau of Land Management, Fish and Wildlife Service, National Park 

Service; U.S. Environmental Protection Agency. https://www.blm.gov/or/plans/nwfpnepa/

FEMAT-1993/1993_ percent20FEMAT_Report.pdf [Accessed November 29, 2016].

Gregory, S.; Ashkenas, L. 1990. Field guide for riparian management, Willamette National 

Forest. U.S. Department of Agriculture, Forest Service, Willamette National Forest. 65 p. 

http://andrewsforest.oregonstate.edu/pubs/pdf/pub1039.pdf [Accessed December 6, 2016].

Gregory, S.V. 1996. History of management of large woody debris in the Pacific Northwest. In: 

Cone, J.; Ridlington, S., eds. The Northwest salmon crisis: A documentary history. Corvallis, 

OR: Oregon State University Press. 384 p.

Gregory, S.V. 1997. Riparian management in the 21st century. In: Kohm, K.A.; Franklin, 

J.F., eds. Creating a forestry for the 21st century: The science of ecosystem management. 

Covelo, CA: Island Press: 69–85. http://andrewsforest.oregonstate.edu/pubs/pdf/pub2334.pdf

[Accessed December 6, 2016].

Gregory, S.V.; Swanson, F.J.; McKee, W.A.; Cummins, K.W. 1991. An ecosystem perspective 

of riparian zones: Focus on links between land and water. BioScience. 41(8): 540–551. http://

andrewsforest.oregonstate.edu/pubs/pdf/pub954.pdf [Accessed December 6, 2016].

Meehan, W.R.; Swanson, F.J.; Sedell, J.R. 1977. Influences of riparian vegetation on aquatic 

ecosystems with particular reference to salmonid fishes and their food supply. In: Johnson, 

R. Roy; Jones, Dale A., tech. coords. Importance, preservation and management of riparian 

habitat: A symposium; 1977 July 9; Tucson, AZ. Gen. Tech. Rep. RM-43. Fort Collins, 

CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range 

Experiment Station: 137–145. http://andrewsforest.oregonstate.edu/pubs/pdf/pub523.pdf 

[Accessed December 6, 2016].

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management of streamside communities. New York: Elsevier. 430 p.

Swanson, F.J.; Gregory, S.V.; Sedell, J.R.; [et al.]. 1982. Land-water interactions: The riparian 

zone, Chapter 9. In: Edmonds, R.L., ed. Analysis of coniferous forest ecosystems in the 

western United States. US/IBP Synthesis Series. 14. Stroudsburg, PA: Hutchinson Ross 

Publishing Co.: 267-291. http://andrewsforest.oregonstate.edu/pubs/pdf/pub626.pdf 

[Accessed December 6, 2016].

Triska, F.J.; Sedell, J.R.; Cromack, K.; [et al.]. 1984. Nitrogen budget for a small coniferous 

forest stream. Ecological Monographs. 54: 119–140. http://andrewsforest.oregonstate.edu/

pubs/pdf/pub1354.pdf [Accessed December 6, 2016].
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Vannote, R.L; Minshall, G.W.; Cummins, K.W.; [et al.]. 1980. The river continuum concept. 

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of a 4th-order mountain stream. I: Hydrologic processes. Journal of the North American 

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32 USDA Forest Service RMRS-GTR-377. 2018

Chapter 3. Impacts of Interacting Fire, Climate, and 

Hydrologic Changes on Riparian Forest Ecosystems in 

the Southwest

D. Max Smith and Deborah M. Finch

Introduction

Changes in human populations, water use, climate, and related disturbances are 

impacting riparian ecosystems throughout the western United States. Nowhere is this 

more pronounced than in the arid American Southwest (Gutzler 2013; Molles et al. 

1998; Webb et al. 2007). Changes in southwestern riparian ecosystems are often visible 

to the casual eye in the form of reduced and channelized water along stream courses, 

loss or changes in riparian vegetation, fire, and urbanization. To manage these changes 

and improve ecosystem resiliency for the future, a better understanding of the impacts 

of stressors and disturbances on southwestern riparian ecosystems, and especially on 

resources of high value from human and ecological perspectives, is needed. We focus 

on aridland riparian forests in this chapter owing to their values for recreation, wildlife 

habitat, and energy and nutrient input.

Aridland riparian forests are composed of species in plant guilds that vary in 

their response to surface flows and groundwater (fig. 3; Stromberg and Merritt 2015). 

a)

Figure 3—Examples of southwestern riparian forests: (a) along the Gila River in southwestern New Mexico, with a Fremont 

cottonwood and Arizona sycamore canopy and Goodding’s willow, boxelder, and other woody species in the subcanopy;
USDA Forest Service RMRS-GTR-377. 2018 33

Figure 3—Examples continued. (c) along the San Juan River in southeastern Utah with Fremont cottonwood, Russian olive, 

and saltcedar (background), which has been defoliated by leaf beetles; 

Figure 3—Examples continued. (b) near the Verde River in central Arizona with a mesquite bosque (foreground) and 

cottonwood canopy (upper center);

b)

c)
34 USDA Forest Service RMRS-GTR-377. 2018

Taxa in hydroriparian guilds include cottonwoods (Populus spp.) and willows (Salix

spp.). Mesoriparian and xeroriparian taxa include boxelder (Acer negundo), mesquites 

(Prosopis spp.), and invasive, nonnative species such as Russian olive (Elaeagnus 

angustifolia) and saltcedar (Tamarix spp). The extent of riparian forests is small relative 

to other southwestern plant communities and is largely determined by complex interactions among human activity, climate, and disturbance processes. Fremont cottonwood 

(Populus fremontii), Arizona sycamore (Platanus wrightii), and other woody species 

provide birds and other animals with nesting sites and foraging opportunities that are 

often absent in upland plant communities (Bock and Bock 1984; Carothers et al. 1974). 

The structurally diverse, species-rich vegetation along many southwestern streams 

supports high densities of territories and nest sites for a variety of birds including the 

Federally endangered southwestern willow flycatcher (Empidonax traillii extimus), the 

threatened western population of yellow-billed cuckoo (Coccyzus americanus), and 

other species of high conservation priority such as Lucy’s warbler (Oreothlypis luciae) 

(Finch et al. 2006; Friggens and Finch 2015; Smith and Finch 2014).

Given the acceleration of human influence at local to global scales and changes in 

climate, fire severity and frequency, and other stressors, it is critical to assess the effects 

and interactions of natural and altered disturbance regimes on riparian forest ecosystems 

and species in the Southwest. Such information will help us understand ecosystem vulnerability and develop actions to manage ecosystems for improved resiliency (Friggens 

et al. 2013). In this chapter we examine these effects through the lens of native and nonnative riparian woody vegetation. The Middle Rio Grande riparian forest, known as the 

“bosque,” is highlighted as a case study.

Figure 3—Examples continued. and (d) along the Middle Rio Grande in central New Mexico with Rio Grande cottonwood, 

saltcedar, and Russian olive.

d)
USDA Forest Service RMRS-GTR-377. 2018 35

Disturbance and Woody Riparian Vegetation

Flood, drought, and wildfire are primary components of disturbance regimes affecting aridland riparian forests. These disturbances can be caused by natural climate 

and hydrological cycles, human activity, or their interactions. To evaluate the state of 

knowledge on these disturbances, we searched for papers describing the response of 

native and nonnative woody species to flood, drought, and wildfire. We focused on 

seven taxa common to these forests: Fremont cottonwood and Rio Grande cottonwood 

(Populus deltoides ssp. wislizenii), combined hereafter as “cottonwood”; Goodding’s 

willow (Salix gooddingii); velvet mesquite (Prosopis velutina) and honey mesquite 

(Prosopis glandulosa), combined hereafter as “mesquite”; Arizona sycamore; boxelder; 

Russian olive; and saltcedar. We found studies of drought and flood effects by searching 

for the common and scientific names of each species, along with “flood” and “drought” 

in the online citation service Web of Science. To find studies of wildfire effects we 

used the same procedure in Web of Science and expanded our search by using Google 

Scholar.

Effects of Flooding and Flood Reduction

Flood is the most frequently studied disturbance for all taxa but mesquite (fig. 4). 

In the Southwest, floods can occur throughout the year, fueled by snowmelt, rain, or 

a combination of the two. High-magnitude flows occasionally create enough shear 

stress, erosion, and sedimentation to kill individuals or remove entire stands of shrubs 

or trees (Bock and Bock 1989; Friedman and Auble 1999; Minkley and Clark 1984). 

Conversely, floods are in some way instrumental for reproduction of most, if not all 

woody riparian species. Vegetative reproduction occurs when above- or below-ground 

portions of woody plants are transported by flood and buried by sediment (Rood et al. 

2007). Floods also create opportunities for germination of pioneer species including cottonwood, Goodding’s willow, and Arizona sycamore (Stromberg and Merritt 2015) and 

produce germination sites by delivering damp sediment and removing litter and competing vegetation (Braatne et al. 1996; Stromberg 1997, 2002). By wetting unsaturated 

Figure 4—Number of publications reporting effects 

of flood, drought, and wildfire on seven woody 

riparian species in North America. Flood and 

drought papers were found by searching for 

common and scientific names with “flood” and 

“drought” in Web of Science; wildfire papers 

were found using the same procedure in Web of 

Science and Google Scholar.
36 USDA Forest Service RMRS-GTR-377. 2018

soil layers, flood waters induce germination and support the growth of seedlings until 

their roots have reached the saturated layer of soil, forming a connection necessary for 

survival in arid conditions (Bhattacharjee et al. 2006; Horton and Clark 2001; Mahoney 

and Rood 1998).

Boxelder, mesquites, and Russian olive have relatively large, long-lived seeds that 

can germinate in the presence of litter and competing vegetation. Though their seeds 

do not require flood-exposed sites, flood-dampened soil is needed for germination, 

especially if rainfall is insufficient (Katz and Shafroth 2003; Dewine and Cooper 2007; 

Stromberg 1993). Mesquite seedlings do not survive well in saturated soil, so mesquite 

bosques often occupy the outer portions of riparian zones, beyond the area occupied by 

cottonwood, willow, and sycamore (Stromberg 1993). Where surface flows are intermittent, floods are needed to both replenish groundwater aquifers and deliver nutrients to 

the soil—processes critical to growth and survival of individuals across guilds and age 

classes (Stromberg 2001a).

At many streams, frequency, magnitude, and timing of floods have been altered 

by diversions and dams, with well-documented effects on woody riparian plants (Webb 

et al. 2007). Though cottonwood and willow recruitment often occurs along narrowed 

streams following dam construction, opportunities for further reproduction are limited 

by the reduction of flood scour and sediment deposition (Coble and Kolb 2013; Howe 

and Knopf 1991; Merritt and Poff 2010; Shafroth et al. 2002). Reduction in magnitude 

of floods can also limit recruitment of mesoriparian and xeroriparian species that grow 

at higher elevations of the floodplain (Coble and Kolb 2013; Stromberg 1993). In turn, 

flow modification encourages establishment of Russian olive, which does not require 

flood-dampened soil or exposed sites for germination (Katz and Shafroth 2003).

Timing of flow events, such as floods and baseflows, influences reproduction 

of woody species and riparian forest composition (Beauchamp and Stromberg 2007; 

Birken and Cooper 2006). A shift in timing of peak discharge away from the dispersal 

period of pioneer species’ short-lived seed will prevent their germination (Fenner et al. 

1985). Saltcedar has similar requirements for germination as cottonwood, Goodding’s 

willow, and Arizona sycamore, but it has a longer, more variable period of seed release 

(Stevens and Siemion 2012). Changes in timing of peak discharge, caused by regulation, 

can therefore encourage replacement of native species by saltcedar.

Drought Effects

The water required by riparian trees for reproduction, growth, and survival is 

accessed from surface flows, groundwater, and precipitation (Kolb et al. 1997; Snyder 

and Williams 2000). These water sources are also linked to one another through natural 

hydrological processes (Webb and Leake 2006). Periodic shortages in these sources occur naturally across climatic cycles, but for over a century shortages have been caused 

or exacerbated by surface flow diversion, groundwater withdrawal, and reservoir storage 

(Phillips et al. 2011; Summitt 2013). Studies have examined effects of stream drying, 

both natural and human-induced, in a variety of native and nonnative taxa, with cottonwood, mesquite, and saltcedar receiving the most attention in the literature (fig. 4).

Cottonwood, Goodding’s willow, and Arizona sycamore are hydroriparian taxa 

that can form an extensive forest canopy where groundwater remains accessible to their 
USDA Forest Service RMRS-GTR-377. 2018 37

shallow roots (Lite and Stromberg 2005; Stromberg 2001b). Of these taxa, Goodding’s 

willow is the most dependent on shallow depth to groundwater and often establishes 

closest to the stream channel. Mesquite, boxelder, Russian olive, and saltcedar are mesoriparian and xeroriparian taxa that can establish throughout a floodplain (Dewine and 

Cooper 2008; Stromberg et al. 2007). Mesquites are capable of attaining large size and 

forming extensive stands known as bosques. For mature mesquite bosques to form, trees 

must have access to groundwater at depths of 15 m or less, though 6 m or less is ideal 

(Stromberg 1993).

When depth to groundwater increases, hydroriparian species respond rapidly 

through reduction of growth, branch dieback, and stem mortality (Coble and Kolb 2012; 

Stromberg et al. 2007). Mesoriparian and xeroriparian species can therefore gain a competitive advantage over hydroriparian species when moderate drought occurs (Cleverly 

et al. 1997; Horton et al. 2001a, 2001b; Lite and Stromberg 2005). Severe droughts and 

flow modification can reduce growth, decrease reproduction, and increase mortality of 

mesoriparian and xeroriparian species as well, leading to replacement by small-stature 

upland plants (Coble and Kolb 2012, 2013; Dewine and Cooper 2007; Stromberg 1993).

Fire Effects

The effects of wildfire on riparian vegetation have received little research attention 

relative to effects of flood and drought (fig. 4). Fire has long been studied as a method 

to control the spread of mesquites in rangelands (Blydenstein 1957), but little is known 

about their response to fire in a riparian setting. Most studies of fire in aridland riparian systems have focused on cottonwoods and saltcedar, largely due to concerns that 

wildfire will facilitate replacement of the former by the latter (Busch and Smith 1995; 

Drus 2013; Smith et al. 2009b). Fire effects have been documented infrequently for the 

other woody taxa (fig. 4). The responses of woody plants to fire in these studies generally include rates of topkill (death of above-ground tissues) and resprouting (production 

of basal sprouts, epicormic sprouts, and root suckers). Postfire germination has been 

observed for cottonwoods by Ellis (2001), but direct measurements are not reported in 

the literature. Boxelder is a widespread and important component of wildlife habitat 

(Brodhead et al. 2007; Stoleson and Finch 2003), but we did not find reported effects of 

fire on this species.

Topkill vulnerability varies among taxa, size classes, and fire severity (Bock and 

Bock 2014; Ellis 2001; Stuever 1997). To our knowledge, all deciduous riparian tree 

taxa can recover from topkill by producing basal sprouts, epicormic sprouts, or root 

suckers. The success of vegetative recovery, however, is affected by numerous factors 

that vary among species and wildfire sites (Eillis 2001; Smith et al. 2009b; Stromberg 

and Rychener 2010). Examinations of fire that take into account flow modifications, 

native and nonnative plant responses, and response of animal communities to changes 

in vegetation are needed. Below we provide an example of this type of study, conducted 

along the Middle Rio Grande in central New Mexico.

Middle Rio Grande Case Study

The Middle Rio Grande in central New Mexico is anthropogenically modified, but 

many stretches still support native trees and shrubs, which provide habitat for wildlife 
38 USDA Forest Service RMRS-GTR-377. 2018

including riparian obligates and threatened and endangered species (Friggens et al. 

2013; Smith and Finch 2014). With several decades of research on hydrology, wildfire, 

plants, and animals, this area is an ideal case study of changing disturbance regimes and 

their ecological effects.

Natural Hydrological Regime

The Middle Rio Grande is the section of the Rio Grande that flows north to south 

through central New Mexico. The Rio Grande originates in and receives most of its 

surface flow from the San Juan Mountains of southwestern Colorado (Phillips et al. 

2011). Peak flows, resulting from snowmelt runoff, historically occurred during the late 

spring or early summer (fig. 5). Floods would cause the channel to migrate throughout 

the broad alluvial valley, leaving a mosaic of wetlands and multi-aged stands of riparian 

vegetation (Scurlock 1998). Floods also occurred following heavy thunderstorms during 

the summer monsoon and other times of the year, but spring snowmelt provided the opportunities for pioneer tree establishment.

Changes to the Regime

Alterations to the Middle Rio Grande have occurred over several centuries, but the 

stream became profoundly regulated during a period bookended by the formation of the 

Middle Rio Grande Conservancy District in 1929 and the completion of Cochiti Dam 

in 1974. To reduce flooding and improve agricultural activity, government agencies 

constructed a network of diversion dams, levees, irrigation canals, and drains (Phillips et 

al. 2011; Scurlock 1998). The levees currently prevent the river from meandering across 

most of the natural floodplain. Within the levees, the stream bank has been stabilized to 

limit the movement of the active channel. Cochiti Dam, located at the north end of the 

Middle Rio Grande, has reduced the magnitude of peak flows, but releases from the dam 

have maintained the seasonal timing of fluctuations (Braun et al. 2015). During years of 

low runoff, large stretches of the channel become dry as water is diverted for irrigation. 

Figure 5—Mean daily discharge 

hydrograph at the Rio Grande at 

Otowi, upstream from Cochiti 

Dam. Line values represent mean 

daily discharge for each day of the 

year averaged over the period of 

1919 to 2013. Data were obtained 

from the United State Geological 

Survey National Water Information 

System and are available online 

(http://waterdata.usgs.gov/nwis) 

(photo by Max Smith).
USDA Forest Service RMRS-GTR-377. 2018 39

In years with heavy runoff, flood pulses, released from Cochiti Dam, inundate some 

portions of the area between the levees.

Effects on Riparian Trees

Prior to extensive regulation, riparian forests and herbaceous wetlands were scattered throughout the Middle Rio Grande floodplain in a variety of age classes and seral 

states (Whitney 1996). Midcentury confinement and channelization of the river led to 

the establishment of Rio Grande cottonwood bosque that is now sandwiched between 

the stream channel and the levees. During this time, herbaceous wetlands were lost 

due to the lowering of water tables and the cessation of channel migration (Crawford 

et al. 1993). Russian olive, saltcedar, and other nonnative woody species have been in 

the Middle Rio Grande Basin since the 1930s or earlier and are established throughout 

the bosque (Scurlock 1998). In terms of stem density, saltcedar is now the numerically 

dominant woody species in the bosque, followed by cottonwood and Russian olive 

(figure 3d, Smith and Finch 2014). Cottonwoods and other native trees may die during 

periods of low precipitation and runoff when depth to groundwater increases (Smith et 

al. 2009b).

As a result of modification and invasives, composition and arrangement of riparian 

vegetation are very different from conditions prior to Euro-American settlement. Native 

and nonnative species now grow in narrow, dense stands that extend from Cochiti Dam 

to Elephant Butte Reservoir. Without management intervention, flood magnitude is not 

great enough to scour the forest floor of woody vegetation and deposit sediment for 

establishment of hydroriparian species. In response, cottonwoods and Goodding’s willows may be replaced by nonnative trees in the coming decades (Howe and Knopf 1991; 

Molles et al. 1998).

Historical/Current Wildfire Effects

As the role of flooding has diminished along the Middle Rio Grande, wildfire has 

grown from a minor component of the disturbance regime to an increasingly important 

influence on plants and animals (Crawford et al. 1993; Finch et al. 2006; Stuever et al. 

1995; Williams et al. 2007; Bess et al. 2002). Most fires that enter the bosque are accidentally ignited and burn with mixed intensity until they are contained by firefighting 

crews. The number of ignitions increases with proximity to larger towns and cities, but 

the size of fires increases with distance from these areas (Williams et al. 2007). Most 

fires occur in the dry spring and early summer period and the number of fires tends to 

be greater during years with low precipitation (Stuever et al. 1995). Since the beginning 

of a long-term drought in 2000, at least 40 percent of a 732-ha study area has burned 

and some portions have burned multiple times (Smith and Finch 2017). With the lack 

of high-magnitude flooding, the bosque has accumulated large quantities of litter and 

woody debris. Nonnative plant species, especially Russian olive and saltcedar, have 

spread as soil moisture declined, increasing the density of woody plants in the understory (Bateman et al. 2008). These fuels, combined with the spatial arrangement of the 

forest, have contributed to increasing fire sizes and intensities that are likely outside the 

natural range of variability in the Middle Rio Grande (Bateman et al. 2008; Ellis 2001; 

Johnson and Merritt 2009).
40 USDA Forest Service RMRS-GTR-377. 2018

Response of Woody Plants to Wildfire

Several studies have examined responses of native and nonnative woody vegetation to wildfire in the Middle Rio Grande. Results are varied, indicating that spatial and 

temporal factors interact with fire in shaping riparian forest composition. Several patterns, however, have emerged with the response of cottonwood and other woody plants.

Cottonwoods are extremely vulnerable to topkill from fires that enter the bosque 

(table 1). Most, if not all, above-ground mortality occurs immediately after fire. In areas 

where fire severity is high (all organic matter is consumed on the forest floor), all cottonwoods and other woody species are top-killed. Topkill rates are lower for trees in 

areas burned with moderate severity (78 to 100 percent of trees killed) and light severity 

(52 to 70 percent of trees killed) (Ellis 2001; Stuever 1997). Managed flooding and 

mechanical fuel reduction can increase resistance to topkill of cottonwood, but only 

to a limited extent given their high vulnerability to fire and the difficulty of removing 

all sources of fuel from the understory (Ellis 2001; Johnson and Merritt 2009). Woody 

riparian plants recover vegetatively from wildfire through production of basal sprouts, 

root suckers, and (in the case of cottonwoods) epicormic sprouts. Their production and 

survival, however, vary among species and study sites (Ellis 2001; Smith et al. 2009b). 

We observed epicormic sprouting of cottonwoods only at one site that burned with light 

to moderate severity.

We observed postfire germination of cottonwoods at a site burned in March of 

2008 and partially flooded in June of that year. Ellis (2001) also observed saplings in 

postwildfire sites along the Middle Rio Grande that were flooded within 2 years of 

being burned. A combination of fire and low-magnitude flooding can therefore act as a 

replacement for high-magnitude flooding along this heavily regulated stream.

Ecosystem Implications of the Current Disturbance Regime

Despite its highly regulated state, the Middle Rio Grande supports a unique assemblage of wildlife species. In unburned portions of the bosque, large cottonwoods provide 

nest sites, shelter, and food for reptiles, birds, and mammals (Finch et al. 2006; Smith 

and Finch 2014; Smith et al. 2006). Cavities in cottonwood snags and broken branches 

of live cottonwoods are used for nesting and roosting sites. Smaller woody plants, 

including Russian olive and saltcedar, provide resources such as nest sites for birds in 

the shrub and subcanopy nesting guilds (Finch et al. 2006; Smith and Finch 2014; Smith 

Table 1—Estimates of top-kill and basal sprouting rates at four sites along the Middle Rio Grande. The San Pedro burn sites 

were measured by Ellis (2001) in 1996. We measured the 3-4-6 and Sevilleta burn sites in 2013.

Topkill Basal sprouting

San Pedro 

Burn 1

(1996)

San Pedro 

Burn 2

(1996)

3-4-6 

Burn

(2011)

Sevilleta 

Burn

(2011)

San Pedro 

Burn 1

(1996)

San Pedro 

Burn 2

(1996)

3-4-6 

Burn

(2011)

Sevilleta 

Burn

(2011)

 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .Percent. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Cottonwood 100 97.7 100 95 77 81 21.2 56.8

Saltcedar 100 100 95 NA 53 55 84.9 NA

Russian olive NA NA 97 100 NA NA 85.7 75.7
USDA Forest Service RMRS-GTR-377. 2018 41

et al. 2009a). Wildfire removes cottonwood canopy, creates snags and fallen debris, 

and induces resprouting of woody plants, especially saltcedar. These changes to forest 

structure and composition create habitat features used by many animal species, but 

make the bosque unsuitable for others (Smith et al. 2007, 2012). As postfire succession 

occurs, density of canopy-dependent species will decline if cottonwoods do not recover. 

In addition, cavity-associated species will lose nest and roosting sites if fallen snags are 

not replaced by mature trees.

Under the current disturbance regime, mortality of large riparian trees will continue to increase due to wildfire, drought, and senescence (Smith and Finch 2015). 

Vegetative and sexual reproduction of cottonwood and other native trees may occur 

under certain circumstances, such as fire and flood events and management intervention 

that are limited in spatial scale (Howe and Knopf 1991). Nonnative woody species, such 

as Russian olive and saltcedar, are present throughout the Middle Rio Grande and will 

likely increase in abundance as cottonwood declines. With high densities of mesoriparian and xeroriparian growth, postwildfire sites are vulnerable to additional high-severity 

fires and may enter a positive feedback loop, to the detriment of native hydroriparian 

species (Drus 2013). Postwildfire replacement of cottonwood by Russian olive and 

saltcedar will change the structure of the Middle Rio Grande riparian forest by increasing the density of low-stature vegetation and decreasing canopy height, in turn affecting 

habitat quality for wildlife, including riparian-nesting birds.

Implications of Climate Change

Projected effects of climate change are particularly severe in the Southwest 

(Gutzler and Robbins 2011; Seager et al. 2007). Increasing temperatures and changes in 

precipitation will affect characteristics of streams, which in turn shape aridland riparian 

ecosystems (Seager et al. 2013) and their capacity to support wildlife (Friggens and 

Finch 2015). For example, global climate models predict that, in the Rio Grande Basin, 

increasing temperatures will result in decreased snowpack, decreased runoff, and earlier 

peak discharge (fig. 6). Such changes could further limit reproduction of cottonwoods 

and willows, increase drought mortality, and decrease their ability to recovery from 

wildfires relative to Russian olive and saltcedar (Smith et al. 2009b). In addition, storms 

such as those during monsoons may gain strength in response to warming. Resulting 

floods could induce late summer germination of mesoriparian and xeroriparian species, to the exclusion of species with narrower windows of seed viability (Dewine and 

Cooper 2007; Fenner et al. 1985; Katz and Shafroth 2003). Replacement of hydroriparian species by mesoriparian and xeroriparian species would have cascading effects on 

the riparian forest community.

Conclusions: Disturbance Change and the Future of Aridland Riparian 

Forests

Studies of woody riparian plants show that effects of disturbances on survival 

and reproduction vary among species. As our Middle Rio Grande study shows, native 

hydroriparian pioneer species are vulnerable to changes in streamflow, which, coupled 

with wildfire effects, could open doors to invasion by xeroriparian and upland species. 

As with other aridland rivers, the Middle Rio Grande has created an extensive riparian 
42 USDA Forest Service RMRS-GTR-377. 2018

corridor critical to the successful migration, reproduction, and overwintering of myriad 

wildlife species (Carothers et al. 1974; Knopf and Sampson 1994; Bateman et al. 2008). 

Maintaining the composition of these corridors is necessary to preserve both regional 

and continental biodiversity.

During the previous decades of riparian research, ecologists have highlighted the 

importance of woody riparian vegetation to wildlife and have described the response 

to changes in hydrology. For most riparian taxa, however, response to wildfire, especially in combination with drought, is poorly known or has been examined in only a 

few locations. As climate changes and wildfire becomes more frequent than flooding, 

information on how fire, drought, and climate change affect riparian vegetation will 

be critical for managers in maintaining ecosystem structure and stability, wildlife 

habitat, and the associated animal populations and communities. Because riparian 

dynamics, including recovery from wildfire, are coupled with hydrology of regulated 

streams, we need hydrological projections that incorporate future water use and 

climate change scenarios. With this information, we can determine which species of 

plants will naturally sustain themselves and which will require adaptive management 

in an increasingly arid Southwest.

Figure 6—Temperature and hydrological projections for the Rio Grande Basin. Boxes A–C contain projections by 

29 models run under the CMIP5 RCP8.5 scenario and 16 models run under the CMIP3 A2 scenario. Box D 

contains projections of peak discharge date for the Rio Grande at Otowi from 15 models run under the CMIP3 

A2 scenario. Projections were made by the U.S. Bureau of Reclamation and are available online (http://gdodcp.ucllnl.org/downscaled_cmip_projections/dcpInterface.html).
USDA Forest Service RMRS-GTR-377. 2018 43

Acknowledgments

We appreciate input from Megan Friggens and comments from multiple reviewers. 

Funding for this work was provided by the Desert and Southern Rockies Landscape 

Conservation Cooperatives and the USDA Forest Service Rocky Mountain Research 

Station.

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USDA Forest Service RMRS-GTR-377. 2018 47

Chapter 4. Invasion and Restoration of Western Rivers 

Dominated by Tamarix spp.

Hisham N. El Waer, Annie Henry, Katie Merewether, and Anna A. Sher

Introduction

Scientists, land managers, government, and private institutions in the United States 

have given much attention to invasive control and restoration projects along western 

rivers; in the West, removal of Tamarix spp. (tamarisk, saltcedar) has been a primary 

focus of these projects (Dennison et al. 2009; González et al. 2017a; Harms and Hiebert 

2006; Shafroth et al. 2008). These trees were first introduced to North America during 

the 1800s from Eurasia mainly to decrease erosion, to be wind breaks, and to slow 

down water flow in riparian and agriculture areas (DiTomaso 1998). Since that time, 

this invasive tree has become the third most common woody species and second highest 

tree cover in the southwestern United States (Friedman et al. 2005). Despite the fact 

that invasive plant species removal has long been a priority in restoration of riparian 

ecosystems (González et al. 2015), we are only beginning to understand the ecological 

impact of the removal of invasive species from such ecosystems. Here we will provide 

an overview of Tamarix ecology along rivers in the western United States and the results 

of our research monitoring plant communities in 25 riparian sites over 3 years, with a 

particular interest in the results of removal of Tamarix by various methods including 

biological control by a defoliating beetle.

Tamarix Introduction and Impact

During the 1800s, eight species of Tamarix were first introduced to the United 

States from Europe, Asia, and North Africa. In the 1920s, Tamarix spread and occupied about 4,000 ha of riparian habitat in the southwestern United States. By 1987, 

the area invaded by Tamarix increased to about 600,000 ha (Brock 1994; DiTomaso 

1998; Gaskin and Schaal 2002; McDaniel et al. 2005; Nagler et al. 2011) and it now 

occupies approximately 800,000 ha. The most widely naturalized species are Tamarix 

ramosissima, Tamarix chinensis, and their hybrid (Friedman et al. 2005; Gaskin and 

Schaal 2002).

Tamarix is classified as deciduous with either a tree or shrub growth habit. It is 

also a paradoxical plant from the selection theory perspective as it uses both r and K 

strategies (sensu McArthur and Wilson 1967) with both a large number of offspring and 

high longevity (Sher 2013). Tamarix individuals can live for more than 100 years, and 

one large tree produces about 500,000 seeds per year (McDaniel et al. 2005; Stromberg 

et al. 2007b). It is highly tolerant of drought; its evapotranspiration rates are highly 

variable and adaptable to water availability, and as a facultative phreatophyte, uses both 

surface and ground water (Glen and Nagler 2005). Its deep roots can reach declining 

water tables when other species cannot (Shafroth et al. 2000). Tamarix also has adaptations that allow it to tolerate greater salinities than native vegetation (Shafroth et al. 
48 USDA Forest Service RMRS-GTR-377. 2018

1995). Together, these traits have allowed Tamarix to take advantage of the environmental stresses associated with dammed rivers (Sher 2013; Stromberg et al. 2007b).

Under natural river flow conditions, native cottonwoods (Populus spp.) are typically competitive over invasive Tamarix (Huston 2004; Sher et al. 2000, 2002; Stromberg 

et al. 2007a), suggesting that Tamarix would not thrive under natural conditions. Rather, 

as a result of widespread flow control, many riparian ecosystems of the Southwest no 

longer support native vegetation, allowing Tamarix to dominate (Sher 2013).

Thus, Tamarix dominance is the result of ecosystem change, but it can also be the 

cause (passenger vs. driver, sensu MacDougall and Turkington 2005; Johnson 2013). 

Like many invasive species, once Tamarix establishes it can also alter the ecosystems in 

which it occurs, increasing fire risk (Drus 2013), changing the morphology of the stream 

bank (Auerbach et al. 2013), and increasing the soil salinity (Ohrtman et al. 2012), 

among other effects (Didham et al. 2005; Johnson 2013; Sher 2006). Such changes add 

to the inhospitable quality of the riparian area following river flow regulation to native 

plant communities.

All of these environmental changes mean that Tamarix-dominated stands are associated with lower diversity in plant and animal communities (Sher 2013), making 

Tamarix removal a high priority for many private and public institutions, despite its 

considerable cost (de Wall 1994; Zavaleta 2001). However, the dual role of Tamarix as 

both ecosystem passenger and driver has profound implications for restoration (Bauer 

2012). As a driver of ecosystem change, removing Tamarix should solve many of the 

concerns for conserving riparian areas. No doubt, this was the perspective of early aims 

to remove Tamarix (see Chapter 1). However, as a passenger of ecosystem change, its 

removal alone should affect no positive change; recent research suggests at least in 

some cases without natural flow regimes, native riparian vegetation cannot reestablish 

(González et al. 2017a; Merritt and Poff 2010). Furthermore, some changes that Tamarix

make to riparian ecosystems are difficult to reverse simply by removing Tamarix itself, 

such as soil salinization (Merritt and Shafroth 2012). Thus it is imperative that we carefully monitor the ecological consequences of Tamarix removal projects.

Since Tamarix removal sites are not selected randomly, monitoring at a single 

point in time or combining sites with different periods of time since Tamarix removal 

can prove misleading. Variables such as weather patterns (e.g., dry year coinciding with 

single time monitoring) or variation in time since removal can confound analyses and 

lead to misinterpretation of results. Multiple years must therefore be sampled to determine whether patterns of recovery are real or a product of confounding variables such as 

years since removal, drought years, or flood events at a particular location in a particular 

year. Because of variability associated with geographic location, it is also imperative 

that sites of the same age but different treatments be sampled at the same time. Rarely 

have both been done in a single project.

Vegetation Response to Tamarix Removal

While invasive plant species removal has long been a priority in the restoration of 

river systems, relatively little is understood about the ecological impact of the removal 

of invasive species from the ecosystem (Bay and Sher 2008; Cuevas and Zalba 2010; 

Ostoja et al. 2014; Shafroth et al. 2005). Previous research on plant community response 
USDA Forest Service RMRS-GTR-377. 2018 49

to Tamarix removal is mixed. Harms and Hiebert (2006) surveyed 33 Tamarix removal 

and non-removal sites where only passive revegetation had been done. They found a 

decrease in the cover of invasive Tamarix compared to the control sites and a significant 

increase in native foliar cover in the Mojave region.

However, there was no consistent change in native cover in the two other regions 

sampled, and when Tamarix was excluded from data analysis, they found no difference regarding species composition across all sites. In contrast, in a similar study but 

with active revegetation, Bay and Sher (2008) found increases in native cover after 

tamarisk removal could be explained by several factors, including that the relative 

cover of planted or native species was greater in the sites when the removal period 

was greater than 8 years, when there was lower salinity, when there was coarser 

soil texture, and when closer to a body of water. More recently, Ostoja et al. (2014) 

examined short-term responses of riparian plant communities to Tamarix removal. 

They found increased species richness and diversity, but still very low absolute diversity and abundance of native species. Most of these studies did not control for the 

number of years since Tamarix removal and did not follow the same sites over time. 

Furthermore, none examined the response of plant communities to Tamarix control 

using the biological control agent, Diorhabda spp.

Diorhabda spp. is a beetle herbivore that defoliates Tamarix spp. in its native 

range. After extensive research on host specificity, it was approved for release as a 

biological control agent in 2001 (DeLoach et al. 2003; also see Chapter 5). Since its 

release, Diorhabda has moved farther and faster than anticipated (Nagler et al. 2014). 

Studies so far show that its impact on Tamarix as well as riparian plant communities is 

variable and difficult to predict (González et al. 2017a; Hultine et al. 2015; Kennard et 

al. 2016).

In the largest review to date of the consequences of Tamarix removal, González et 

al. (2017a) compiled monitoring data for plant communities following Tamarix removal 

across the Southwest in 416 sites, including those with Diorhabda spp. biological 

control. They found that for most removal methods, native cover significantly increased 

following Tamarix removal, but the increases were extremely gradual, averaging 

less than 2 percent per year. Native cover was not consistently increased with active 

revegetation, and the community did not typically revert to the mesic plant species that 

had occurred there historically. Exotic species also increased in cover after Tamarix

removal (“secondary invasion”), particularly with high disturbance treatment methods, 

even though exotic cover decreased over time in biocontrol treatment sites (González et 

al. 2017b). However, the González et al. (2017b) study amassed data from regions all 

over the western United States, possibly obscuring the importance of regional effects. 

Nowhere in the literature is there a study that compares regions that were measured at 

the same times in the same ways with comparable removal dates over time.

In Colorado, biocontrol beetles were released in 2005 on the Western Slope and 

have become nearly ubiquitous there. At the time of this study (2010–2012), the biocontrol had been released but not yet well-established on the Eastern Slope. This situation 

provided a unique opportunity to study the impact of Diorhabda spp. alone and in 

combination with other removal techniques on Tamarix cover, as well as the consequent 

effect on the rest of the plant community. This is one of the only studies in which multiple sites in different regions have been sampled by the same methods for multiple years 
50 USDA Forest Service RMRS-GTR-377. 2018

and multiple seasons within a year, thus representing a unique opportunity to investigate 

the impact of Tamarix removal on plant communities.

Sampling Methodologies and Invasive Removal Techniques

We monitored vegetation response after invasive Tamarix removal in sites east and 

west of the Rocky Mountains in Colorado for 3 and a half years beginning in fall 2009. 

This project aimed to establish baseline data for long-term monitoring and to develop 

the best practices to make recommendations for monitoring by land managers.

Our broad interest was understanding the response of plant communities along 

rivers in Colorado to the removal of a dominant invasive species. Our specific questions 

were: (1) How do riparian plant communities differ between the Eastern and Western 

Slope of Colorado that may influence response to restoration? (2) Are removal methods 

effective in reducing percent cover of Tamarix? (3) Does it matter whether the biological control is present or not? (4) How do exotic and native plants respond to Tamarix

spp. removal in each of these regions? Our intent, by measuring the impact of Tamarix

removal in the ecosystem via the measurement of vegetation parameters over time, is to 

help answer some of the controversial questions about the ecological impact of Tamarix

removal on these ecosystems, including the presence of the biological control.

Methods and Site Locations

We monitored vegetation in a total of 25 sites: nine sites in three reaches located in 

western Colorado (fig. 7a) and 16 sites in 5 reaches located in eastern Colorado (fig. 7b). 

The western sites, at approximately 38 ̊1ʹ0ʺ N 108 ̊49ʹ26ʺ W, are located in the Upper 

Dolores Watershed including Big Gypsum Valley, Disappointment Valley, and Slickrock 

Canyon. The eastern sites, at approximately 37 ̊ 33ʹ 0ʺ N 103 ̊38ʹ 21ʺ W, are located in 

the Purgatory Watershed including Chacuaco Creek, Plum Creek, and Apishapa River. 

These sites were mostly on private land in the East and BLM land in the west, selected 

by land managers for our group to survey either because they were candidates for 

tamarisk removal, or they represented an un-invaded ecosystem. Although they were 

not randomly selected, the sites represent a range of representative riparian ecosystem 

conditions, from degraded to fairly un-impacted. Native-dominated sites are referred to 

throughout this chapter as positive reference sites (González et al. 2015). The Society 

for Restoration International (SERI) Primer states that it is imperative to identify sites 

that represent the goal (positive references) of restoration (SERI 2004). Sites dominated 

by Tamarix where there was no active removal are referred to as negative reference 

sites. Active removal refers to the removal of Tamarix by the following methods: cut 

stump (chainsaws are used cut down Tamarix and herbicide is applied to the remaining 

stump); track hoe (Tamarix is removed using large machinery); or lastly, spraying herbicide by helicopter.

By employing both sampling over time and reference sites, we address potential 

issues of confounding time and space, referred to as a BACI approach: Before vs. After 

treatment (within the same site) plus Control vs. Impact (comparing sites with versus 

without active removal within the same point in time) (Bernhardt et al. 2007; Palmer 

et al. 2005). In this way, we address the shortcomings of most other studies in which 

results could be attributed to either time or space rather than the treatment itself.
USDA Forest Service RMRS-GTR-377. 2018 51

Within each site, a 20 m x 50 m plot was established, within which five stratified 

random transects were established that were sampled at 10 cm intervals using the line 

point intercept method. All sites were sampled in this way in spring of 2010 and in 

spring and summer of 2011 and 2012. Sites in the East were also sampled in the summer 

of 2010. The Modified Whittaker Plot method was used to record species richness at 

each sampling period in each site (Stohlgren et al. 1995).

All specimens that could not be positively identified in the field were collected 

and taken to Denver Botanic Gardens’ Kathryn Kalmbach Herbarium (KHD) for identification. A total of 412 specimens were collected by the end of the last season of data 

collection on August 13, 2012. Pictures were taken of all specimens and stored in a 

digital form to serve as a backup for the original.

Specimens were identified with the assistance of local plant experts and the following books: Colorado Flora Eastern Slope (Weber et al. 1996a), Colorado Flora Western 

Slope (Weber et al. 1996b), Illustrated Key to the Grasses of Colorado (Wingate 1994), 

Shrubs and Trees of the Southwest Uplands (Elmore 1976), and Weeds of the West 

(Whitson and Burrill 2000). All specimens identified by dichotomous key were doublechecked against stored reference specimens and confirmed by staff at KHD.

Digital plant databases were used to confirm current species information. 

Geographic distribution, plant morphology, spelling of scientific and common names, 

nativity status, growth habit, and plant functional group of each specimen were determined using United States Department of Agriculture Plant Database (USDA 2010), 

Figure 7—(a) Site locations for west slope of Colorado, a total of 9 sites. (b) Site locations for eastern plains of Colorado, a 

total of 16 sites (map by Annie Henry).

D o l o r e s R i v e r

. 0 2.5 5 10 15 20

Kilometers

a

A p i s h o p a R i v e r

P u r g a t o i r e R i v e r

C h a c u a c o C r e e k

b
52 USDA Forest Service RMRS-GTR-377. 2018

Colorado State University (CSU Herbarium 2001), and Southwest Environmental 

Information Network (SWEIN 2012).

Research Findings and Implications

Regional Differences

We identified a total of 145 different species in 45 plant families within the 25 

sites from 2010–2012. The Eastern Slope sites had greater species richness than the 

Western Slope sites during this study period; 111 species in 41 families on the Eastern 

Slope compared to only 53 species in 25 families on the Western Slope. The most 

important families in the study area as indicated by highest numbers of species, in descending order are: (1) Poaceae—35; (2) Asteraceae—34; (3) Chenopodaceae—10; (4) 

Fabaceae—7; and (5) Brassicaceae—6.

The plant communities differed in composition between the two areas (fig. 8). 

Understory vegetation cover in western Colorado was dominated by shrubs (Ericameria 

nauseosa, Chrysothamnus linifolius Greene, Artemisia tridentate, Atriplex canescens, 

Sarcobatus vermiculatus, Forestiera pubescens, Rhus trilobata) while exotic herbaceous 

species (Bromus tectorum, Bromus japonicas, Kochia scoparia) dominated eastern 

Figure 8—Mean (+/- 1 SE) cover for east and west slope sites by functional group. (Chi-square, Pearson; N = 699, DF = 8, X2 

= 26.34, P < 0.0009.)
USDA Forest Service RMRS-GTR-377. 2018 53

Colorado. With regard to trees, both regions were dominated by Tamarix spp., Populus

spp., and Salix exigua, but the eastern sites also contained Juniperus scopulorum and 

Celtis reticulata, while in the west the only other species was Acer negundo.

The Jaccard Index of similarity indicated that there was 36 percent similarity 

between eastern and western sites using all vegetation types, and 50 percent similarity 

in tree and shrub communities. East and west sites differed significantly in frequency 

of different functional groups (Chi-square, Pearson; N = 699, DF = 8, X2 = 26.3, P < 

0.0009). While there were distinct regional differences in the understory plant community by functional group, both regions had similar starting absolute cover of Tamarix at 

about 30 percent, and less than 10 cover percentage of native woody vegetation, as well 

as similar cover of understory natives and exotics.

The Response of Tamarix to Removal Efforts

There was a significant decrease in Tamarix cover after active removal with a 

more dramatic decrease on the Western Slope (repeated measures ANOVA, before/

after*slope: F = 4.13, DF = 1/83, P < 0.05). Over time, we can see the decrease in the 

total absolute cover of Tamarix immediately following active removal in the spring 

2011, but the trajectories differ between east and west sites after that (fig. 9). Western 

Slope Tamarix cover remained low, whereas Tamarix increased slightly in the eastern 

Figure 9—Mean (+/- 1 SE) cover of Tamarix spp. of active removal (light grey) or no active removal (dark 

grey). West slope sites have beetles present whereas east slope sites do not. 
54 USDA Forest Service RMRS-GTR-377. 2018

Figure 10—Mean (+/- 1 SE) 

number of Tamarix (i.e., density) 

per site of Tamarix spp. on 

removal vs. non removal sites, 

in the eastern plains and west 

slope of Colorado before active 

removal in 2010 and for 2 years 

following; spring season, cut 

stump method. ANOVA test 

(F = 0.340, DF = 2, P value < 

0.004). 

sites in non-active removal sites. When we compare active removal sites with non-active 

removal at the same point in time, we see even more dramatic evidence of the impact of 

removal. By 2012, however, this difference was much greater in the east sites compared 

to the west where Diorhabda spp. is known to occur.

There was also a decrease in total density (i.e., number of individuals) of Tamarix

in the removal sites compared to non-active removal sites from the first to third year, 

also with a significant region effect, but the opposite pattern from what was observed 

with cover (fig. 10). Both regions saw a decrease in Tamarix individuals with active removal, but by 2012 there was an increase in the number of individuals in the west sites.

Differences between regions in successful control of Tamarix are multifaceted. 

Although Tamarix cover remained low in the Western Slope sites where biological 

control is ubiquitous, the Tamarix density was almost the same by the end of our 3 years 

of sampling. This suggests that even though there was new recruitment of Tamarix

seedlings after active removal in the Western Slope, defoliation by the beetles was likely 

keeping overall Tamarix cover lower than in the Eastern Slope. The lower recruitment 

of seedlings in the east is likely because there was less overbank flooding than in the 

west, which is generally associated with regeneration by seed for Tamarix; especially 

high flows were observed by our team in several areas along the Dolores River in 2011, 

but not in the eastern sites. Water availability is the primary determinant of Tamarix establishment by seed (Sher and Marshall 2003). It is also possible that in the eastern sites 

new individuals were killed with follow-up treatments or by cattle grazing or trampling, 

given that all of the eastern sites were on private land and thus subject to more intense 

use and management.
USDA Forest Service RMRS-GTR-377. 2018 55

The Response of Plant Communities to Tamarix Removal Efforts

After Tamarix removal, we observed slightly greater cover of natives and a decrease in introduced species relative to before removal, but with different magnitudes 

between the two regions (fig. 11). Overall, we observed much more dramatic positive 

changes to the understory in the western sites.

Removal method mattered as well. Other than the positive reference sites, the 

highest relative percent cover of native species was found in the cut stump sites 

(fig. 12). Relative understory native cover decreased over time only in Tamarix control 

sites and helicopter spray treatment sites, while it dramatically increased in native control, cut stump, and track-hoe treatment.

Species richness was also strongly impacted by Tamarix removal, but a different 

picture is painted whether we consider before vs. after or control vs. impact. In the east, 

the number of species decreased after removal, likely due to negative effects of the disturbance of the Tamarix removal methods (fig. 13). This is in contrast with the increase 

in richness after removal in the west. However, when we compare species richness at 

the same point in time, east and west active removal sites have very similar species 

richness, but this is low relative to positive reference sites in the east and high for those 

in the west (fig. 14). In the west, but not in the east, Tamarix-dominated stands (negative reference sites) were associated with much lower species diversity than positive 

Figure 11—Mean (+/- 1 SE) cover of understory vegetation by nativity (native “N” and introduced “I”) for east and west 

sites by active removal (“has Tamarix been removed yet”).
56 USDA Forest Service RMRS-GTR-377. 2018

reference sites. The difference between regions in the response to Tamarix removal is 

likely due to the difference in initial plant communities; most of understory vegetation were shrubs in the Western Slope whereas the Eastern Slope was dominated by 

herbaceous species. Herbaceous species are much more likely to be killed by the heavy 

machinery, use of herbicides, and trampling associated with Tamarix removal relative 

to shrubs. In particular, helicopter-applied herbicides in some of the eastern sites likely 

negatively affected the plant community.

Given that very different patterns appeared in the sites where helicopter spray 

was applied, we used another quantitative measurement, the importance value (IV), 

to investigate the community response in these sites with more detail. The importance 

value is a combination of the relative cover, density, and frequency, and as such can better explain the change of plant composition than just one parameter (Mueller-Dombois 

and Ellenberg 1974)3. Results of this analysis showed that the increase in understory 

exotics in the helicopter spray sites was primarily due to the exotic forb Bassia scoparia

(a.k.a. Kochia scoparia, burningbush). This exotic forb, which ranked third in IV before 

removal, took first place after removal, shifting the community from native dominated 

(Elymus canadensis) to exotic because of the increase in the relative cover of B. scoparia. This is an exotic species that is considered a nuisance but is not a State-listed weed. 

It is also known to have resistance to commonly used herbicides. Thus, the primary 

problem we observed with the aerial application of herbicide was its direct negative effect on the native species, which decreased significantly when this approach was used.

Figure 12—Mean (+/- 1SE) relative native understory cover over time within different removal methods. (F = 55.9, 

DF = 5,599, P < 0.0001).
USDA Forest Service RMRS-GTR-377. 2018 57

Figure 13—Average richness over time from 

before treatment to 2 years after treatment, 

by slope. 

Figure 14—Mean (+/- 1 SE) number 

of species per transect for positive 

reference sites (“native”), negative 

reference sites (“Tamarix”), and active 

removal sites in both east and west 

slope sites. (ANOVA Treatment*Slope: F 

= 19.7, DF = 2,380, P < 0.0001.)

Conclusion and Implications

In summary, active removal of Tamarix was highly successful and had significant effects on the understory that differed by region. In the western sites where the 

biocontrol beetle Diorhabda spp. was present, Tamarix cover decreased over time in 

both Tamarix active removal sites and non-active removal sites, and the former was 

associated with decreases in other exotics while increasing native species. The eastern 

sites where there was no biological control and no active removal had very high cover 

of Tamarix, but differences between these sites were not as great for the understory, due 

to the poor results seen in those sites with aerial application of herbicide. Generally, 
58 USDA Forest Service RMRS-GTR-377. 2018

Tamarix removal led to an increase in native understory vegetation except when herbicide was applied by helicopter. Spraying herbicide by helicopter was associated with 

the increased importance of Bassia scoparia, an understory exotic. These results are 

consistent with other studies in which secondary invasions were less notable following 

biocontrol than other treatment methods (González et al. 2017a; Sher et al., in review), 

particularly B. scoparia (González et al. 2017b).

It is also important to note that we observed positive reference sites to increase 

in native cover over time, whereas sites with no Tamarix removal decreased in native 

cover, and when the biocontrol was absent, increased in Tamarix cover. This suggests a 

risk of site depredation if nothing is done to reduce this exotic tree.

In conclusion, our results suggest that removal of Tamarix by commonly used 

methods is either neutral or positive for the native plant community. Although 2 years 

post removal is too short a period to reach any broad conclusions about impact, our 

results are consistent with those over longer time periods and greater geographic range 

that suggest that native cover can be promoted through Tamarix removal (González et 

al. 2017a). Although this should not be confused with a return to the mesic riparian forests associated with pre-invasion and pre-damming, it does bode well for improvement 

of these communities.

Acknowledgments

We are indebted to the following partner organizations for their support: Branson 

Trinchera Conservation District (BTCD), The Nature Conservancy (TNC), the Colorado 

Water Conservation Board (CWCB), U.S. Fish and Wildlife Service (USFWS), Natural 

Resources Conservation Service (NRCS), Denver Botanic Garden (DBG), University 

of Denver (DU), Colorado State Forest Service (CSFS), and the Bureau of Land 

Management (BLM). Many thanks to all who contributed to data collection and data 

entry for this project: D. Stahl, R. Anderson, K. Merewether, C. Slattery, B. Peters, S. 

Coles, R. Whitney, F. Aguirre, E. Jackson, and C. Ballinger. Early drafts of this paper 

were improved by feedback from M. Quigley, S. Murphy, and R. Powell. Special thanks 

are directed to all of the landowners and land management agencies, especially the 

Doherty, Wooten, and Larson families, for permitting their properties to be the locations 

for this project and for providing guidance and facilities which made the field work and 

data collection possible.

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62 USDA Forest Service RMRS-GTR-377. 2018

Chapter 5. Unintended Consequences: Tamarisk Control 

and Increasing Threats to the Southwestern Willow 

Flycatcher

Mary Anne McLeod

Introduction

It is well known that nonnative tamarisk (Tamarix parviflora, T. ramosissima, T. 

chinensis, and their hybrids; a.k.a. saltcedar) has replaced native riparian woodland 

vegetation along many streams in the arid Southwest over the last 100 years. Tamarisk 

can form extensive, dense monocultures and may alter not only the physical structure 

of the riparian woodland but also soil salinity and fire frequency (Sher 2013). There 

is significant debate, however, over whether tamarisk is the driver or a passenger of 

ecological change (Johnson 2013). The decline in the numbers and range of native 

riparian wildlife has been concurrent with the spread of tamarisk, and numerous studies show that tamarisk-dominated stands may support a lower density and/or diversity 

of wildlife than do native habitats (Bateman and Ostoja 2012; Sogge et al. 2008; 

Strudley and Dalin 2013). Consequently, tamarisk is often portrayed as the primary 

cause for declines in riparian wildlife (e.g., DeLoach et al. 2003a). Although it is now 

recognized that water use by native vegetation compared to tamarisk depends on site 

conditions (Zavaleta 2013), tamarisk was also widely blamed for water consumption 

in excess of native species (Nagler and Glenn 2013). Tamarisk control efforts, many 

of which were driven by the desire to make more water available for human use, 

began in the 1940s (Douglass et al. 2013) and continue to the present day, with improvement of wildlife habitat often cited as a goal of tamarisk removal.

Many wildlife species successfully inhabit tamarisk, however, and evidence 

mounted in the 1980s and 1990s that tamarisk provides important habitat, particularly for birds, in many southwestern riparian systems (see review in Sogge et al. 

2008), especially if a small component of native vegetation remains (van Riper et al. 

2008). Perhaps the best known, and most controversial, avian occupant of tamarisk 

habitats is the southwestern willow flycatcher (Empidonax traillii extimus; hereafter 

flycatcher), a riparian obligate songbird that was listed as endangered in 1995 by the 

U.S. Fish and Wildlife Service. In recent decades, an effective biological control for 

reducing the vigor and reproductive success of tamarisk was found, and biological 

control agents were released beginning in 2001. As a result of this biocontrol effort, 

there have been unintended consequences to the flycatcher and other riparian wildlife. 

This paper addresses the history of the biocontrol effort and the failure of the scientific community and regulatory agencies to accurately predict the impact of biocontrol 

releases.
USDA Forest Service RMRS-GTR-377. 2018 63

The Search for Biocontrol for Tamarisk

Tamarisk is difficult to control by mechanical means, as it resprouts vigorously 

following cutting or fire. Heavy equipment can be used to remove entire plants, but this 

is expensive, causes extensive soil disruption, and is often impractical in riparian areas 

that are difficult to access. There are several herbicides that are widely used for tamarisk control; however, aerial applications are non-selective, and cut stump treatments, 

while very effective, are labor intensive (DiTomaso et al. 2013). Given the difficulty 

and expense of mechanical and chemical control methods, biological control, which 

would involve using a natural predator, parasite, or pathogen to suppress the tamarisk 

population, seemed like an attractive alternative. Tamarisk was an ideal candidate for 

biocontrol; the Tamaricaceae family has no species that are native to North America, 

and the prospects were therefore good of finding a biocontrol agent that would act 

selectively on tamarisk and leave other species untouched. Successful biocontrol was 

expected to reduce the use of conventional pesticides and provide cost effective, selfsustaining, target-specific suppression for tamarisk (USDA Animal and Plant Health 

Inspection Service 2015a).

The search for biological control for tamarisk began in the late 1960s, led by 

personnel at the U.S. Department of Agriculture’s Agricultural Research Service (USDA 

ARS). Researchers gathered information on tamarisk and the insects that feed on it 

throughout its native range to identify potential biocontrol agents—insects that feed 

selectively on tamarisk and are capable of reducing the population of the host plant 

(DeLoach et al. 2003a). In 1989, Jack DeLoach of the USDA ARS and colleagues submitted a petition to the Technical Advisory Group for Introduction of Biological Control 

Agents of Weeds (TAG), a committee of the USDA Animal and Plant Health Inspection 

Service (APHIS) that is tasked with “providing guidance to researchers and recommendations to regulating agencies for or against the release of nonindigenous biological 

control agents” (USDA Animal and Plant Health Inspection Service 2015b), asking the 

TAG’s advice on whether proceeding with a biological control program on tamarisk was 

in the best national interest (DeLoach et al. 2003b). The TAG recommended that the 

program could proceed, and the search for suitable biocontrol agents continued, with the 

assistance of several overseas cooperators.

The tamarisk leaf beetle (Diorhabda spp.) was identified as one of the promising 

candidates for biocontrol. It was known to occur in high densities and completely defoliate large areas of tamarisk, and beetles were considered a pest in China in areas where 

tamarisk was planted for sand dune control (DeLoach 1994). The beetle was brought 

back to the United States where it underwent extensive host-specificity testing starting 

in 1992. Tests conducted in 1992 and 1993 showed that beetle larvae developed readily on Tamarix hosts and would feed but rarely developed into adults on shrubs in the 

Frankenia (seaheath) genus, the only plants native to the Western Hemisphere that are 

in the same order as Tamarix (Lewis et al. 2003a). Larvae failed to develop on all other 

host species that were tested (DeLoach 1994).

ARS submitted a petition to the TAG requesting approval for the field release of 

beetles in Texas and Wyoming in May 1994 (DeLoach 1994). Of the 16 TAG members 

who responded, nine recommended release with no reservations, six recommended 

release but had reservations, and one did not recommend release. The most common 
64 USDA Forest Service RMRS-GTR-377. 2018

concern was the possibility that beetles could feed on Frankenia, some species of which 

were rare. Other concerns were the impact beetles could have on athel (T. aphylla, a 

tree form of tamarisk that is far less invasive than the other Tamarix species), which 

was used as a shade tree in the southwestern United States and Mexico; whether native 

vegetation would replace tamarisk following biocontrol; and whether there was a means 

to control Diorhabda, if needed. A letter from Jack DeLoach to APHIS on December 

22, 1994, clarified that larvae that developed into adults on Frankenia were unable to 

reproduce; this was later confirmed by further host testing (Lewis et al. 2003a). The 

letter cited the relative rarity of athel among ornamental shade trees as evidence of the 

limited potential effects to athel.1 DeLoach conceded in his letter that “Replacement by 

native vegetation after control ... is circumstantial and not well supported. It is based 

mostly on biocontrol of other weed species, where native vegetation, desirable range 

plants, etc., came back strongly after the weed was controlled.” As far as beetle control 

was concerned, DeLoach cited information from China that local beetle populations 

could be controlled with insecticides. The letter was sufficient to relieve the concerns 

of the TAG members, and the TAG chairman recommended to APHIS on June 1, 1995, 

that Diorhabda be approved for release into North America.

Tamarisk leaf beetles feed on the foliage of tamarisk as adults and in each of three 

larval stages. Beetles overwinter in the leaf litter as adults and emerge in the spring in 

response to warming temperatures. They aggregate on tamarisk plants to feed, mate, and 

lay eggs. After progressing through the larval stages, beetles descend to the leaf litter 

to pupate. The entire life cycle takes 30 to 40 days (Lewis et al. 2003b), and beetles 

typically go through multiple generations during a growing season. When adult beetles 

emerge from the leaf litter, they either become reproductively active or are triggered to 

go into diapause, an overwintering state of suppressed development, by shortening of 

the photoperiod in late summer (Bean et al. 2007a; Lewis et al. 2003b). If local food 

resources have been depleted by the previous generation of beetles, adults will disperse, 

sometimes in large flights, in search of green foliage. Adults that are destined for 

diapause feed and then descend into the leaf litter instead of becoming reproductively 

active (Bean et al. 2013).

Tamarisk leaf beetles are sensitive to chemicals released by tamarisk and also to 

chemicals released by other Diorhabda individuals (Bean et al. 2013). This sensitivity 

allows beetles to occur in large aggregations that can exceed 1,000 beetles per plant 

(Bean et al. 2013; Jashenko n.d.), resulting in rapid and complete defoliation. The defoliation period can last for several weeks, but the tamarisk then typically puts on new 

leaves. An individual plant can be defoliated multiple times within a growing season. 

The response of tamarisk to multiple defoliation events is variable and is likely influenced by numerous factors, including age of the plant, access by the plant to resources 

such as water, seasonal timing of defoliation, and plant genetics (Bean et al. 2013). A 

single defoliation event can result in mortality, but more typically several defoliation 

events are required, if the plant is killed at all. When plants do refoliate, they often 

exhibit reduced vigor, with dieback of terminal branches, reduced foliage volume, and 

1 Further research suggested that beetles would not affect athel as strongly as they affected other Tamarix species, although 

the researchers conceded the possibility of “transient but substantial damage” to athel (Moran et al. 2009). Damage 

to athel in northern Mexico was described as “conspicuous” (Estrada-Muñoz and Sánchez-Peña 2014), and treating 

individual athel trees with insecticide was recommended for beetle control (Muegge 2010).
USDA Forest Service RMRS-GTR-377. 2018 65

reduced flowering. The response of other vegetation to tamarisk dieback and mortality 

is also variable and is influenced by soil condition, hydrological regime, and local seed 

sources. 

Conflict With the Southwestern Willow Flycatcher

By the time beetles had been recommended for release by the TAG, another 

concern had surfaced. The southwestern willow flycatcher was listed as endangered 

on February 27, 1995 (60 FR 10694-10715). The flycatcher breeds in Arizona, New 

Mexico, southern California, southern Nevada, southern Utah, and western Texas, placing its nests in dense, wet thickets of trees and shrubs approximately 4–7 m or more 

in height. The southwestern willow flycatcher was historically a common species in 

riparian areas throughout its range, but flycatcher numbers dwindled as southwestern wetlands and riparian habitats, particularly those vegetated by cottonwood and 

willow, suffered large-scale losses during the 1900s as the result of dams, diversions, 

livestock grazing, increase in agriculture, urbanization, and wood cutting. By the time 

the flycatcher was listed, the range-wide population was estimated at around 500 pairs 

(60 FR 10694-10715). The decline in flycatcher populations coincided with the spread 

of tamarisk, which proliferated with the modifications in the natural hydrograph caused 

by dams and diversions. Tamarisk was regarded by some researchers as providing poor 

habitat in comparison to native vegetation for various bird species because of reduced 

structural diversity, changes in the arthropod community (Carothers and Brown 1991), 

and a hotter microclimate (Hunter et al. 1987). At the time of listing, flycatchers were 

known to nest in thickets dominated by tamarisk, but it was unclear whether the longterm reproductive success of flycatchers nesting in native vegetation differed from the 

success of those nesting in tamarisk.

Because of the potential effect of beetle-caused defoliation on flycatchers, consultation with the U.S. Fish and Wildlife Service (USFWS) was required prior to release 

of the beetles. ARS and APHIS submitted a draft Biological Assessment (BA) to the 

USFWS in October 1997 (DeLoach and Tracy 1997). The action proposed in the draft 

BA was to release both the leaf beetle and a mealybug, Trabutina mannipara, at sites 

across seven States. All proposed release sites were at least 100 miles from areas where 

flycatchers were known to nest in tamarisk. The BA considered direct and indirect effects of biological control agents on flycatchers not only at the release sites themselves 

but also across the region, after the control agents had resulted in an estimated 75–85 

percent reduction in the density and cover of tamarisk.

The analysis of effects in the BA rested on several assumptions: (1) biocontrol 

agents would not spread more than 2–4 miles each year; (2) the decrease in the density 

of tamarisk would be gradual and would be accompanied by a “consequent and concurrent increase in the native plant community”; (3) biocontrol agents would provide food 

for flycatchers; and (4) tamarisk is “only partially suitable as habitat for flycatchers” 

and was “a major factor in [the flycatcher’s] extirpation from ... the lower Colorado and 

the lower Gila rivers.” In particular, the BA proposed that tamarisk provides unsuitable 

habitat for flycatchers at lower elevations because tamarisk stands are “intrinsically 

hotter” than cottonwood/willow vegetation, resulting in the exposure of flycatchers to 

temperatures that are lethal for eggs and nestlings. It then followed that, if biocontrol 
66 USDA Forest Service RMRS-GTR-377. 2018

resulted in the replacement of tamarisk with native vegetation, the suitability of formerly tamarisk-dominated areas for nesting flycatchers would increase. The BA concluded 

that biological control of tamarisk “is not likely to adversely affect the southwestern 

willow flycatcher.”

The authors of the BA overlooked a few key factors about the effects of the beetles 

on tamarisk. Although mortality of tamarisk might be gradual, defoliation by tamarisk 

leaf beetles was known to be sudden and complete. If, as the project proponents maintained, exposure of flycatcher nests to lethal temperatures was a cause of extirpation of 

flycatchers from some areas, surely the removal of shade during the height of flycatcher 

breeding season should have been a concern.

In addition, there was no evidence, at the time the BA was written, that native 

vegetation would return to areas where tamarisk had been controlled. The BA did not 

completely address the role that changes in hydrology (e.g., the construction of Hoover 

Dam) likely played both in the invasion of tamarisk and in the reduction of habitat 

suitability for willow flycatchers via stream channelization and the reduction in spring 

floods that create the dense thickets of young vegetation preferred by flycatchers.

Despite assurances by beetle proponents, some scientists gave unequivocal warnings against beetle releases (see Appendix IV in DeLoach and Tracy 1997). In response 

to concerns voiced within and outside of the USFWS and APHIS, a meeting was called 

in June 1998 with the USFWS, ARS, and other Federal agencies to discuss the BA. 

As a result of these concerns, ARS and APHIS submitted a revised research 

proposal to the USFWS in August 1998 (DeLoach and Gould 1998). This proposal 

requested the release of tamarisk beetles and mealybugs at 13 sites in seven western 

States, with one of the criteria for site selection being a distance of at least 200 miles 

from areas where flycatchers nest in tamarisk. The proposal specified two phases—

research and implementation—of the biocontrol program. The research phase included 

both cage and field releases and was intended in part to evaluate whether remedial 

actions, such as manual revegetation, would be needed prior to the arrival of beetles 

in areas occupied by flycatchers. The proposal stated that biocontrol agents would be 

contained within the defined release sites; some of these release sites, however, encompassed over 100 miles of river. The implementation phase would entail release of 

biocontrol agents at additional sites, with the goal of widespread control of tamarisk, but 

would still exclude releases in areas where flycatchers nest in tamarisk.

The proposal acknowledged, however, that the control insects would eventually 

disperse to areas occupied by flycatchers, though this was projected to take 10 to 20 

years. This estimate of dispersal speed (200 miles in as little as 10 years, or an average 

of 20 miles in a year) was quite different from the 2–4 miles per year stated in the 1997 

BA, but there was no direct acknowledgment in the proposal that the dispersal estimate 

had changed, and subsequent analysis by the USFWS did not consider this faster 

dispersal rate. The proposal expressly stated, “Unprecedented safeguards and precautions are herein proposed to insure that biological control does not adversely affect the 

[flycatcher], especially through the process of reducing its potential breeding habitat 

(which presently includes saltcedar) before the recovery of its native habitat (willows, 

cottonwoods and other native trees and shrubs) can occur.” 

The proposal refers to the BA for these safeguards and precautions, and the BA 

contains only the very general statement that in the “very improbable” event that beetles 
USDA Forest Service RMRS-GTR-377. 2018 67

needed to be controlled, “...the areas of the most dense and most damaging control 

agents populations will be treated by ground or aerial applications of appropriate insecticides at sub-lethal or borderline lethal rates in order to slow the reproduction and 

spread, but without exterminating all control agents at the release site.” There was no 

consideration of the possibility that by the time it would become obvious that control 

was progressing too rapidly, the beetles would occupy too large an area for insecticide 

to be a reasonable solution.

The Flycatcher Recovery Team reviewed the revised research proposal and 

submitted a letter to the Southwest Regional Director of the USFWS, expressing its 

concerns that tamarisk might not be replaced by native vegetation with equal function 

with respect to the needs of the flycatcher, particularly in areas with site conditions such 

as altered flooding regimes, high salinity, and grazing. There was also concern that the 

biocontrol agents might escape the proposed release areas, either by natural dispersal or 

by transport by humans (USFWS 2014). 

Despite the concerns of the Flycatcher Recovery Team and the presence of flycatchers nesting in tamarisk along the Rio Grande (USFWS 2002), the USFWS issued 

concurrence on December 28, 1998, that the proposed releases at all 13 sites would not 

adversely affect the flycatcher. It cited safeguards such as the 200-mile distance from areas where flycatchers nested in tamarisk. The concurrence was rescinded the following 

April, based on “new information” regarding flycatchers nesting in tamarisk along the 

Rio Grande. An amended concurrence was issued on June 3, 1999, following a meeting 

between USFWS and USDA personnel and agreement from the project proponents that 

all sites along the Rio Grande would be removed from consideration. 

The concurrence required a separate consultation before the implementation phase 

of the project could be started, and it also required separate consultation for the addition 

of any new sites. Despite the prediction in the project proposal that beetles could spread 

as much as 20 miles per year, the concurrence cited “convincing argument[s]” presented 

by USDA personnel at the meeting that the geographic isolation of the remaining release 

sites would prevent the control agents from reaching areas where flycatchers nested 

in tamarisk, given that “research ... indicates [the control agents] should move on the 

order of tens of feet per year.”

A final draft Environmental Assessment (EA) on the biological control of tamarisk 

was released by APHIS in July 1999. The EA addressed the release only of the tamarisk 

leaf beetle, Diorhabda elongata Brulle, and did not include the mealybug. The proposed 

action included the release of beetles at 10 sites, all of which were at least 185 miles 

from flycatcher nesting areas, and stated that the spread of beetles would be slowed 

through chemical or mechanical means if the beetles appeared to “consume saltcedar 

too rapidly.” The EA did not predict a rate of spread for these particular beetles but 

implied that the rate would be slow, given that other chrysomelid beetles “appeared to 

spread relatively slowly at a maximum of several tens of meters per year,” and stated 

that invasion of flycatcher nesting areas by beetles and the death of tamarisk faster than 

native plants could regenerate was “highly improbable.” A Finding of No Significant 

Impact (FONSI) was signed on July 7, 1999, and APHIS immediately began issuing 

permits for the establishment of field cages.
68 USDA Forest Service RMRS-GTR-377. 2018

Beetle Releases

Northern Beetles

Field cages were established as soon as permits were obtained, but because permits were not issued until mid-July, research at the field cages during 1999 was largely 

limited to one generation of beetles. Despite having data only from a single generation 

of beetles in a single year, a report issued later that year (Gould 1999) concluded that 

“the data ... indicate that D. elongata will require at least several years to reduce even 

local saltcedar stands to a significant degree.” A second summer of cage studies was 

completed in 2000, and beetle populations at several of the sites increased to very high 

numbers, resulting in the consumption of all foliage in the cages and causing complete 

mortality of some large plants (DeLoach et al. 2003b).

In the spring of 2001, APHIS issued permits for the open release of beetles, and 

beetles were released in May at sites in Texas, Colorado, Wyoming, Utah, Nevada, 

and California. Little defoliation was seen during the summer of 2001, but the summer 

of 2002 produced “spectacular defoliation” at some sites (DeLoach et al. 2003b). By 

September of 2002, beetles had dispersed no farther than approximately 100 m from the 

release points.

The beetles that were released into cages in 1999 and then into the field in 2001 

all originated either from Fukang, China, or Chilik, Kazakhstan (classified at the time 

as Diorhabda elongata deserticola and later reclassified as D. carinulata, the northern 

tamarisk beetle; Tracy and Robbins 2009). The cage and field tests revealed that beetles 

originating from these northerly latitudes did not overwinter successfully at southerly 

latitudes (DeLoach et al. 2003b). Laboratory experiments and field observations on the 

northern tamarisk beetle showed that this ecotype entered diapause at a day length of 

≈ 14 h 39 minutes; south of 36° 20’ N (approximately 50 miles south of the northern 

borders of Arizona and New Mexico), the longest day is shorter than this critical day 

length. Researchers therefore concluded that this northern ecotype was unlikely to succeed as a biocontrol agent south of 38° N (approximately 70 miles north of the northern 

borders of Arizona and New Mexico) because premature diapause would increase mortality (Bean et al. 2007b).

Other Beetle Species

Beetle proponents began testing other ecotypes, which were also later reclassified as separate species, originating from Uzbekistan (D. carinata, the larger tamarisk 

beetle), Crete (D. elongata, the Mediterranean tamarisk beetle), and Tunisia (D. sublineata, the subtropical tamarisk beetle) (Tracy and Robbins 2009) for control of tamarisk 

at latitudes below 38° N. By the end of the summer of 2002, all four beetle species had 

been tested in cages, and a request was submitted on February 14, 2003, to the USFWS 

for the release of beetles of all four subspecies at the 10 existing sites as well as 20 

new sites, many of which were in Texas and New Mexico and included sites on the Rio 

Grande where flycatchers were known to breed (DeLoach et al. 2003b). Beetle releases 

at these flycatcher sites were requested as a demonstration that beetles would have no 

adverse effects on flycatchers that nested in areas with abundant willows. 
USDA Forest Service RMRS-GTR-377. 2018 69

Subsequent discussion between ARS and the USFWS resulted in all sites along the 

Rio Grande that were within 200 miles of occupied flycatcher habitat being removed 

from consideration, and the USFWS issued a letter on June 13, 2003, concurring that 

release of beetles at the additional sites “may affect, but is not likely to affect” the 

flycatcher. This determination relied on distance and geographic barriers (i.e., expanses 

without tamarisk) to be “effective as a means of keeping the control agents from sites 

on the Lower Colorado River in Arizona and the Rio Grande in New Mexico where 

flycatchers are nesting in saltcedar.” Concurrence was also given by the USFWS on 

July 23, 2003, for the release of multiple species of beetle near Kingsville, Texas, and 

at several sites along the Pecos River. Concurrence for release at sites in western Texas 

followed on September 1, 2004. Between 2003 and 2009, beetles of all four species 

were released at about 70 sites in Texas, and the larger, subtropical, and Mediterranean 

species became established at various sites (DeLoach et al. 2011). Beetles were first 

released on the Rio Grande in Texas in 2007, and subtropical beetles became established 

at multiple sites along the Rio Grande in 2009.

Implementation Phase of Biocontrol Program

While additional release sites were being approved by the USFWS, USDA personnel were also working toward the implementation phase of the biocontrol program. 

The implementation phase considered only the northern tamarisk beetle, D. carinulata. 

Under the implementation phase, beetles would be established at “nursery” sites in up to 

13 western and midwestern States, north of 38° N latitude. Beetles would then be available for distribution anywhere in those States. Utah was excluded from the plan since 

beetles had already been distributed there. 

A BA for the implementation phase was released in March 2005 and an EA followed 

in June. The effects analysis in both documents determined that there would be no effect 

to flycatchers, and this determination relied on critical day length for D. carinulata precluding beetles from establishing populations in areas south of 38° N. The EA did concede 

that beetles from the northern ecotype could eventually adapt to conditions below 38° but 

stated that this adaptation would not be rapid because beetles were documented to disperse 

only 1.5 miles in 3 years. Later in the same document, however, dispersal was noted as 

being over an order or magnitude larger than this, 50 miles in 4 years.

In the concurrence letter, the USFWS indicated that APHIS provided additional 

information and clarification to the USFWS before the concurrence was issued. The 

USFWS concluded that D. carinulata is “capable of occurring between 36 and 38 

degrees north latitude in substantial numbers at some locations” and also acknowledged 

that the beetles “are expected to eventually adapt to diapause at lower latitudes” but 

still maintained that beetles were expected to have “little success in reducing saltcedar 

stands in areas at or south of 38° north latitude.” The concurrence letter also recognized 

that suppression of tamarisk without active management to restore native vegetation 

might result in riparian habitat of degraded quality, thereby reducing habitat quality 

for flycatchers. Despite these apparent misgivings, the USFWS concurred on July 11, 

2005, that the proposed action was not likely to adversely affect the flycatcher, and the 

FONSI, which predated the concurrence, was issued for the EA in June 2005. 
70 USDA Forest Service RMRS-GTR-377. 2018

Beetles Spread Into Flycatcher Habitat

Virgin, Muddy, and Lower Colorado Rivers

The 2005 EA for the implementation phase of the beetle program acknowledged 

that the eventual outcome of the no action alternative and the proposed action alternative were the same; the only difference was the rapidity with which beetles would 

spread. As was soon to be demonstrated, however, beetles are capable of rapid spread 

even without human assistance. In 2004, beetles were released along the Colorado River 

near Moab, Utah. This release was not done under the USDA program but rather by the 

Grand County weed management department, which had collected beetles (D. carinulata, from the Chilik, Kazakhstan strain) from Delta, Utah, one of the original 10 test 

sites, and re-distributed them near Moab. The first major defoliation along the Colorado 

River was recorded in 2006 and encompassed around 990 acres over approximately 20 

miles of river. By the end of the following year, around 10,000 acres were defoliated 

over approximately 70 miles of the Colorado River, and beetles had spread into the 

Green and Dolores Rivers (Tamarisk Coalition, n.d.). In subsequent years, these beetles 

continued to spread in the upper Colorado River Basin.

In 2006, another D. carinulata release occurred, this time by the city of St. George, 

Utah. St. George lies along the Virgin River just above 37° N latitude and has a small 

population of nesting flycatchers. Although beetle researchers had claimed that D. carinulata was unlikely to do well below 38° N, widespread defoliation around St. George 

occurred beginning in August 2008 (Dobbs et al. 2012). As the beetles progressed 

southward, it became abundantly clear that the assumption that northern beetles would 

not succeed as a biocontrol agent below 38° N was wrong. Further experiments in 2007 

and 2008 on D. carinulata showed that critical day length had decreased over the previous 5 years, with the magnitude of the decrease being inversely proportional to latitude 

(Bean et al. 2012).

The area around St. George was again defoliated in 2009, this time beginning in 

June, and by the end of that summer, complete defoliation had spread over 25 miles 

downstream of St. George, to a point between Littlefield, Arizona, and Mesquite, 

Nevada. Beetles continued to spread downstream on the Virgin River, traveling another 

20 miles and encompassing another flycatcher breeding site in 2010, reaching the Gold 

Butte area by the end of the summer, and extending another 20 miles to reach a third 

flycatcher breeding site and encompass the entire lower Virgin River to the Overton 

Arm of Lake Mead (approximately 36.4° N) by the end of 2011. By the end of 2012, 

beetles occurred all along the Muddy River, including at another flycatcher site, and had 

reached the lower end of Lake Mohave. By the end of 2013 they were at Big Bend State 

Park, south of Laughlin, Nevada, at 35.1° N, having dispersed a straight-line distance of 

150 miles in the 7 years since their release in St. George (fig. 15). This was an average 

dispersal distance of slightly over 20 miles per year, exactly what had been stated in the 

1998 project proposal as the maximum likely dispersal rate.

Over the next 2 years, beetles spread very little south of Big Bend State Park. In 

July 2015, extensive defoliation was apparent at Big Bend State Park (M.A. McLeod, 

Biologist, SWCA Environmental Consultants, personal observation, Big Bend State 

Park, July 23, 2015), and a few individual beetles but no defoliation were observed 
USDA Forest Service RMRS-GTR-377. 2018 71

approximately 5 miles south of Big Bend (Tom Dudley, University of California Santa 

Barbara, personal communication, August 7, 2015). Some people speculated that the 

northern beetles had finally reached the latitude where southerly dispersal would be 

limited by critical day length. However, by July 2016, beetles and defoliation were 

apparent at Topock Marsh, on both sides of Lake Havasu, and at the mouth of the Bill 

Williams River, over 60 miles from Big Bend and within a quarter mile of nesting flycatchers along the Bill Williams River (M.A. McLeod, Biologist, SWCA Environmental 

Consultants, personal observation, Lower Colorado River, July 21 and 22, 2016).

C A L I F O R N I A

N E V A D A

A R I Z O N A

Las Vegas

Kingman

Virgin River

St. George

L a k e M e a d

Pahranagat Wash

L a k e M o h a v e

T o p o c k M a r s h

L a k e H a v a s u

Colorado River

Bill Williams River

2008

2010 2009

2011

2012

2013

2016

Littlefield

Muddy River

Laughlin

Figure 15—Spread of tamarisk 

beetles along the Virgin 

and Lower Colorado 

Rivers in 2008 to 2016 

(map by Glenn Dunno, 

SWCA Environmental 

Consultants). 
72 USDA Forest Service RMRS-GTR-377. 2018

Rio Grande

While D. carinulata were spreading along the Virgin River and downstream along 

the Lower Colorado River, they were also spreading elsewhere. By the end of 2014, D. 

carinulata were found throughout much of the upper Colorado River Basin, including 

the San Juan River, as well as through Grand Canyon and the Little Colorado River as 

far east as Holbrook, Arizona. D. carinulata were also defoliating tamarisk along the 

Rio Grande around Albuquerque and as far south as Socorro, New Mexico, near 34° N 

(Tamarisk Coalition 2014). By the end of 2015, northern beetles had moved southward 

on the Rio Grande almost to Elephant Butte Reservoir. Subtropical beetles expanded 

northward along the Rio Grande from Texas at the same time as the northern beetles 

advanced southward. Subtropical beetles were found north of Las Cruces at Rincon, 

New Mexico, by fall of 2014; in August of 2015 beetles and defoliation were noted at 

Caballo Reservoir, about 30 miles south of Elephant Butte (James Tracy, Texas A&M 

University, College Station, Texas, personal communication, August 6, 2015).

Beetles, presumed to be of the subtropical species but awaiting species verification at the time of this writing, arrived at Elephant Butte during the summer of 2016 

and began defoliating tamarisk close to areas with large numbers of flycatchers nesting 

in tamarisk (Dave Moore, U.S. Department of the Interior, Bureau of Reclamation, 

Denver, Colorado, personal communication, July 2, 2016). Nesting flycatchers are found 

along the Rio Grande from north of Albuquerque south to Radium Springs near Las 

Cruces. The Elephant Butte area hosts one of the largest populations of nesting flycatchers, with over 300 territories documented in 2014 (Moore 2015). Nesting habitat along 

the Middle Rio Grande was dominated by native vegetation in 2002, but native habitat 

declined in subsequent years, partly as the result of drought, and by 2014 less than 40 

percent of flycatcher nest sites were dominated by native vegetation (Moore 2015).

The subtropical beetle had been projected to reach Elephant Butte Reservoir by 

fall of 2014 and the middle Gila River in Arizona by the spring of 2017 (Tracy 2014). 

Given that the arrival of beetles at Elephant Butte was later than initially thought, 

beetles will likely not arrive on the Gila River as soon as was anticipated. The middle 

Gila River also hosts a large flycatcher population, and almost all flycatcher breeding 

sites are dominated by tamarisk (Graber et al. 2012; Heather English, Salt River Project, 

Phoenix, Arizona, personal communication, April 30, 2013).

Cessation of Biocontrol Releases

The Center for Biological Diversity (CBD) and Maricopa Audubon Society filed 

suit against the USFWS and APHIS on March 27, 2009, seeking reinitiation of consultation between APHIS and the USFWS, after beetles defoliated flycatcher habitat around 

St. George and research showed that northern beetles were adapting to more southerly 

latitudes. Consultation was reinitiated later that year. In May 2010, APHIS submitted 

a new BA to the USFWS, asking for concurrence that cessation of the beetle release 

program would have no adverse impacts to the flycatcher. An official moratorium on 

the beetle program, cancelling all existing permits for release and interstate transport of 

all Diorhabda species and discontinuing issuance of new permits, was announced by 

APHIS on June 15, 2010. However, beetles were already widely established by then, 

and shutting down the release program had no effect on preventing the spread of the 
USDA Forest Service RMRS-GTR-377. 2018 73

beetle. CBD and Maricopa Audubon Society filed a second suit in 2013, seeking, in 

part, a “mitigation plan to address impacts of the beetle populations on flycatchers and 

their critical habitat.”

Response by Flycatchers to Tamarisk Defoliation

St. George, Utah

Breeding flycatchers and beetle defoliation first overlapped in St. George, Utah, 

in late July of 2008. This was, coincidentally, also the year that the Utah Division of 

Wildlife Resources began intensive monitoring of nesting flycatchers in St. George 

after doing general presence/absence surveys since 2001. Flycatchers along the Virgin 

River typically arrive on their territories beginning in early May, with late arrivals 

coming in the middle of June. The average start of incubation is in the middle of June, 

and the latest nests fledge by the middle of August (Bureau of Reclamation, n.d.; Sogge 

et al. 2010). Therefore, defoliation that begins in late July has little effect on flycatchers. Eight breeding pairs and 10 nesting attempts were documented in St. George in 

2008; seven of the 10 nests were built in areas dominated by tamarisk (Edwards and 

Woodhouse 2015). The eight pairs produced 17 young (fecundity, or young produced 

per female, was 2.1; fig. 16) (Fridell et al. 2009).

The following year, defoliation was noted beginning in early June, after territories 

had been established and when nesting was already underway, and a second defoliation event occurred in late July to August. The Utah Division of Wildlife Resources 

documented 10 breeding pairs and 18 nests, three of which were abandoned before eggs 

were laid. As was the case in 2008, the majority of nests (11 of 15 nests with flycatcher 

eggs) were placed in areas dominated by tamarisk (Edwards and Woodhouse 2015). 

Defoliation of the tamarisk was complete, leaving nests without shade or concealment 

(fig. 17). Productivity was markedly lower than in 2008, with a total of two young 

produced (fecundity = 0.2) (Fridell et al. 2009; fig. 16). Primary causes of nest failure 

were depredation and failure of the eggs to hatch after at least 18 days of incubation. 

Six nests (46 percent of nests that failed after eggs were laid) failed to hatch, indicating 

that embryos died in the egg (Dobbs et al. 2012). Failure to hatch is not typically a common reason for a nest to fail, accounting for less than 9 percent of failures on the lower 

Virgin River in 2003 to 2010 (McLeod and Pellegrini 2013). The high rate of addled 

nests in 2009 could have been the result of eggs at unshaded nests being exposed to 

lethal temperatures.

In 2010, defoliation was again noted in St. George beginning in early June, with 

a second defoliation in late July to August (Dobbs et al. 2012). Nine breeding pairs 

of flycatchers were detected in 2010. The marked difference between 2010 and prior 

years was where the flycatchers established territories. In previous years, flycatchers 

had nested primarily in tamarisk-dominated sites; in 2010, 16 of 20 nests were built in 

willow-dominated areas that had been previously either unoccupied or sparsely occupied by flycatchers, including one site where willow planting had occurred following a 

large flood in 2005. Fecundity in 2010 rose to 1.3 young per female (Dobbs et al. 2012). 

Tamarisk around St. George has been defoliated each year since 2008, although the 

timing and number of defoliation events has varied among years. Despite annual defoliation, the tamarisk shows little sign of mortality or reduced vigor (Hultine et al. 2015; 
74 USDA Forest Service RMRS-GTR-377. 2018

Figure 16—Number of female southwestern willow flycatchers and annual fecundity (young produced per female) documented 

at St. George, Utah, and Mormon Mesa, Nevada, 2008 to 2015. 

Figure 17—Female southwestern willow 

flycatcher panting as she shades her nest in 

a defoliated tamarisk stand, St. George, Utah 

(photo by Pam Wheeler, Utah Division of 

Wildlife Resources).

0

0.5

1

1.5

2

2.5

3

0

5

10

15

2008 2009 2010 2011 2012 2013 2014 2015

fecundity (young per female)

# females

Mormon Mesa # females

St. George # females

Mormon Mesa fecundity

St. George fecundity
USDA Forest Service RMRS-GTR-377. 2018 75

M.A. McLeod, Biologist, SWCA Environmental Consultants, personal observation, St. 

George, Utah). Flycatchers around St. George continued occupying native-dominated 

areas in 2011 to 2015, though flycatchers again occupied tamarisk-dominated areas in 

2014 and 2015 (Edwards and Woodhouse 2015). The total number of flycatcher pairs 

fluctuated between 7 and 12, and fecundity varied between 1.4 and 3.0 young per female (fig. 16; Dobbs and Edwards 2012; Dobbs et al. 2012; Edwards and Dobbs 2013; 

Edwards and Woodhouse 2014, 2015).

Flycatchers typically demonstrate a high degree of site fidelity, with half of all 

adults that are detected in multiple years returning to nest within 40 m of the place 

where they nested the previous year (McLeod and Pellegrini 2013). However, site fidelity is strongly influenced by reproductive success, with flycatchers that fail to produce 

offspring being far more likely to attempt breeding at a different site in the following 

year (McLeod and Pellegrini 2013; Paxton et al. 2007). This local plasticity in site selection, as demonstrated by the flycatchers in St. George, is what should be expected from 

a species that was adapted to the southwestern riparian system, which was characterized 

by scouring floods and a constantly shifting mosaic of vegetation.

Mormon Mesa, Nevada

When beetles progressed downstream on the Virgin River in 2009 to 2011, they 

affected the two flycatcher breeding sites on the lower Virgin River. Flycatchers along 

the Virgin River in Arizona and Nevada have been monitored annually since 1997 under 

a contract funded by the Bureau of Reclamation. Two areas, one in the vicinity of the 

town of Mesquite, Nevada, and the other approximately 19 miles downstream, across 

Mormon Mesa from Overton, Nevada, have been consistently occupied by breeding 

flycatchers since monitoring began (McLeod and Pellegrini 2015). Although flycatcher 

occupancy and nest success at Mesquite declined markedly from 2010 to 2013 (McLeod 

and Pellegrini 2014), this was likely a response to the site being dry for extended periods in 2011–2013 and the resulting decline in vegetation health.

The flycatcher breeding site at Mormon Mesa contained patches of coyote willow and a few emergent Goodding’s willows (Salix gooddingii) surrounded by a sea 

of tamarisk. Defoliation was first observed at Mormon Mesa in mid-July 2011, and the 

breeding area was fully defoliated by early August. In 2012, widespread defoliation 

was observed in late May, and a second defoliation event occurred at the beginning of 

August. No defoliation events occurred during the flycatcher breeding season in 2013, 

but defoliation from the previous 2 years had resulted in 84 percent dieback at Mormon 

Mesa by 2013 (Hultine et al. 2015). This dramatic change in the vegetation is readily 

apparent on satellite imagery (fig. 18).

As was the case in St. George, the first defoliation event at Mormon Mesa occurred 

late in the breeding season, after most flycatcher nesting activity was over. The number 

of female flycatchers (13) and fecundity (1.3 young per female) documented in 2011 

were typical of previous years (fig. 16; McLeod and Pellegrini 2013). In the following 

year, defoliation occurred as flycatchers were establishing territories and choosing nest 

sites. Fourteen female flycatchers nested at Mormon Mesa in 2012, and every nest found 

was placed in or on the edge of a willow patch. 
76 USDA Forest Service RMRS-GTR-377. 2018

Despite nests being placed in willow patches, fecundity in 2012 was 0.6, just half 

that recorded in the prior year (McLeod and Pellegrini 2013). Nest desertion during 

the laying phase, without any sign of cowbird parasitism or partial depredation of the 

clutch, accounted for half (4 of 8) of the nest failures; in 1997 to 2010, only two other 

instances of this were observed on the lower Virgin River, accounting for 1 percent 

of nest failures (Bureau of Reclamation, n.d.). Nest desertion may be indicative of 

poor habitat quality; flycatchers in Arizona were observed to forego nesting altogether 

under severe drought conditions (Ellis et al. 2008). In 2013, only seven females nested 

at Mormon Mesa, producing one fledgling among them (fecundity = 0.14 young per 

female; fig. 16; McLeod and Pellegrini 2014). Four nests were abandoned before eggs 

were laid, and two of the four nests (50 percent) that failed after eggs were laid had 

addled clutches.

Flycatchers at Mormon Mesa were not monitored in 2014 because all Federal 

agency personnel and their contractors were barred from visiting the lower Virgin River 

Figure 18—The Mormon Mesa area of the Virgin River in August 2010 before tamarisk beetles arrived (left) and in May 

2013, after 2 years of defoliation (right). The 2013 image was not taken during a defoliation event and depicts the 

mortality and dieback of tamarisk caused by the prior 2 years of defoliation. The areas in the 2013 image that appear 

grayish-green are vegetated by arrowweed (Pluchea sericea), and the bright green areas are coyote willow (Salix exigua) 

or Goodding’s willow (S. gooddingii). Yellow outline shows the area surveyed for flycatchers (base photo by Google 

Maps, outlines by SWCA Environmental Consultants).
USDA Forest Service RMRS-GTR-377. 2018 77

because of safety concerns following the Bureau of Land Management’s attempt to 

remove Cliven Bundy’s trespass cattle. The Nevada Department of Wildlife (NDOW) 

completed surveys and intermittent monitoring for flycatchers at Mormon Mesa in 2015. 

Only two breeding male flycatchers were detected, each of which was confirmed to have 

one mate and suspected to have two (NDOW, n.d.). Two of the females were confirmed 

to have produced three fledglings between them, for a minimum fecundity of 0.8 young 

per female (fig. 16).

It is unlikely that a substantial number of the flycatchers at Mormon Mesa moved 

to other locations, where they went undetected, in the years following 2012. Estimated 

annual survival of adult flycatchers along the Virgin River in 1997–2012 was 61 percent (McLeod and Pellegrini 2013). In 2012, 20 adult flycatchers were individually 

identified, via unique color-bands, at Mormon Mesa. Thirteen (65 percent, essentially 

what would be expected given annual mortality) of those flycatchers were detected 

again in 2013, and all 13 were detected at Mormon Mesa; none were detected at other 

flycatcher sites despite all known flycatcher breeding areas in southern Nevada and 

western Arizona being monitored as part of various projects. In addition, all of the adult 

flycatchers detected at Mormon Mesa in 2013 had been there the year before; i.e., for 

the first time in 17 years of monitoring, no new adults entered the population at Mormon 

Mesa (McLeod and Pellegrini 2014; Bureau of Reclamation, n.d.).

St. George vs. Mormon Mesa

Flycatchers at both St. George and Mormon Mesa had a sharp decline in productivity in the first year when defoliation significantly overlapped the flycatcher breeding 

season. The response after that first year differed dramatically between the two areas, 

however, and the two populations took very different trajectories. Flycatchers in St. 

George moved into native-dominated stands in subsequent years, the total number of 

breeding flycatchers remained the same, and productivity recovered from the low observed in the first year of defoliation; flycatchers at Mormon Mesa continued attempting 

to breed in the same stands but the number of breeding flycatchers dropped sharply and 

productivity remained very low through the second year, despite no defoliation events 

occurring during that breeding season.

There are two key differences between St. George and Mormon Mesa that likely 

contributed to the differing trajectories of the flycatcher population. One is the response 

of tamarisk to beetle defoliation. Tamarisk around St. George has experienced little 

mortality or dieback as a result of repeated defoliation; unless a defoliation event is occurring, the tamarisk still provides shade and concealment. In the Mormon Mesa area, in 

contrast, two seasons of defoliation resulted in widespread mortality and dieback. 

The other striking difference is the overall species composition of the vegetation 

in the riparian zone. St. George is on the upper Virgin River, where native trees are 

common. Prior to 2009, St. George had stands of native trees that looked like suitable 

flycatcher habitat and had been surveyed, but where few or no resident flycatchers were found (Rob Dobbs, Utah Division of Wildlife Resources, Hurricane, Utah, 

personal communication). Flycatchers moved into these existing native sites in 2010, 

following defoliation and widespread nest failure at tamarisk-dominated sites in 2009. 

The lower Virgin River, in contrast, has relatively little native vegetation; there were 
78 USDA Forest Service RMRS-GTR-377. 2018

no sites, consisting of either native vegetation or tamarisk, that looked like they could 

be suitable flycatcher habitat but were unoccupied (M.A. McLeod, Biologist, SWCA

Environmental Consultants, personal observation, lower Virgin River). When tamarisk 

was defoliated, flycatchers at Mormon Mesa had nowhere nearby to go. The closest 

flycatcher sites to Mormon Mesa were at Mesquite (~19 miles away) and at Overton on 

the Muddy River (~7 miles away). Both were in poor condition; Mesquite had been dry, 

and Overton was suffering both from changes in streamflow that left a portion of the 

breeding site dry and from tamarisk defoliation (McLeod and Pellegrini 2015).

Short-Term Consequences

The proponents of the tamarisk biocontrol program claimed that decline in tamarisk 

would be gradual and would be accompanied by a concurrent increase in the native plant 

community. They also cited examples of how quickly native vegetation could become 

suitable habitat for nesting flycatchers as evidence that no gap in habitat availability 

would occur. However, beetles can cause complete defoliation that results in greater solar 

insolation, lower humidity, and higher temperatures (Bateman et al. 2013), which have 

been associated with low reproductive output in flycatchers (Dobbs et al. 2012; McLeod 

and Pellegrini 2013). While it is true that riparian vegetation can grow into suitable 

flycatcher nesting habitat in as little as 2 or 3 years (Paxton et al. 2007; M.A. McLeod, 

Biologist, SWCA Environmental Consultants, personal observation, Virgin and lower 

Colorado rivers), this typically occurs in areas that have been scoured by a flood or on 

sediments that are newly exposed when water levels in a reservoir decline. 

A resurgence of willows on the stream margins of areas that have been defoliated 

by tamarisk beetles has been observed in some areas, such as along the upper Colorado 

River (S. Carothers, SWCA Environmental Consultants, personal observation, upper 

Colorado River; Graham et al. 2016), but in other areas, the reduction in canopy cover 

of tamarisk has allowed weed species to proliferate. At Mormon Mesa, beetles caused 

rapid dieback and mortality of tamarisk, without concurrent recovery of native vegetation. No expansion of the willows was apparent as of the summer of 2016, 5 years after 

beetles became established in the area, but whitetop (Lepidium sp.), an invasive weed, 

had become much more widespread than it was prior to the arrival of beetles (M.A. 

McLeod, Biologist, SWCA Environmental Consultants, personal observation, Mormon 

Mesa, June 16, 2016). In short, it is now clear that beetles can cause a gap in habitat 

availability that lasts for several years, and possibly much longer.

Flycatchers, like most small passerines, are relatively short-lived birds. Estimated 

juvenile mortality along the Virgin River in 1997–2012 was 68 percent (i.e., only 

32 percent of fledglings survive to their second summer; McLeod and Pellegrini 2013), 

and estimated annual adult mortality over that same period was 39 percent. If flycatchers are unable to reproduce, the flycatcher population experiences a precipitous decline. 

In the complete absence of reproduction, the adult flycatcher population would decline 

by 77 percent in just 3 years and 91 percent in 5 years, given the estimated annual mortality rate. Even if each adult female produces, on average, one fledgling per year, the 

population would decline 73 percent after 5 years. Flycatchers do not have the longevity 

to be able to wait out several years of poor reproduction if tamarisk defoliation renders 

entire river reaches unsuitable as breeding habitat.
USDA Forest Service RMRS-GTR-377. 2018 79

In areas where flycatchers nest primarily in tamarisk-dominated habitats, alternative breeding sites are needed so that local populations can persist during the period 

when tamarisk no longer provides nesting habitat but native vegetation has not recovered. These alternative sites should be close to existing breeding sites; in multiple 

drainages where flycatchers have been studied, the vast majority of between-year movements by adult flycatchers resulted in distances moved of less than 25 miles (McLeod 

and Pellegrini 2013; Paxton et al. 2007). 

In addition, when a large area of formerly suitable habitat became unsuitable, 

no large-scale dispersal to other breeding locations was observed (Paxton et al. 2007), 

illustrating the need for local refugia. Following biocontrol, any areas that recover native vegetation in sufficient density to become flycatcher breeding habitat should be 

readily colonized by flycatchers if a local population is present. Colonization of new 

habitats that are adjacent to occupied areas has been observed in multiple flycatcher 

studies (McLeod and Pellegrini 2014; Moore and Ahlers 2008; Paxton et al. 2007), and 

flycatcher populations can increase rapidly in response to an increase in suitable habitat 

(Graber et al. 2012; Moore and Ahlers 2008; Paxton et al. 2007).

Long-Term Prospects

Beetles are expected to be eventually found in all areas of North America that 

have tamarisk (Bean et al. 2013), and these areas completely overlap the breeding range 

of the southwestern willow flycatcher. In the long term, tamarisk biocontrol may have 

exactly the effect on tamarisk that its proponents advertised: reducing the density of 

tamarisk by up to 85 percent. The authors of the 1998 biocontrol proposal suggested this 

would “reduce the abundance of saltcedar to below the level where it causes important 

damage to western riparian ecosystems.” At this level of suppression, however, tamarisk 

would also no longer provide much ecological value. Recovery of native vegetation is 

unlikely to occur in many areas that will be affected by the tamarisk beetle, and in these 

places, biocontrol will result in a long-term reduction in habitat quality.

Tamarisk is a symptom as well as a cause of the degradation of riparian ecosystems, and removing the tamarisk does not address the underlying changes that limit 

native riparian species and allow tamarisk to proliferate. The Flycatcher Recovery 

Team identified this concern early on and reiterated it in the letter sent to the Regional 

Director of the USFWS in 1998 expressing their concerns about the proposed beetle releases, pointing out that without extensive regional changes in the management of water 

and land, existing conditions would continue to preclude the establishment of native 

riparian vegetation. 

The long-term effects of tamarisk beetles on vegetation conditions are likely to 

vary widely between river systems and between reaches, depending on the prevalence 

of tamarisk and on the many factors, such as flood regimes, groundwater levels, and 

soil and water salinity, that influence whether native riparian vegetation can become 

established and persist. The fact that some of the largest flycatcher populations occur 

in sites dominated by tamarisk makes the flycatcher, among riparian obligate wildlife, 

particularly susceptible to the detrimental effects of tamarisk beetles. Based on the 

observed responses of flycatchers to tamarisk defoliation, flycatcher productivity will 

almost certainly decrease immediately following the arrival of beetles in any area 
80 USDA Forest Service RMRS-GTR-377. 2018

where flycatchers nest in vegetation with a significant tamarisk component, including at 

Elephant Butte Reservoir and on the Gila River. 

Whether beetles cause a long-term reduction in, or even extirpation of, the 

flycatcher population in these areas remains to be seen. The health, and possibly persistence, of flycatcher populations in reaches that can still support native vegetation may 

depend on active restoration of native vegetation prior to and immediately following the 

arrival of beetles. Even beetle researchers have recently acknowledged that defoliation 

and subsequent dieback of tamarisk can have adverse effects on flycatchers, including 

in stands that are not dominated by tamarisk (Tracy et al. 2014). Researchers have advocated restoration efforts in advance of the arrival of beetles (Tracy 2014).

Beetle proponents and opponents differ, often passionately, on whether tamarisk biocontrol was a good idea, but both sides ostensibly have the same goal: the preservation and 

improvement of riparian health. Now that the beetle has been set loose and is spreading 

rapidly, the common focus should be on mitigating the detrimental effects and maximizing the beneficial results of the inevitable arrival of beetles. The lawsuit that was filed in 

2013 against APHIS and the USFWS by CBD and Maricopa Audubon Society sought, 

in part, a declaration that the defendants had violated the Endangered Species Act and 

the development of “an appropriate mitigation plan to address the impacts of the beetle 

populations on flycatchers and their critical habitat.” The District Court found that APHIS 

was in violation of section 7(a)(1) of the Endangered Species Act, which requires Federal 

agencies to take actions to preserve endangered species (Center for Biological Diversity et 

al. v Vilsack et al., Dkt. 87), and issued a remedial order instructing the defendants to take 

several measures, including considering funding intensive third-party restoration efforts 

(Center for Biological Diversity et al. v Vilsack et al., Dkt. 104). 

Flycatchers have received much attention because of their status as a Federally 

endangered species, but they are, of course, not the only species affected by tamarisk 

defoliation. A study of riparian-nesting birds on the Virgin River showed that species richness and abundance were higher in 2009 and 2010, prior to the arrival of beetles, than they 

were in 2013, after 2 or 3 years of defoliation, with yellow warblers (Setophaga petechia) 

being particularly affected (Johnson 2015). Similarly, herpetofauna were less abundant 

after defoliation in both monotypic tamarisk and stands of mixed vegetation (Bateman et 

al. 2014). No study of riparian-nesting birds has been undertaken to compare the pre- and 

post-beetle nest success of species other than the flycatcher.

For the sake of all riparian obligate wildlife, restoration of native vegetation is urgently needed wherever tamarisk constitutes a significant portion of the woody riparian 

vegetation and the arrival of beetles is imminent or has already occurred. Restoration 

is often costly and labor intensive, and it will likely be prohibitively so in places where 

soil treatments are needed or depth to groundwater is such that irrigation would be 

required in perpetuity. Restoring habitat where flycatchers can nest successfully is even 

more difficult, given their propensity to select dense vegetation close to surface water. 

One strategy in places where beetles are already present is to target monotypic or 

mixed tamarisk stands that supported breeding flycatchers prior to the arrival of beetles. 

These stands likely still have the same surface water conditions that attracted breeding flycatchers but no longer have suitably dense vegetation. This strategy has been 

employed around St. George, and two successful flycatcher nests were located in one 

of the restored areas in 2017 (Christian Edwards, Utah Division of Wildlife Resources, 
USDA Forest Service RMRS-GTR-377. 2018 81

personal communication, September 11, 2017). Prior to the arrival of beetles, restoration 

efforts should target sites that can support dense vegetation in proximity to breeding 

flycatchers, as nearby sites are the most likely to be colonized. This approach is being 

used on the Gila River in the Safford Valley. Either approach requires careful planning 

and close coordination with the USFWS. These strategies represent a change from 

more traditional flycatcher management, which required avoiding nesting sites and their 

surroundings, but this kind of proactive management may be critical to the long-term 

success of the southwestern willow flycatcher.

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USDA Forest Service RMRS-GTR-377. 2018 85

Chapter 6. Beavers, Livestock, and Riparian Synergies: 

Bringing Small Mammals Into the Picture

Jennifer K. Frey

Introduction

Riparian ecosystems provide the anchor for their associated aquatic habitats and 

the structure for a unique assemblage of life found in these exceptionally productive 

ecosystems. Much of upland life also is tied to this zone, particularly in arid regions. For 

instance, on National Forest lands in the Southwest Region, 57 percent of all vertebrates 

occur in riparian ecosystems, but these systems make up <2 percent of these lands 

(Rickel 2005a). However, it has been estimated that 90 percent of riparian ecosystems 

have been lost or degraded in parts of the western United States through human-mediated factors (Ohmart and Anderson 1986). The inception of much of this degradation 

occurred during the 1800s when trappers flooded into the West exploring each river 

and tributary in pursuit of the American beaver (Castor canadensis; hereafter beaver). 

The pelts of these animals were made into felt that was used for the manufacture of 

gentlemen’s top-hats. Trade in the pelts of these and other furbearers constituted a 

major economic export to Europe, which helped fuel the economy of the young country 

(Weber 1971). As a consequence, beavers were extirpated from many streams and the 

population of beavers in North America fell from more than 60 million before the arrival of Europeans to near extinction by 1900 (Naiman et al. 1988). Prior, beavers were 

an important force that influenced the hydrology and hence overall ecology of streams 

and rivers. Beavers cut trees to build dams on smaller streams and on side channels of 

larger rivers. Beaver dams and the coarse woody debris introduced into the waterways 

due to beaver activities are a strong force that alters the stream planform by spreading 

the water into a multitude of smaller channels and by creating ponds, pools, and backwaters (Polvi and Wohl 2013). With the loss of beavers, there was a simplification of 

the complex hydrology maintained by beavers resulting in single channels of water with 

relatively high stream power and erosive force, as well as reduced storage of water that 

otherwise would be released to sustain riparian plants during dry periods and droughts. 

This resulted in the inception of stream channel incision and a narrowing and simplification of the riparian zone (Naiman et al. 1988).

Of course, other factors besides the demise of beaver also have contributed to the 

loss of riparian habitats. Nearly all western rivers have had natural flow regimes altered 

by dams and diversions. Free water is an essential limiting factor for human survival 

and hence most human settlements are located along streams and rivers. Agricultural areas were initially developed in the fertile valleys along low gradient reaches of streams 

and rivers where water could be harnessed for irrigation. In turn, the larger of these 

valuable agricultural areas helped to fuel the growth of cities, which in turn are run by 

a complex infrastructure of water delivery and waste often originating and ending in 

rivers. Exotic species have been introduced, either intentionally or by accident, some of 
86 USDA Forest Service RMRS-GTR-377. 2018

which, such as saltcedar (Tamarix spp.), have nearly wholly displaced native riparian 

communities in some areas in favor of novel systems consisting of near monotypes of 

the alien species (Shafroth et al. 2005).

Another agricultural product that exerts an influence on riparian ecosystems is 

domesticated livestock, such as cattle, horses, sheep, and goats. In the American West, 

livestock grazing is one of the few economic uses of much of the land and hence they 

are nearly ubiquitous on larger tracks of private lands and on the extensive tracks of 

public lands managed for multiple uses (e.g., U.S. Forest Service [USFS]; Bureau of 

Land Management [BLM]). These exotic animals were introduced into western North 

America in the late 1500s and hence some areas have been grazed by these animals for 

over 400 years (Bowling 1941). The behaviors and managed distributions of livestock 

are substantially different from native ungulates such as elk (Cervus elaphus) and deer 

(Odocoileus spp.). The long-term and differential use of riparian zones by livestock 

has substantially changed the composition and structure of riparian plant communities, 

typically resulting in an overall drying and simplification of the ecosystem (Belsky et 

al. 1999). And lastly, the world has embarked on a period of rapid climate change that 

is just beginning to exert its influence. In the American West, the future is expected to 

bring periods of prolonged drought as well as intense flooding. How riparian zones will 

respond to these changes is not precisely known, but it is likely to cause the continued 

deterioration of riparian ecosystems.

Water is vital for human interests and ecosystem function. Consequently, there is 

increasing interest in restoring riparian ecosystems and their associated aquatic habitats. 

Regardless of the goal (e.g., restore stream hydrology to improve fish habitat, improve 

specific riparian habitat elements) and regardless of the mechanics (e.g., use of big 

machinery to recontour streambeds, installation of instream structures, planting specific 

species), nearly all such restorations ultimately consider aspects of the riparian plant 

community. It is the deep-rooted riparian plants that ultimately stabilize the riparian and 

aquatic habitats and provide habitat for associated animals. However, to a large extent, 

these efforts focus on woody plants and their associated bird communities. This narrow 

perspective may fail to result in projects that restore full ecosystem function.

Thus, the ultimate purpose of this chapter is to call attention to an overlooked but 

vital element of riparian zones: the small mammal community, which is associated with 

herbaceous vegetation near ground level. I argue that taxonomic biases in the study of 

riparian ecosystems and frequent goals of riparian restoration have resulted in overlooking one of the most important elements of the riparian zone. Abundant and diverse small 

mammal communities support a vast array of ecosystem services. However, such communities are only fully expressed when riparian zones support a productive and diverse 

herbaceous riparian community. This may be maximally expressed through a synergism 

between healthy native riparian vegetation and beaver activities. However, this synergism can be disrupted, especially by livestock grazing, which ultimately can cause a loss 

of diversity and function to these ecosystems. I conclude by making recommendations 

on needed research to help improve understanding of these relationships and the management of these systems. Riparian restorations that fail to consider these aspects are 

not likely to reap full ecosystem benefits.
USDA Forest Service RMRS-GTR-377. 2018 87

Taxonomic Bias in Riparian Restoration

There appear to be taxonomic biases in our knowledge about riparian ecosystems 

and the typical goals and monitoring of riparian restoration. People often seem to assume that birds are the group of animals most impacted by loss and degradation of 

riparian systems, perhaps with the exception of the impact to fishes by the concomitant 

degradation of the aquatic systems. Many references on riparian restoration explicitly 

link vegetation restoration with the needs of birds or use birds as the basis for understanding existing or future conditions (e.g., Eubanks 2004; Gardner et al. 1999). 

Monitoring the success of riparian restorations typically involves birds, and sometimes 

other taxa such as reptiles or bats, but rarely includes small mammals (e.g., Bateman et 

al. 2008; but see Queheillalt and Morrison 2006).

For instance, a USFS riparian restoration guide, while noting the importance of 

riparian systems for all wildlife, singles out only the analysis of threatened and endangered species and bird communities for establishing existing conditions of a riparian 

zone (Eubanks 2004). Furthermore, bird communities are often promoted as an index 

for planning and monitoring riparian condition and restoration, to the exclusion of other 

taxa (Bryce et al. 2002; Rich 2002; Young et al. 2013). As an example, it has been proposed that evaluation of breeding bird communities should be used as a means to assess 

“Proper Functioning Condition,” which is the main method that Federal land management agencies, including the USFS, BLM, and Natural Resources Conservation Service, 

use to evaluate riparian health (Rich 2002).

I believe there are numerous reasons for this seeming taxonomic bias, although 

few have a biological basis. First, birds are viewed as charismatic and valuable animals. 

Besides the human fascination with flight, most birds are relatively easy to observe due 

to their flight, diurnal behavior, and repertoire of often loud, beautiful, and distinctive 

vocalizations. Pronounced morphological and plumage variation in this diverse group 

allows for relatively easy identification of species and often identification of different 

genders and ages, allowing even casual observers to understand something a bit deeper 

about their biology. As a consequence, birds are enormously popular with the public, 

which has led to formation of powerful lay advocacy groups (e.g., National Audubon 

Society). Such organizations convey numerous tangible and intangible benefits to birds 

by bringing attention to issues, garnering resources, and influencing legislation.

Further, the accessibility of birds allows citizens to participate in the collection of 

biologically meaningful data (e.g., winter bird count, rare bird alert, eBird) that help to 

expand the knowledge base of these organisms. Just as birds are popular with the public, 

they also are popular study organisms for scientists. For instance, consider that in North 

America there is only a single professional society dedicated to the study of mammals 

(American Society of Mammalogists), while in contrast there are at least four professional societies dedicated to the study of birds (American Ornithologists Union, Cooper 

Ornithological Society, Association of Field Ornithologists, Wilson Ornithological 

Society). One important consequence of this heightened attention and knowledge is that 

birds might be more likely to be listed as threatened or endangered than other taxonomic 

groups, and the perceived bar for “endangerment” might be lower. In turn, threatened or 

endangered listing can stimulate research and management focuses.
88 USDA Forest Service RMRS-GTR-377. 2018

While there is no doubt that birds can be impacted by changes to riparian habitats 

and that they provide ecosystem services (Whelan et al. 2015), the taxonomic bias 

focuses attention on a suite of species that are primarily associated with the woody 

components of riparian ecosystems. This approach parallels another apparent bias in 

how humans tend to perceive ecosystems. Although trees and shrubs are usually a minor 

component of the diversity of riparian zones, they are usually the largest species present and hence they receive our differential attention. As an example, most vegetation 

classification schemes are based, in large part, on the woody species present (e.g., U.S. 

National Vegetation Classification). In addition, because these plants are relatively longlived, they are used as a benchmark or proxy for assessing the “health” of a riparian 

system or success of a restoration project. The main focus of many riparian restoration 

projects is planting woody species or adjusting hydrology to encourage their natural 

regeneration (e.g., Dreesen et al. 2002). Thus, a focus on avian species seems to dovetail 

nicely with a corresponding viewpoint of the importance of the woody component of 

riparian habitats.

Although birds are diverse, their influence on the structure and function of riparian ecosystems may be relatively weak compared to other taxonomic groups. There 

are several reasons for this. First, due to territoriality and high vagility, birds tend to 

be relatively sparsely distributed. As a result of this rarity, and in combination with 

their relatively small body size, birds usually account for but a minor proportion of the 

animal biomass in a given area (Turner and Chew 1981). Consequently, birds have relatively little influence on higher trophic levels. On the other hand, due to the relatively 

high diversity of birds, their greatest ecosystem influence might be a consequence of 

some foraging behaviors (Whelan et al. 2015). For instance, granivores are considered 

important agents of seed predation and dispersal, even though when compared directly 

with rodents, the impacts of birds are relatively weak (e.g., Hulme and Benkman 2002; 

Mares and Rosenzweig 1978). Some birds, such as hummingbirds, transfer pollen that 

benefits some plants. Insectivores may help regulate invertebrate communities, while 

raptors may help regulate some rodent communities (Whelan et al. 2015). Other services 

provided by some kinds of birds, such as creating cavity holes in trees by woodpeckers, 

while important, are of more minor impact to the overall ecosystem.

Although the restoration of riparian habitats is considered important for the 

maintenance of bird diversity (Gardner et al. 1999), the taxonomic bias on birds focuses 

attention on species that may be only weakly linked to overall ecosystem function or 

may cause other important aspects of ecosystem function to be missed. In contrast, 

other elements of the riparian zone, in particular small mammals, have been mostly 

overlooked and yet they may have more strong influence on overall ecosystem structure 

and function. Consequently, riparian habitat restorations without consideration of small 

mammals, and the riparian habitat elements they require, will be incomplete and may 

not provide the full range of ecosystem services.

Terrestrial Small Mammal Riparian Communities

In western North America, terrestrial small mammals strongly associated with 

riparian zones include members of the order Eulipotyphla (e.g., shrews [Soricidae], 

moles [Talpidae]); Rodentia (e.g., deer mice [Cricetidae], cotton rats [Sigmodontidae], 
USDA Forest Service RMRS-GTR-377. 2018 89

voles [Arvicolidae], jumping mice [Zapodidae]); and some small members of the orders 

Carnivora (e.g., short-tailed weasel [Mustela erminea; Mustelidae]) and Lagomorpha 

(e.g., some Sylvilagus). These species can be categorized into two groups: those that 

are relatively specialized on riparian ecosystems and those that become disproportionately abundant within riparian ecosystems. Examples of species in the first category, 

the riparian specialists, include: marsh shrew (Sorex bendirii), cordilleran water shrew 

(Sorex navigator), water vole (Microtus richardsoni), muskrat (Ondatra zibethicus), and 

some jumping mice (e.g., Zapus luteus luteus and Z. hudsonius preblei). Examples of 

species in the much larger second category include: montane shrew (Sorex monticola), 

Townsend’s mole (Scapanus townsendii), white-footed deermouse (Peromyscus leucopus), North American deermouse (Peromyscus maniculatus), western harvest mouse 

(Reithrodontomys megalotis), hispid cotton rat (Sigmodon hispidus), tawny-bellied 

cotton rat (S. fulviventer), California vole (Microtus californicus), long-tailed vole 

(Microtus longicaudus), montane vole (Microtus montanus), meadow vole (M. pennsylvanicus), white-footed vole (Phenacomys albipes), western heather vole (Phenacomys 

intermedius), western jumping mouse (Zapus princeps), Pacific jumping mouse (Zapus 

trinotatus), brush rabbit (Sylvilagus bachmani), and short-tailed weasel.

In addition, other species that are more typical of other vegetation types may also 

occur in the riparian zone when the communities closely abut (e.g., small order stream 

in coniferous forest), such as the cinereus shrew (Sorex cinereus), mountain cottontail 

(Sylvilagus nuttalii), least chipmunk (Neotamias minimus), red squirrel (Tamiasciurus 

hudsonicus), Botta’s pocket gopher (Thomomys bottae), northern pocket gopher (T. talpoides), Mexican woodrat (Neotoma mexicana), and southern red-backed vole (Myodes 

gapperi). In the Mountain West, riparian small mammal communities become more 

unique in comparison with uplands as elevation increases (Olson and Knopf 1988). 

Finally, it should be recognized that many species exhibit geographic variation in their 

habitat associations such that a species may be a riparian associate in a more mesic 

region, but become more of a riparian specialist in a more xeric region (e.g., California 

vole; Conroy et al. 2016).

Although terrestrial small mammals are important members of riparian communities, they have been mostly overlooked both in terms of knowledge of their ecology 

and focus in riparian restorations (but see for example Rickel 2005b and Golet et al. 

2008). There are a number of reasons for this. First, small mammals tend to be relatively 

difficult to study. Most are nocturnal and live in burrows or other hidden places such 

that they are not easily observed. Study usually requires capturing individuals, which is 

labor intensive and necessitates specialized equipment, permits, and expertise. Because 

most small mammals have generalized, nondescript body plans, accurate identifications 

can be difficult, often requiring examination of cranial characters (necessitating collection and preparation of series of museum specimens) or DNA (which is expensive and 

requires specialized equipment).

These difficulties largely preclude the public, and even many scientists, from informal or formal study of these organisms. Thus, relative to some other taxonomic groups, 

such as birds, fishes, and big game mammals, there are relatively few scientists that 

specialize on studying small mammals. Perhaps more importantly, most of these species, with the exception perhaps of some squirrels, simply lack charisma or perceived 

value (e.g., few are considered game species). To many people, these “rats and mice” 
90 USDA Forest Service RMRS-GTR-377. 2018

are simply vermin. Consequently, no public organizations that promote these species 

exist and there is limited funding and political will. Taken together, these factors cause 

knowledge about riparian small mammals to lag far behind other species. For instance, 

the recent comprehensive Beaver Restoration Guidelines reference manual (Pollock et 

al. 2015), which reviewed the ecological impacts of beavers on other organisms, included sections on birds, fishes, and invertebrates, but did not include mention of ecological 

impacts to other mammals.

Ecological Roles of Riparian Small Mammals

The ecological roles of most small mammals have been poorly studied. Yet, limited research indicates that the ecosystem services provided by small mammals are strong 

and important to diverse ecosystems, including riparian zones. By way of example, I 

highlight four essential aspects of the role of small mammals in riparian ecosystems, 

including: (1) dominance of animal biomass, (2) prey base for diverse carnivore communities, (3) influence on soil condition, and (4) influence on plant composition and 

succession.

Animal Production and Biomass

Although data are limited, small mammals likely constitute the dominant proportion of vertebrate animal production and biomass in healthy riparian systems. For 

instance, Turner and Chew (1981) found that production of terrestrial small mammals in 

arid environments of southwestern North America far outweighed production by other 

groups of animals (in contrast, birds were among the lowest). In part, this is because 

small mammals, while having small body size compared with other mammals, are on 

average larger than most other kinds of organisms. In addition, terrestrial small mammals are year-round residents and they have relatively small and overlapping home 

ranges. For example, home ranges of the meadow vole may be as little as 160 m2 (Van 

Vleck 1969).

Further, small mammals are prone to population irruptions that can produce 

extremely high densities and biomass. Generation of exceptionally high biomass is 

particularly true for the graminivorous (grass eating) riparian species, including the cotton rats at lower latitudes and elevations and the voles at higher latitudes and elevations; 

both groups may display population cycles or strong annual variation in population 

densities (Fagerstone and Ramey 1996; Grant et al. 1982; Rickel 2005b; Taitt and Krebs 

1985). For instance, densities during population highs for these species can exceed 369 

cotton rats/ha (Guthery et al. 1979) and 7,400 voles/ha (Spencer 1958). Production of 

small mammals in vole-dominated communities can exceed 5,000 kcal/ha in high quality habitats (Grant et al. 1982).

Small mammal populations are renowned for their volatility (Witmer and Proulx. 

2010). Whitford (2002) identified episodic species as those that respond to periods 

of high primary production with high rates of reproduction and population growth, 

and hence driving high rates of secondary production. Riparian zones may serve as 

important refugia for episodic species in arid environments. During wet periods, episodic species experience population growth and immigration into marginal habitats, 

where they can exhibit explosive growth relative to corresponding dry periods. Thus, 
USDA Forest Service RMRS-GTR-377. 2018 91

population dynamics of small mammals in riparian ecosystems can have far reaching 

and direct impacts on adjacent terrestrial ecosystems.

Fuel for Predator Communities

The high density and biomass of small mammals that concentrates in healthy 

riparian zones provides fuel for supporting a diverse predator community that includes 

hawks, owls, snakes, and myriad mammalian carnivores such as weasels, minks, 

bobcats, foxes, and coyotes (Hamilton et al. 2015). Starvation is a real and constant 

threat for many kinds of predators. Consequently, a high threshold of prey abundance 

is required for many kinds of predators to persist in an area. For instance, weasels have 

particularly high energy demands due to their active lifestyle and their lean, narrow 

body plan. Meals pass through their short digestive tracks in just a few hours, meaning they must eat frequently. It has been estimated that long-tailed weasels (Mustela 

frenata) must consume 20−40 percent of their body weight in small mammal prey 

every day. However, the energetic demands for smaller weasels are even more extreme. 

For instance, captive least weasels had to eat meals every 2.5 to 3 hours, totaling 5−10 

meals per day. Small weasels are not likely to be able to survive more than 24 hours 

without eating (Gillingham 1984). The energetic demands of females that successfully 

raise young are even higher. Voles are the main group of small mammals in the temperate zone that can generate the densities and biomass of animal flesh required to support 

predators that have high energy demands (Frey and Calkins 2013; Rickel 2005b). Thus, 

it is no surprise that many species of owls also specialize on voles preferential to other 

kinds of small mammals (e.g., Colvin and McLean 1986).

Exceptionally high abundance of suitable prey is thought to be necessary to permit coexistence of some predators. For instance, short-tailed weasels and long-tailed 

weasels are sympatric across much of the western United States. However, models 

indicate that it might not be possible for both species to coexist in a local habitat unless 

the prey populations are high and diverse (Powell and Zielinski 1983). Riparian zones 

also provide an important source of alternate prey for more specialized predators such as 

Canada lynx (Lynx canadensis), river otters (Lontra canadensis), martens (Martes spp.), 

fishers (Pekania pennati), wolverines (Gulo gulo), and wolves (Canis lupus). Alternate 

prey serves as a critical resource for these predators during certain seasons or years or 

by certain demographic groups (e.g., nursing mothers). Thus, small mammals contribute 

to the diversity of riparian zones, not only through the presence of a unique assemblage 

of riparian small mammals, but also by supporting diverse predator communities. By 

supporting diverse predator communities, riparian small mammals provide an important 

link between the riparian zone and adjacent upland communities.

Soil Condition

Soil is formed over time primarily through the interactions of the parent material, 

topography, climate, vegetation, and invertebrate animals. However, small mammals 

also play an important role in the structure and function of soils through their burrowing 

activities, underground caching of seeds and other plant parts, and decomposition of 

latrines and carcasses. Most terrestrial small mammals construct subterranean burrows 

or modify and utilize the burrows constructed by other species of small mammals, at 
92 USDA Forest Service RMRS-GTR-377. 2018

least for some aspects of their life cycle (e.g., hibernacula, maternal nest chambers). 

Some species, such as the pocket gophers and moles, spend most of their lives within 

these burrows. Burrowing activity has an important role in improving soil structure by 

loosening soil particles and in mixing soils by bringing soils from lower strata to the 

surface, to the extent that sometimes this mixing can obliterate the upper soil horizons 

(Hendricks 1985). For instance, estimates of soil excavated by pocket gophers can exceed more than 100 Mg/ha (Cox 1990; Grinnell 1923).

Burrow systems also affect the soil climate by enhancing infiltration of oxygen 

and water. Subterranean chambers made by small mammals to store food caches, or 

serve as nest sites or latrines, create local concentrations of key limiting nutrients such 

as salts, nitrogen, phosphorus, and potassium (Hendricks 1985). Finally, like woodpeckers creating tree cavities that may be used by other species, the burrows constructed by 

small mammals can harbor an array of non-burrowing organisms such as fungi, spiders, 

amphibians, and snakes, thus supporting overall biodiversity (Scheffer 1945).

Plant Community Composition and Succession

Although understudied, terrestrial small mammals may have a profound impact 

on the structure and function of plant communities. For instance, Bryce et al. (2013) 

demonstrated that the long-term impacts of vole burrowing activities and herbivory result 

in a patchwork of different plant successional stages in riparian systems. Further, small 

mammals impact the structure and function of plant communities through their predation 

on seeds and seedlings and concomitantly through the dispersal of seeds and mycorrhizal 

fungi. Seeds are an important part of the diet of most small mammals, including species 

such as shrews, which are normally thought of as strictly insectivorous (Hallett et al. 

2003). Small mammal seed predators include species such as harvest mice and jumping 

mice that specialize on harvesting seeds from grasses and other herbaceous plants prior to 

seed dispersal (e.g., Wright and Frey 2014), as well as species such as chipmunks, pocket 

mice (e.g., Chaetodipus spp.), and deer mice that forage on a wide range of seed types 

from both herbaceous and woody plants in the seed rain or seedbank.

In the temperate zone, most seed removal is due to small mammals (Hulme and 

Benkman 2002; Mares and Rosenzweig 1978). Thus, granivory by small mammals can 

have a profound impact on seed populations, and hence plant communities (Hulme and 

Benkman 2002). For instance, the extent of small mammal seed predation can be so high 

(>95 percent of seeds sown) that it can hamper efforts to restore forests and other ecosystems via direct seeding (Hallett et al. 2003). Similarly, seedlings and saplings are also 

vulnerable to predation by rodents, including those of conifer trees that are especially 

vulnerable under cover of snow when herbaceous plants are less available (Hallett et al. 

2003). Thus, small mammal herbivory tends to impede succession, thereby maintaining 

early successional habitats that provide favorable food and cover (Davidson 1993).

Few studies have examined the role of mammals in seed dispersal. However, the 

number of fruit seeds dispersed into a plant population by medium-sized mammals, such 

as foxes, may be twice that mediated by frugivorous birds (Jordano et al. 2007). Further, 

although small mammals are efficient predators on some seeds, many also gather, move, 

and store these seeds in underground burrows. Unrecovered seeds cached by these 
USDA Forest Service RMRS-GTR-377. 2018 93

rodents may be essential to the establishment of some plants (e.g., Hallett et al. 2003; 

Longland and Ostoja 2013).

Perhaps even more important than the role mammals play in seed dispersal and 

germination is their crucial relationship with mycorrhizal fungi. Mycorrhizal fungi form 

symbiotic relationships with the roots of most vascular plants. Growth and survival of 

many plants is dependent on this relationship, which provides for the uptake of water 

and nutrients (Molina 1994). Concurrently, mycorrhizal fungi are an important component of the diet for many small mammals, but the spores of these fungi pass through the 

digestive tracts of the small mammals with no loss in viability (Maser et al. 1978). Thus, 

small mammals have been implicated as the dominant means for dispersal of mycorrhizal inocula, which in turn must be present to support plants. Small mammals, therefore, 

may control some aspects of succession.

For instance, an interesting study by Terwilliger and Pastor (1999) concluded 

that small mammals regulated the succession of trees into “beaver meadows.” Beaver 

meadows are herbaceous dominated ecosystems that form in the silt that is left behind 

after a beaver dam has been breached and the pond drained. These are exceptionally important habitats, especially in coniferous forest dominated regions, and they can persist 

for many decades, although an explanation for their longevity was lacking. Terwilliger 

and Pastor (1999) found that the soils of beaver meadows lacked the mycorrhizal fungi 

necessary for the growth of conifers and attributed that to long-term inundation of the 

soils by the former pond. In their study system, the primary consumer of mycorrhizal 

fungi associated with conifers was the red-backed vole (Myodes gapperi), which mostly 

occurs in forested habitat. In contrast, they speculated that meadow voles, which are 

associated with the graminoid habitats of the beaver meadows, competitively excluded 

red-backed voles and their spore containing feces from the meadows. Thus, it was 

speculated that reestablishment of conifers in the beaver meadows was limited both by 

patterns of mycorrhizal fungal distribution and by use by small mammals and the competitive interactions among them.

Terrestrial Small Mammal Community Habitat Relationships

Terrestrial small mammals serve as prey for a host of predators, such as snakes, 

hawks, owls, minks, foxes, bobcats, and bears. Consequently, appropriate concealment 

cover from predators is one of the most important overriding microhabitat component 

required for the production of abundant and diverse small mammal communities (e.g., 

Longland and Price 1991). In riparian ecosystems, these communities are mainly associated with early seral plant communities typified by tall, dense herbaceous ground cover, 

often in conjunction with riparian shrubs such as willow and alders (e.g., Dickson and 

Williamson 1988; Golet et al. 2008).

Because these communities are often associated with riparian shrubs, vegetation 

maps are likely to underestimate the extent of this vegetation type and there are no estimates for the proportion of riparian systems that support herbaceous riparian vegetation. 

These herbaceous plant communities also provide necessary food for small mammals. 

Small mammals have high energetic demands and utilize a wide range of foods. Many 

small mammals eat the seeds of grasses and forbs. Other small mammals, such as the 
94 USDA Forest Service RMRS-GTR-377. 2018

voles and cotton rats, eat the vegetative parts of graminoids, while pocket gophers utilize the roots of herbaceous plants.

In well-developed herbaceous riparian habitats, graminoid food is not limiting, 

which contributes to potential for exceptionally high biomass of voles and cotton rats. 

Invertebrates are supported by diverse and abundant riparian vegetation and these 

are utilized as food by shrews and moles, and to a lesser extent by many rodents. In 

contrast, late seral riparian vegetation types are dominated by trees (e.g., cottonwood 

gallery forest) and consequently may have lower densities and diversity of mammals 

due to the drier soils and lack of adequate ground cover (Andersen and Nelson 1999). 

Furthermore, in late seral riparian habitats that have been substantially degraded, such 

as via conversion to monotypes of saltcedar, small mammal communities may be indistinguishable from upland communities (i.e., due to loss of riparian adapted species and 

invasion by upland species) or become dominated by disturbance adapted species such 

as deer mice (e.g., Ellis et al. 1997).

Riparian habitats that offer tall, dense, diverse herbaceous vegetation are maximally developed in locations with relatively low topographic relief, low gradient, broad 

floodplains, high soil moisture, and high exposure to solar radiation (low tree canopy 

cover or lack of shading by adjacent mountainsides; Dickson and Williamson 1988). To 

a large extent these characteristics are determined by local landscape features. However, 

beavers are unique in that their activities can increase the distribution and abundance of 

riparian characteristics that support early seral stage herbaceous communities. Beavers 

increase these habitats through a variety of mechanisms, including felling trees, creating 

dams that retain sediments, creating ponds that kill trees (due to submergence), increasing the area of sub-irrigated soils, storing water that is released as base flows during dry 

periods or drought, and creation of beaver meadows following abandonment of ponds 

(McMaster and McMaster 2000; Naiman et al. 1986, 1988; Rosell et al. 2005). For 

instance, Naiman et al. (1986) found that beaver dams increased the wetted surface area 

of the channel by several hundred-fold, while in Wyoming the width of riparian zones 

was 33.9 m in streams with beaver ponds but only 10.5 m in streams without such ponds 

(McKinstry et al. 2001).

Although research is limited, studies indicate that beaver activities can have a profound impact on small mammal communities. For instance, in Idaho, Medin and Clary 

(1991) compared riparian small mammal communities in an area within a complex of 

beaver ponds versus a control reach that did not have beaver ponds. They found that 

the standing crop biomass (g/ha) of small mammals was about 2.7 times higher at the 

beaver complex. This was due to exceptionally higher density of shrews and voles in the 

beaver modified habitats. In addition, the western jumping mouse was only found in the 

beaver modified habitat. Medin and Clary (1991) attributed differences in small mammal communities to the dense and structurally complex vegetation created by beavers, 

which provide food and cover. Similarly, in Oregon, Suzuki (1992) found significantly 

more deer mice (Peromyscus), voles in genus Microtus, Pacific jumping mice, and 

certain species of shrews (Sorex) on stream reaches occupied by beavers compared to 

stream reaches lacking beavers. In the American Southwest, beaver dams were implicated as important for the occurrence of the short-tailed weasel, which is of conservation 

concern (Frey and Calkins 2013), and the Federally endangered New Mexico meadow 

jumping mouse (Z. l. luteus), which is a riparian obligate (Frey and Malaney 2009).
USDA Forest Service RMRS-GTR-377. 2018 95

In contrast to the positive influence of beavers on herbaceous habitat and small 

mammal communities, livestock (and sometimes overabundant native ungulates such as 

elk [Cervus elaphus]) can have a deleterious influence on herbaceous riparian habitats. 

Although the specific impacts of livestock grazing on riparian ecosystems are dependent 

on a number of variables and some studies have been based on poorly designed methods 

(Milchunas and Lauenroth 1999; Sarr 2002), the vast majority of evidence indicates that 

riparian systems are especially prone to excessive grazing that can lead to a disruption 

of ecosystem structure and function (e.g., Belsky et al. 1999; Fleischner 2002; Trimble 

and Mendel 1995). As large-bodied nonnative grazers and browsers, livestock can negatively impact or obliterate herbaceous riparian vegetation through herbage removal, soil 

compaction, and trampling (Giuliano and Homyack 2004; Johnston and Anthony 2008; 

Kauffman and Krueger 1984). Once deep-rooted riparian plants have been reduced or 

eliminated from riparian communities, stream banks can erode and channels can downcut, further exacerbating changes to the riparian plant and animal communities.

Small mammals are sensitive indicators of disturbances and changes to plant communities, particularly changes that impact the herbaceous layer (Fagerstone and Ramey 

1996). For instance, in Oregon Moser and Witmer (2000) found significantly higher 

abundance, species richness, and species diversity of small mammals in ungrazed areas 

versus areas grazed by elk and cattle, while no differences were exhibited in similar 

metrics of the bird or plant community. While abundance of some upland or generalist mammal species, such as the North American deermouse, can increase in riparian 

areas that are grazed, most studies show a marked decrease in diversity and abundance 

of small mammals in grazed riparian areas. This pattern is especially pronounced for 

species that are more restricted to these areas such as shrews, voles, cotton rats, harvest 

mice, and jumping mice (e.g., Fagerstone and Ramey 1996; Frey and Malaney 2009; 

Medin and Clary 1989; Schulz and Leininger 1991).

For instance, voles select areas with high vegetation cover, which provides concealment from predators, reduces antagonistic interactions among individuals, provides 

food, facilitates subnivean spaces during winter, and moderates temperature and humidity. A threshold of vegetative cover may be necessary to support a population and allow 

for population buildups (Fagerstone and Ramey 1996). Consequently, voles tend to be 

relatively intolerant of livestock grazing that reduces cover, and their populations may 

be greatly depressed or extirpated in locations where grazing has greatly reduced herbaceous cover (Sullivan and Sullivan 2013).

Besides causing changes to riparian habitats that directly impact small mammal 

communities, livestock grazing can also influence suitability of riparian habitats for beavers. Beavers prefer herbaceous plants for food, but will use certain deciduous trees and 

shrubs for food when herbaceous plants are not available, such as under ice in winter 

(Müller-Schwarze and Sun 2003). In western North America where there are relatively 

few species of deciduous trees and shrubs to choose from, beavers exhibit a mutualistic 

relationship with willows, wherein beavers benefit from willows for food and building material, while willows benefit from beavers via the increased area of wetted soil 

created by their dams and asexual reproduction by resprouting of cut limbs (Kindschy 

1989; Peinetti et al. 2009).

As strict herbivores, beavers enter into direct competition with livestock and native ungulates for food. Livestock preferentially graze on herbaceous vegetation during 
96 USDA Forest Service RMRS-GTR-377. 2018

the early part of the growing season but then switch to browsing on willows during the 

latter part of the growing season (Pelster et al. 2004). Thus, excessive livestock grazing 

can result in loss of both herbaceous vegetation and willows, creating depauperate riparian zones that resemble upland communities (e.g., Small et al. 2016). A similar process 

can occur due to abundant native ungulates when top predators have been removed from 

an ecosystem. For instance, in Yellowstone National Park an increase in beavers occurred after restoration of wolves, likely due to decreased use of riparian zones by elk, 

which allowed for increased growth of willows and other riparian plants (Ripple and 

Beschta 2012).

Beavers are sometimes able to exist, at least temporarily, in marginal habitats 

that appear to offer scant resources for food and building material. In New Mexico for 

instance, I have observed beavers building dams and living on small order streams at 

the opening of a narrow sheer-sided canyon in the Chihuahuan Desert, and occurring in 

other locations where the only available dam building material was saltcedar, sagebrush 

(Artemisia), or cattails (Typha). On the Rio Grande in the Chihuahuan Desert, beavers 

can persist in small isolated pools of water after river flows are shut off by upstream 

dams (Barela and Frey 2016).

Given that beavers can exist in marginal environments, if conditions are inadequate for beavers, they also are likely inadequate to support healthy small mammal 

communities. Thus, the status of beavers may suggest the concomitant status of small 

mammal communities. For example, in New Mexico, Small et al. (2016) found only 

38 active primary beaver dams on Federal public lands throughout the State, despite 

historical efforts to restore the species. The near absence of beavers was attributed to 

the loss of riparian habitat as a consequence of nearly ubiquitous cattle grazing. Given 

that riparian habitat conditions are mostly not adequate to support beavers, this suggests 

that riparian habitats are also mostly not capable of supporting diverse and abundant 

small mammal communities. The recent listing of a riparian habitat specialist, the New 

Mexico meadow jumping mouse, as endangered supports this idea and suggests the 

need for managers to more carefully consider the needs of riparian small mammals (and 

beavers) in management plans.

Conclusions

Small mammals are an often overlooked but vitally important component of 

healthy riparian ecosystems. These species provide critical ecosystem services to riparian zones by virtue of their high biomass, support of diverse predator communities, 

physical alterations of the soil, and regulation of plant communities. Small mammal 

communities are best developed in riparian systems that provide an abundance of tall, 

dense, and diverse herbaceous vegetation. Beavers are capable of increasing the capacity of riparian systems to produce these early seral plant communities and hence benefit 

small mammal communities. In contrast, excessive livestock grazing can disrupt small 

mammal communities by causing loss of tall, dense, and diverse herbaceous vegetation, 

and can also limit the capacity of riparian systems to support beavers. This negative 

synergism can result in riparian ecosystems that are depleted and fail to support critical 

ecosystem services.
USDA Forest Service RMRS-GTR-377. 2018 97

Thus, restoration of full ecosystem services of riparian zones requires consideration of the herbaceous plant and small mammal communities. In contrast, riparian 

restoration that focuses mainly on woody plants might restore habitats for birds and 

stabilize stream banks, but may fail to provide full restoration of crucial ecosystem 

function. Additional research is needed on livestock grazing management that can 

enhance herbaceous riparian vegetation and thereby support beavers and healthy small 

mammal communities. In addition, there is need for more research on the roles of small 

mammals in riparian ecosystem function and the patterns of riparian mammal diversity 

and abundance in relation to various disturbances and management actions.

Acknowledgments

I thank T.C. Frey, R. Goljani, I. Perkins-Taylor, and several peer-reviewers for 

comments on previous versions of this manuscript.

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102 USDA Forest Service RMRS-GTR-377. 2018

Chapter 7. Euro-American Beaver Trapping and Its LongTerm Impact on Drainage Network Form and Function, 

Water Abundance, Delivery, and System Stability

Suzanne C. Fouty

Introduction

Euro-American (EA) beaver trapping was a regional and watershed-scale disturbance that occurred across the North American continent. This concentrated removal of 

beavers altered drainages by creating thousands of localized base-level drops as beaver 

dams failed and were not repaired. These base-level drops led to the development of 

channels as ponds drained and water eroded the fine sediment trapped behind the dams 

(Dobyns 1981; Fouty 1996, 2003; Parker et al. 1985). The speed at which drainages 

transformed from beaver-dominated to channel-dominated varied as a function of 

climate, upland and riparian vegetation, and the subsequent land uses. As the drainage 

network pattern changed, flood magnitudes and frequencies increased and base flows 

decreased, creating stream systems much more sensitive to climatic variability.

In most places, trapping predates EA settlement. The one exception is New 

England in the 1600s where settlement and trapping co-existed in time. The journals 

from this time period provide intriguing but limited references to vegetative and 

ecological changes as a result of beaver removal (Cronon 1983). This limited historic 

documentation of system response to the widespread removal of beavers resulted in our 

missing the significance of early EA beaver trapping on drainage network development, 

water abundance, delivery, and system stability. However, the abundant and high quality 

written records left by the early General Land Office (GLO) surveyors and scientific and 

military expeditions of their observations of the Southwest and Intermountain West hold 

promise because they entered these areas only 10 to 30 years post-trapping. Their observations when combined with later research suggest that EA trapping and grazing, though 

temporally and spatially separated, led to extensive drainage network development in 

the West and an increase in the sensitivity of stream ecosystems to climate variability. 

This transformation pre-dates the installation of stream gages and the data collection 

that forms the basis of our understanding of hydraulic geometry, channel form, and 

fluvial processes. Consequently, current fluvial geomorphic relationships and our understanding of stream sensitivity to climatic variability reflect highly disturbed watersheds 

and ecosystems, not healthy intact systems. Without realizing it, the GLO surveyors 

and the early military and scientific expeditions had captured not virgin territory, but a 

landscape in transition.

Using current research and historic observations, I developed a conceptual model 

describing the geomorphic and hydrologic response of a drainage basin to the entry of 

beavers and then their removal or abandonment (Fouty 2003). The conceptual model is 

similar in structure to Cooke and Reeves’ (1976) deductive model of arroyo formation 

in the Southwest in that both models examine the hydrologic and geomorphic response 
USDA Forest Service RMRS-GTR-377. 2018 103

of streams to EA disturbances. Cooke and Reeves (1976) focused on post-settlement 

EA disturbances such as livestock grazing, logging, agriculture, and road building. The 

conceptual model presented here steps further back in time to examine the impact of EA 

beaver trapping on drainage systems.

Currently, the most visible impact of beavers would occur on first- through fourthorder streams because beavers can build their dams across these stream sizes (Naiman et 

al. 1986; 1988) and thus have a direct effect on the physical appearance and hydrologic 

behavior of the drainage network in these areas. However, historically beaver dams 

would have occurred on higher order streams. The current widths and depths we see 

for larger order streams is an artifact of and response to the magnitude and abundance 

of disturbances that have occurred. Prior to trapping and subsequent EA disturbances, 

streams in many places were anabranching, narrower and had high stream-valley floor 

connectivity (Sedell and Froggatt 1984; Walter and Merritts 2008).

With the conceptual model as the framework for analysis, this paper then (1) explores why EA beaver trapping as a major disturbance was missed and absent from the 

discipline of fluvial geomorphology until recently, (2) reexamines the early observations 

in the Southwest and Intermountain West in the context of EA beaver trapping, and (3) 

concludes with a brief discussion of Leopold and Maddock’s (1953) hydraulic geometry 

relationships, one of the foundations of fluvial geomorphology. When these relationships are put into their historic disturbance context, one that includes beaver trapping 

and later land uses, the magnitude and longevity of the impact of EA beaver trapping on 

drainage networks, water abundance, and system resiliency becomes clear.

Conceptual Model Part 1: Watershed Response to Long-Term Beaver 

Presence

The fluvial processes and sequence of events that occur when beavers enter a drainage area and establish a long-term stable presence are shown in figure 19. An excellent 

discussion of the changes in more detail can be found in Naiman et al. (1988) and so are 

only briefly discussed here. Changes begin with dam building and foraging for vegetation 

around the pond area. As beaver dam complexes expand, changes occur to (1) water abundance and vegetation, (2) local and downstream hydrographs, and (3) the character of the 

drainage network. In figure 19, the local effects are broken down into low-energy environments and high-energy environments. The distinction is made because dams in low-energy 

environments tend to be stable, allowing wetlands to form and evolve into meadows (if 

sediment is abundant), while dams in high-energy environments (i.e., steeper gradient 

streams, higher discharge streams) are more prone to periodic failure and wetlands and 

meadows less likely to form (Meentemeyer and Butler 1999).

In terms of conceptual models, there are key differences between beaver-dominated 

and channel-dominated systems. In beaver-dominated systems, stream channels are interrupted by dams that develop ponds across the channel and valley floor and lead to the 

creation of wetlands. In contrast, the conceptual model of a channelized drainage network 

is visualized as a set of interconnected channels “where physical variables present a continuous gradient of physical conditions from the headwaters to the mouth” (Naiman et al. 

1986). This conceptual model of a watershed of interconnected channels is implied in the 

hydraulic geometry relationships of Leopold and Maddock (1953) and other researchers 
104 USDA Forest Service RMRS-GTR-377. 2018

Localized channelization.

Possible 

downstream scour. Knickpoint forms.

Ponds drain over 

time.

Ponds drain abruptly.

Dams fail over time. Dams fail abruptly.

New pond forms.

Dams repaired or new dam built in vicinity.

Sediment behind 

dams erodes.

 

Figure 19—Conceptual model of how beavers influence fluvial systems at the watershed scale (source: Fouty 2003; 

portion in the shaded box is from Naiman et al. 1998).

BEAVER ENTER A CHANNELIZED UPPER WATERSHED

Decrease in 

suspended sediment 

load downstream.

Rise in local 

base level.

Sediment 

captured in 

ponds.

Low Energy Environments 

(Dams tend to hold)

Riparian and wetland vegetation communities 

expand in distribution and diversity creating a 

mosaic of different communities on valley floor.

Water table rises and great volume of 

water stored in the subsurface

Greater and more frequent recharge of valley 

floor sediments and water table.

LOCAL EFFECTS

Increase in the temporary storage of water on the 

valley floor in upper watershed.

Valley floor floods at lower magnitude flows.

Decrease in distance 

between water surface 

and valley floor.

Increase in lateral 

flow from pond 

into bank.

Composition and density of vegetation around 

the pond changes as foraging begins and the 

water table rises and stabilizes. Water ponds behind dams.

Dams built.

Higher base flows 

and cooler low flow 

stream temperatures.

Improved water 

quality as less 

sediment transported 

downstream and 

cooler low flow 

stream temperatures.

Decrease in flood 

magnitudes and 

increase in flood 

durations for a given 

storm.

LOWER WATERSHED EFFECTS

Tributaries adjust 

to higher base 

level by aggrading

when sediment is 

available.

High Energy Environment 

(Dams frequently fail)

Erosive power of 

stream flow decreases.

Reduction in 

Available Channel 

capacity (ACC).

Pond New 

Pond

Emergent 

wetland

Stream 

Bog

Meadow

Forested 

wetland

Old Pond
USDA Forest Service RMRS-GTR-377. 2018 105

(Knighton 1998). It is also an assumption in the discharge-drainage area relationships that 

have been developed and used to guide and inform stream restoration efforts.

The most visible change once beaver dams are built is the increase in the amount 

of surface water stored within a given section of stream, but the increase extends 

beyond surface water to include groundwater. I modeled the potential increase in the 

total amount of water that would be stored in a 29 acre meadow with 0.76 miles of 

stream, and available to the river, using a point-in-time volume estimate of the amount 

of water held in the channel during the summer as baseline conditions. The reach occurs 

on Camp Creek, a tributary to the North Fork Burnt River (NFBR), located in eastern 

Oregon. The goal was to predict potential water-related benefits if beaver returned to 

the creek. Beavers are currently present in the NFBR but not in Camp Creek, though 

past signs of their presence exist and beaver habitat restoration efforts are underway. 

Commercial and recreational trapping currently occur in the NFBR watershed and the 

area is a mix of public and private land.

The calculations used existing cross-section data, information about summer water 

levels, an average bank height of 5 ft, and expected soil types (silt loam and sandy loam). 

The meadow was assumed to be one or the other soil type resulting in a range in values. 

Using information from Dunne and Leopold (1978), the soil porosity values used were 

0.46 for silt loam and 0.44 for sandy loam and the field capacity values used were 0.31 

for silt loam and 0.15 for sandy loam. These values were used to estimate the amount of 

groundwater that could be stored in the meadow under saturated conditions and capable 

of draining freely to the river. The calculations assumed that the dam heights equaled the 

average bank height.

Without beaver dams (baseline condition), the estimate of in-channel surface 

water was 0.4 acre-ft for the 0.76 miles of stream. Summer contributions from 

groundwater to base flows are limited because the channel is incised and over-wide. 

With beaver dams, the volume of in-channel surface water would increase from 0.4 

acre-ft to 9 acre-ft. A saturated meadow could contribute an additional 22 acre-ft (if 

silt loam) to 42 acre-ft (if sandy loam) of groundwater to the river, contributing to 

summer baseflows. If the dams extended across the valley floor and were a foot tall, 

an additional 29 acre-ft of surface water would be held in temporary storage. When 

all the values are totaled, the amount of water that could be stored within this 29-acre 

meadow and available to drain freely into Camp Creek if beaver dams were present 

increases from 0.4 to as much as 80 acre-ft of water.

A landscape-scale example of actual change in response to beavers expanding their 

range occurred between 1940 and 1986 on the Kabetogama Peninsula, a 17.4 mile2 area 

in Minnesota. Using aerial photos, Naiman et al. (1988) counted 71 beaver dams in 1940 

and 835 dams in 1986. The cover types were lumped into four categories (forest, wet, 

moist, ponds). In 1940, acres in the moist, wet, and pond categories were 640. By 1986 the 

number had increased to 9,308 acres with a large portion of that change achieved by 1961. 

Ponds increased from 40 acres to 3,388 acres reflecting large increases in surface water. 

The moist and wet categories increased from 600 to 5,921 acres reflecting a large increase 

in groundwater stored. Other examples exist within the literature that capture the water 

storing capability of beavers (Beedle 1991; Hood and Bayley 2008; Demmer and Beschta 

2008), leaving little doubt about the ability of abundant beaver populations to effectively 

help mitigate the effects of climate variability.
106 USDA Forest Service RMRS-GTR-377. 2018

Reductions in Downstream Flood Magnitudes

Beaver ponds have long been credited with reducing flood magnitudes and stream 

power through pond storage, valley-floor storage, or both (Dobyns 1981; Naiman et 

al. 1988; Parker et al. 1985)—assumptions that have been “based primarily on qualitative observations in the literature from the first half of the century” (Meentemeyer and 

Butler 1999). Actual studies quantifying the influence of beavers on flood magnitudes 

are few, have focused on small headwater streams, and have only considered the role 

of pond storage in flood peak reductions. Burns and McDonnell (1998) compared two 

stream hydrographs. One hydrograph captured flows in a 102 acre drainage containing 

a perennial stream and a single 3.2 acre beaver pond at its downstream end. The other 

hydrograph captured flows for a 151.2 acre watershed containing an ephemeral stream 

and no beaver pond. They found that the single pond provided minimal retention during 

several large runoff events. Beedle (1991) explored how storm hydrographs responded 

to increasing amounts of pond storage as the size and numbers of ponds in series increased. The drainages studied were 1,532 acres or less and his maximum pond size was 

1.48 acres. He found that the amount of reductions varied with storm size, pond size, 

pond numbers, and available storage capacity of the ponds prior to the flow event.

A single full beaver pond was found to theoretically reduce peak flows by no more 

than 5.3 percent regardless of the return interval or watershed size. The shape of the outflow hydrographs were the same as the inflow hydrographs, with only a 10 or 15 minute 

delay in the time to peak and slightly increased duration. Reductions in peak flows 

became increasingly large as the number of ponds in a series increased. Five large-sized 

beaver ponds in series reduced the storm peak flow by 14 percent for a 2-year event, but 

only 4 percent for a 50-year event (Beedle 1991).

However, the greater contribution of in-channel ponds as it pertains to reducing 

flood magnitudes is that they reduce the available channel capacity (ACC). This causes 

streams to overtop their stream banks at lower flows allowing flood waters to access the 

valley-floor where temporary flood storage is greater.

The degree to which overbank flooding decreases flood magnitudes and increases 

flood durations varies as a function of valley-floor roughness (Campbell et al. 1972; 

Leopold and Maddock 1954; Shankman and Pugh 1992), the amount of storage area 

(Campbell et al. 1972; Osterkamp and Costa 1987), and the location of unmodified 

sections of river with respect to the flood wave (Campbell et al. 1972; Hillman 1998). 

The mix of channelized and nonchannelized reaches results in a discontinuity in flood 

magnitudes, durations, and frequencies as a flow moves downstream. Some areas will 

experience increased flooding while others (e.g., downstream of a wetland) will show 

minimal changes for the same precipitation or dam-bursting event (Campbell et al. 

1972; Hillman 1998).

Reductions of flood magnitudes as a result of valley-floor storage in systems where 

the streams are hydrologically connected to their valley floors but without beavers have 

been documented by a number of researchers. Osterkamp and Costa (1987) estimated 

water depths at three valley cross-sections on Plum Creek in Colorado, which drains 

328.2 miles2. During a 900 to 1,600-year recurrence interval, flood water depths averaged 

from 7.9 to 9.5 ft but were as great as 19 ft. Dunne and Leopold (1978) examined runoff 

from four large drainage basins ranging in size from 7,411 to 203,000 miles2. They found 
USDA Forest Service RMRS-GTR-377. 2018 107

that the channel and valley floor stored 57 to 80 percent of the runoff generated by large 

storms. Campbell et al. (1972) used two flood-routing methods to determine the effect of 

channel straightening on flood magnitudes, durations, and attenuation of the flood peak for 

58 miles of the Boyer River in Tennessee. The river drains 1,188 miles2. Channel straightening and the building of dikes increased discharge downstream by limiting access to the 

valley floor and increasing the stream gradient. In contrast, under partial straightening they 

found that the unmodified sections of the stream substantially reduced the magnitude of 

flood peaks because the flood waters overflowed onto the valley floor. The floodplain in 

the unmodified section averaged 1.3 miles wide.

Campbell et al. (1972) reported the following results of their modeling: “The 

unmodified reach, even though short, provides tremendous storage, which can nullify 

the effects produced by the upstream straightening. As indicated in figure 8c, 16 miles 

of unmodified river reduced the increase in peak discharge from 90 percent to 15 pecent 

for the condition of high flood plain roughness coefficient. The increase in peak discharge at section 30 [the most downstream section and below the unmodified section] 

is 35 percent with high n [0.10] and 30 percent with low n [0.04] as compared with 

190 percent and 90 percent respectively for complete straightening.”

A smaller scale example comes from Rocky Creek in central Alberta. It drains 

an area of 4,621 acres. Hillman (1998) observed the influence of a 222.4-acre wetland 

on downstream flood flows after a beaver dam failed. The wetland contained a sedge 

meadow, willows, a small lake, and several beaver ponds. By the time the flood waters had 

passed through the wetland and reached the main gage, about 4 miles downstream of the 

dam failure, the flood peak was only 6 percent of the peak estimated to have entered the 

wetland. He concluded that wetlands, especially when large, are very effective at regulating high flows, even more so than beaver ponds because the dams often wash out during 

high floods.

These examples show that valley-floor storage, beaver ponds, and wetlands can 

significantly reduce flood peaks at all scales of drainage area and storm size. As such, the 

loss of wetlands and beaver ponds and the severing of stream-valley floor hydrologic connections has the reverse effect, increasing flood peaks for all sizes of storms and drainage 

areas. Nowhere has the impact of lost storage potential been more frequently on display 

than in the Mississippi River Basin. Hey and Philippi (1995) estimate that beaver ponds 

and wetlands historically made up 11 percent and 10 percent respectively of the 450 million acres of the Missouri and Upper Mississippi River Basins. They currently make up 

1 percent and 4 percent. The reduction of these features, along with the development of a 

channelized drainage network and levee system, has led to increases in the magnitude and 

frequency of downstream flooding with large economic and social costs.

Conceptual Model Part 2: Watershed Response to Beaver Trapping After 

a Long-Term Presence

The fluvial processes and sequence of events that occur when beavers are removed 

after a long-term presence, but prior to another large-scale disturbance, are shown in figure 20. Decreases or elimination of beavers from drainages means that dam failures are 

not repaired. The lack of dam maintenance at multiple sites within a drainage and across 

a region sets in motion changes that allow the channels to expand beyond the point of 
108 USDA Forest Service RMRS-GTR-377. 2018

Figure 20–Conceptual model of how beaver trapping or site abandonment influence fluvial systems (source: Fouty 

2003).

BEAVER TRAPPED OUT OR ABANDONED AN UPPER WATERSHED DRAINAGE

Dams leak

and ponds

drain.

Dams do NOT 

leak.

May set in motion 

complex response of 

aggradation and 

incision and widening 

downstream of failure.

Discontinuous 

channels begin to 

coalesce and drainage 

network expands.

Incision crosses a 

tributary junction.

Exposed sediments 

begin to vegetate.

Sediment behind 

dams erode.

Localized 

channelization and 

concentrated flow.

Pond drains over time. 

Available channel 

capacity increases

Possible 

downstream 

scour.

Dams do not 

breach.

Ponds drain abruptly. 

Available channel 

capacity increases.

Dams fail abruptly.

Tributary junction 

experiences a 

base level drop 

and begins 

headcutting.

Development of 

discontinuous channels 

on main stem. Some 

channel widening.

Sections downstream of 

dam failure aggrade due to 

deposition of eroded 

sediment where sediment 

trapping mechanisms exist.

DOWNSTREAM EFFECTS

Vegetation 

becomes denser

around old pond

sites.

Discontinuous channels do NOT 

coalesce because they encounter 

reaches with high resistance to erosion.

Drainage network a mix 

of channelized and 

nonchannelized reaches

Knickpoint forms at breach.

Dams fail over 

time.

Foraging around the 

pond stops.

Incision does NOT cross 

a tributary junction.

Increase in stream velocities, 

increase in erosive power.

Increase in speed at which 

water is transferred from 

upper to lower watershed.

Increase in flood magnitudes and 

frequency of large magnitude floods in 

lower watershed due to concentration of 

flows in upper watershed channels.

Decreased water quality (i.e. sediment loads, 

warmer low flow stream temperatures).

Stream-valley floor hydrologic connection 

decreases. Frequency of valley flooding 

decreases. Water table and low-flow base flows 

drop. Sediment once underwater now exposed. 

Downstream of a wetland 

or unmodified channel

reach minimal changes in 

flood frequencies or 

magnitudes.

Downstream of 

incised, widened

channel sections

flood frequencies 

and magnitudes

increase for all 

sizes of 

precipitation events.

Meadows, Bogs, or 

Wetlands form

creating areas with 

high resistance to 

erosion.

Continue 

to pond 

water.

Fill with 

sediment.
USDA Forest Service RMRS-GTR-377. 2018 109

dam failure. With failures, (1) ponds drain and sediment erodes, (2) available channel 

capacity (ACC) increases, (3) local base-levels drop, (4) knickpoints form, and (5) the 

stream-valley floor hydrologic connection becomes increasingly disconnected.

The impact of beaver trapping and subsequent dam failures on ACC and thus the 

stream-valley floor hydrologic connection was captured at Price Creek in southwestern 

Montana (Fouty 2003). Fourteen monumented cross-sections were established in 1995. 

Twelve occurred inside a cattle exclosure that contained beaver dams. Two occurred 

downstream of the exclosure in a section without beaver dams. In 1995, the percent 

ACC at the dam-controlled cross-sections ranged from 0 percent (channel full) to 

62 percent depending on dam integrity (table 2), which was already varying as a result 

of beaver being trapped out between 1994 and 1995. By 1998 all dams were failing or 

gone and the percent ACC had increased at all sites.

Table 2—Changes in available channel capacity (ACC) as a result of dam failure and pond drainage, Price Creek, Montana. 

Cross-sections 17 and 18 were downstream of the beaver dam controlled reaches (source: Fouty 2003.)

Cross-section 

number

% Available channel 

capacitya

Dominant reason for 

the change in available 

channel capacity

Distance upstream of a beaver dam (ft) 

and timing of dam failure

1995 1998

19 0 69 Pond drains 115 ft. Begins breaching post 1995 but still exerting 

some influence in 1997 and 1998. 

20 0 84 Pond drains and 

sediment erodes

30 ft. Begins breaching post 1997 but still exerting some 

influence in 1998.

21 38 81 Pond drains 39 ft. Begins breaching post 1997 but still exerting some 

influence in 1998. 

22 39 91 Pond drains 52 ft. Begins breaching post 1997 and gone by 1998.

23 17 68 Pond drains and 

sediment erodes 

49 ft. Begins breaching post 1997 but still exerting some 

influence in 1998. 

24 26 72 Pond drains and 

sediment erodes 

10 ft. Begins breaching post 1995. Completely breached 

by 1998. 

25 66 84 Pond drains 62 ft. Intact in 1997, gone by 1998.

26 63 80 Pond drains 89 ft. Begins breaching post-1997 but still exerting some 

influence by 1998. 

27 35 71 Pond drains 36 ft. Begins breaching post-1997 but still exerting some 

influence by 1998.

28 62 84 Pond drains 72 ft. Begins breaching post 1995 but still exerting 

influence in 1997 and 1998. Below this dam was a 

submerged dam acted as a base-level control.

29 62 94 Pond drains and 

sediment erodes

10 ft. Begins breaching post 1995. Gone by 1998. 

30 69 95 Pond drains and 

sediment erodes

10 ft. Begins breaching post 1995. Some influence still in 

1997 but gone by 1998.

No beaver dam influence

17 94 94 Minimal changes in 

sediment or water levels

n/a

18 87 89 Minimal changes in 

sediment or water levels

n/a

a % Available channel capacity = (1-(water XS area/channel XS area)) *100.
110 USDA Forest Service RMRS-GTR-377. 2018

Changes in ACC between 1995 and 1998 inside the cattle exclosure varied as a 

function of proximity to a dam, dam integrity, and the length of time the dam had been 

in place prior to the building of upstream dams (Fouty 2003). ACC increased at all 

12 cross-sections. Seven cross-sections had increases due to pond drainage, and five 

cross-sections due to sediment erosion and pond drainage. In contrast, the ACC at the 

two downstream cross-sections (17, 18) changed very little despite the large influx of 

sediment from upstream sites as dams failed and the sediment previously trapped behind 

them eroded. The lack of sediment trapping mechanisms at these two cross-sections 

resulted in sediment moving through the system as suspended load, giving the appearance of a sediment-starved system rather than one simply lacking sediment trapping 

mechanisms.

Unlike Price Creek where repeated beaver trapping has prevented the development 

of wetlands, streams with a long-term beaver presence and abundant sediment inputs 

would in time develop complex wetland vegetation communities (Johnston and Naiman 

1990). Post-trapping but in the absence of another disturbance, these well-established 

wetlands would persist after the loss of beavers and impede the headward migration of 

the knickpoints initiated when the dams failed. Their location, characteristics, and stability would prevent the channels from coalescing into a single interconnected system. 

The resulting drainage pattern would be one in which channelized reaches are spatially 

separated by wetlands. The hydrologic impact of this drainage pattern would be a discontinuity in flow magnitudes and durations, similar to that described by Campbell et al. 

(1972) and Hillman (1998).

The effectiveness of the wetlands at impeding channel development comes from 

(1) their enhanced roughness that reduces flow velocities (Cooke and Reeves 1976; 

Hendrickson and Minckley 1984), (2) temporary storage potential that reduces flood 

peaks (Campbell et al. 1972; Dunne and Leopold 1978; Hillman 1998; Osterkamp and 

Costa 1987), and (3) their enhanced subsurface cohesion (Cooke and Reeves 1976; 

Finch et al. 2003; Hendrickson and Minckley 1984; Smith 1976). If, on the other hand, 

the wetlands develop channels or are eliminated by human activity, the drainage pattern 

becomes increasingly channelized (fig. 20). Examples of the amount and speed at which 

channelization can occur are found in table 3.

As a drainage becomes channelized, vegetation shifts from water-dependent to 

drought-tolerant species. This shift occurs as a result of decreased soil moisture. As 

a channel incises, widens, or both, the increased ACC leads to a reduction in the frequency of valley-floor flooding (Campbell et al. 1972; Schumm et al. 1984; Shankman 

and Pugh 1992) and enhances the flow of groundwater toward the channel by steepening 

the hydraulic gradient between the water table and stream (Knighton 1998 referencing 

Dunne 1980, 1990). The result is a lowering of the water table.

The speed at which vegetation changes varies as a function of climate, land use, 

incision depth, groundwater depth, subsurface stratigraphy, and vegetation requirements (Cooke and Reeves 1976). In areas where precipitation is distributed throughout 

the growing season, it may partially compensate for a decline in water tables if its 

abundance and distribution are sufficient to maintain soil moisture levels. In areas 

where precipitation is strongly seasonal, declines in water tables are not compensated 

for by precipitation, and vegetation responds more quickly to channelization and the 

lowering of the water table (Bryan 1928b). The speed at which vegetative communities 
USDA Forest Service RMRS-GTR-377. 2018 111

Table 3—Examples of the speed at which channelization occurs, and the depth, width, and length of the channelization 

when known (source: Fouty 2003.) 

Location Dates Type and amount of change Time interval Source

Rio Salado, NM Between 1882 

and 1918

Channel widens from 11.88 to 48.84 ft wide to 

330.1 to 550 ft wide.

<36 years Bryan 1927

Felipe Gilbert 

Creek, NM

One storm event Channel headcuts for a distance of 40 to 75 ft. 1 day Bryan 1927

Whitewater Draw, 

AZ

The rainy season 

in 1910

Channel headcuts for a distance of 0.25 miles. Up to a 

couple of 

months

Cooke and 

Reeves 1976

Kanab Creek, UT Between 1883 

and 1885 

Channel incises 60 ft and widens nearly 70 ft for a 

distance of 15 miles.

<2 years Gregory 1917

Walker Creek, AZ Between 1894 

and 1913

Channel incises 80 ft deep. <19 years Gregory 1917

Chinle Creek, AZ Between 1894 

and 1913

Channel incises 100 ft deep. <19 years Gregory 1917

Mountain 

Meadows, UT

1884 – in one 

series of storms

Channel incises into what was once a wet 

meadow during a series of storms and continued 

to widen after 1884. Gullies fingering out to 

nearly all parts of meadow. No numbers given.

Up to one 

month 

for initial 

incisions

Cottam and 

Stewart 1940

Crane Creek, OR Between 1925 

and 1930

Channel incises to a depth of 10 ft. 5 years Schaffer 1941

Santa Cruz River 

near Tucson, AZ

Between Aug 5 

and 9, 1890

Channel incises some unknown depth for 1.5 

miles between August 5 and 7. Between August 

7 and 9 channel begins to fork and headcut in 

multiple directions.

4 days Cooke and 

Reeves 1976 

(p. 51) 

Sonoita Creek, AZ Between 1891 

and 1912

Channel incises 18 to 20 ft deep and widens to 

250 ft.

<21 years Bryan 1928b

Gila River near 

Safford, AZ

Between 1905 

and 1917

Channel widens from an average of less than 330 

to 2000 ft for about 45 miles.

<12 years Burkham 1972

Cimarron River 

in southwestern 

Kansas

Between 1874 

and 1939

Channel widens from average of 50 to 1200 ft for 

about 175 miles.

 <65 years Schumm and 

Lichty 1963

Rio Puerco, NM 

(between mouth 

and Cabezon)

Between about 

1885 and 1892

Channel incises and the incision migrates 

upstream for 110 miles Discontinuous incision 

existed prior to 1885 and this may have facilitated 

rapid headward migration of the incision.

<7 years Bryan 1928a

Douglas Creek, CO 

(East Fork)

Between 1882 

and 1900

Channel has incised 16.4 ft. <18 years Womack and 

Schumm 1977

Wolf River near 

Memphis, TN

Between 1964 

and 1999

Channel incises migrates upstream for 10.6 miles. 

Headcutting is episodic in nature with an average 

rate of headward migration of 0.37 miles/yr. Some 

areas have had a 19.7 ft drop in bed level and the 

channel has widened to twice its original width.

<35 years Wiens 2001

Price Creek, MT 1995 and 1998 Channel incises 0 to 2.6 ft deep for sites less than 

50 ft upstream of beaver dams as a result of dam 

failures.

1 to 3 years Fouty 2003

Obion River, TN Post 1960s Channel has undergone headward migration 

of knickpoints as much as 0.62 miles/yr and 

channel widening as much as 3.28 ft/yr due to 

channelization of the river by U.S. Army Corps of 

Engineers in the 1960s.

1 year Shankman and 

Pugh 1992
112 USDA Forest Service RMRS-GTR-377. 2018

Table 4—Examples of the speed and character of vegetation changes as a result of channel incision. Unless beaver trapping or 

area abandonment is explicitly mentioned, the cause of the incision is Euro-American settlement activities (source: Fouty 

2003.)

Location

Time 

interval Vegetation change Total time Source

Santa Cruz 

River near 

Tucson 

1880 to 

1928

From area covered by sacaton grass with groves of mesquite and 

swampy areas of tule (bulrushes) prior to 1880 to dense mesquite 

forest by 1928. Arroyo forms in 1880.

Less than 48 

years

Bryan 

1928b

Sonoita River 

of Sonora 

pre-Aug 6, 

1891, to 

1928

From swampy area prior to August 6, 1891, to a dense mesquite 

forest by 1928. Arroyo forms in August 6, 1891. 

Less than 37 

years

Bryan 

1928b

Yancy 

Meadows, 

Yellowstone 

NP

1903 or 

1904 to 

1921

Beavers began to desert area in 1903 or 1904. By 1912 the colony 

was abandoned. Changes from ponds to well-formed meadows 

to solid ground by 1921 with little evidence of the earlier beaver 

ponds.

17 or 18 

years

Warren 

1926

Crane Creek, 

OR

1925 to 

1936

Beavers trapped out in 1924. Channel incises in 1925 and 

vegetation changes from meadows of “stirrup-high native” grasses 

subirrigated by beaver ponds to meadows nearly gone, with 

clumps of new sagebrush and sparse remnants of the original 

grasses by 1936.

11 years Schaffer 

1941

Near Little 

Summit 

Ranger 

Station area, 

OR

1925 to 

1929

Area was formerly full of beavers, but the last appear trapped out 

by 1925. “From that date to 1929 (4 years) the old ditch and the 

entire meadow were fast becoming a dust bed. During 1928 and 

1929 no water ran out at the lower end of the station”

4 years Bailey 

1936

Mountain 

Meadows, 

southern UT

1884 to 

sometime 

prior to 1900

Channel incises into what was once a wet meadow during a series 

of storms and continues to widen since 1884. Gullies fingering 

out to nearly all parts of meadow. Shift in vegetation from a wet 

wiregrass meadow surrounded by numerous springs and a dry 

grass meadow as meadows drain to desert shrub. 

< 16 years Cottam 

and 

Stewart 

1940

change further accelerates where land uses increase runoff rates, decrease infiltration 

rates, surface roughness, and stream-bank vegetation, and damage soil structure. A 

self-enhancing feedback loop is triggered as increased runoff and loss of bank stability 

and floodplain roughness facilitate channel enlargement during high flows furthering 

a lowering of the water table and a shift to more drought tolerant, less densely rooted 

vegetation. This change to more drought-tolerant species in turn sets another feedback 

loop in motion by decreasing the water-holding capability of the soil as below-ground 

root biomass declines (Fitch et al. 2003). In the examples presented in table 4, major 

changes occurred between 4 to less than 50 years, the time frames constrained by the 

next observation.

The conceptual model represented by figures 19 and 20 is the backdrop for the 

remaining sections that explore why beaver trapping as a regional EA disturbance, and 

beavers as a major component of stream systems, were missed by early researchers and 

remained absent from the discipline of fluvial geomorphology until recently. This chapter concludes with a discussion of how adding beavers and EA beaver trapping back 

into the story of EA disturbances changes the discipline of fluvial geomorphology and 

current stream restoration efforts.
USDA Forest Service RMRS-GTR-377. 2018 113

The Geographies of Euro-American Beaver Trapping, General Land 

Office Surveys, and Early Expeditions

Euro-American (EA) beaver trapping was temporally and spatially concentrated. 

It began in the 1600s on the East Coast and along the Mississippi and Missouri Rivers, 

in the early 1700s on the West Coast, and into the Interior United States in the late 

1700s and early 1800s (Phillips 1961). The arrival of trappers in an area predated most 

settlement and scientific and military surveys by at least several decades, with a few 

exceptions. One exception is the East Coast where settlement and trapping co-existed 

in time (Cronon 1983) and numerous writings exist from the 1600s and 1700s on the 

local natural history of those areas (Meisel 1924). Other exceptions are the Lewis and 

Clark (1804 to 1806), Long (1819 to 1820), and Pike (1805 to 1807) expeditions in the 

West (Phillips 1961). These written observations, combined with later trappers’ journals 

and records from fur companies, reveal complex, multi-channeled rivers abundant with 

beavers and beaver dams.

Yet, it was not these early records but those of the later GLO surveys and scientific and military expeditions that have been used to reconstruct the geomorphic and 

ecological characteristics of watersheds prior to settlement. Embedded in this reliance 

on the GLO notes for pre-settlement conditions has been an unspoken assumption that 

“the public land surveys were carried forward in virgin territory—unexplored and unmapped—in advance of settlement” (Clements 1985). However, regardless of the area 

examined, changes to the drainage network were well underway by the time the GLO 

surveys and various expeditions arrived, with the degree of change observed influenced 

by the number and type of EA disturbances that had occurred post-trapping.

The GLO surveys began in 1785 with the passage of the Land Ordinance. The 

first survey took place in Ohio in 1785 with subsequent surveys proceeding westward 

in response to pending EA settlement (Clements 1985; White 1996). It was a formalized gathering and storing of information. However, even by 1785 the area east of 

the Missouri and lower Mississippi Rivers had already been heavily trapped (Phillips 

1961). In the New England landscape, beavers had ceased to be a dominant feature 

as early as the late 1600s (Cronon 1983). By the late 1700s to early to mid-1800s the 

beaver dams had been replaced by thousands of water-powered mill dams in New 

England and the Mid-Atlantic States. The mill dams trapped sediments eroding off the 

valley and hillslope in response to agriculture and logging, burying the beaver-created 

wetlands (Walter and Merritts 2008). In the Southwest and Intermountain West, trapping and the GLO surveys were more coincident in time but still separated by a decade 

or more (table 5). Some early researchers in the West acknowledge the occurrence of 

beaver trapping in their study areas and its implications (Dobyns 1981; Gregory 1917; 

Gregory and Moore 1931; Hendrickson and Minckley 1984; Leopold 1951), but the 

loss of beavers as a regional disturbance with major geomorphic and hydrologic significance was missed.

The temporal differences between when trapping occurred versus the GLO surveys 

and early expeditions was compounded by differences in their spatial geographies. 

Trappers followed streams in their search for beavers. In contrast, the GLO surveys 

recorded information about the land and its resources along linear grid lines spaced 
114 USDA Forest Service RMRS-GTR-377. 2018

Table 5—Estimated timing of beaver trapping, observations of discontinuous arroyos and incised tributaries, and baseline 

GLO survey (Source: Fouty 2003.) 

Site: San Pedro River, AZ

Dates trapped: 1826-1827 (Pattie 1831)

Pattie and his party trap the river in March 1826 and take 200 beavers. They trap the river again in October 1827. 

No numbers given for the second time.

Next observation: Military expeditions: 1846, 1852, 1859 (Leopold 1951).

1846 – Description of vegetation patterns in area (Johnston 1847).

1852 near Pomerene: The stream banks not less than 8 to 10 feet high (Bartlett 1854).

1859 – There is a discontinuous gully near Pomerene. The river has a “width of about twelve feet and a depth of 

twelve inches [water depth], flowing between clay banks ten or twelve feet deep, but below it widens out and 

from beaver dams and other obstructions overflows a large extent of bottom land, forming marshes densely 

timbered with cottonwood and ash Hutton (1859).”

Estimated time between trapping and next observation: 19 to 20 years.

Baseline GLO surveys: 1851, 1865, 1867 (White 1996).

Site: Diablo Range, CA about 10.6 to 15.5 miles west of the San Joaquin River (Bull 1964)

Dates trapped: 1829 to 1843 (Phillips 1961).

Hudson’s Bay Company trapped in California beginning in 1829 until 1843, returning “every year to trap the 

Sacramento-San Joaquin River systems and the area around the San Francisco Bay (p. 544).” The company took 

from the Bay area alone 10,860 beaver between 1830 and 1839.

Next observation: GLO surveys in early 1850s (Bull 1964).

GLO surveyors noted the existence of “traces of older gullies on some of [alluvial] fans indicate that entrenched 

channels existed before sheep were brought into California in 1853 and before large-scale cattle ranching was 

introduced in western Fresno County.”

Estimated time between trapping and next observation: 9 to 25 years

Baseline GLO surveys: 1852 to 1854 (Cooke and Reeves 1976).

Site: Rio Puerco, NM, a tributary to the Rio Grande (Bryan 1928a)

Dates trapped: 1823 to about 1838 (Weber 1971).

“In 1823, however the fur trade from New Mexico had scarcely begun....most trappers certainly centered their 

operations on the virgin streams of the Pecos and Rio Grande valleys. The beaver supply in this convenient area 

was already being depleted” and by 1824 trappers were heading west. In 1827 American fur trappers were 

floating down the Rio Grande trapping as they went. 1832 to 1838 trapping occurs around the settlements 

along the Rio Grande valley.

Next observation: Military expedition 1846 to 1847, 1849 (Bryan 1928a).

Abert (1847): banks were 10 or 12 feet high and vertical at a point west of Albuquerque. Banks were 30 feet 

further upstream near a ruined town. Simpson (1849): channel was 100 feet wide, contained stagnant pools 

of water; banks were 20 to 30 feet high about 5 miles above Cabezon (small village on the river). Late 1880s: 

many settlers testify that in many places the river had no banks or only small ones and in flood the river spread 

out over the entire valley floor.

Estimated time between trapping and next observation: 9 to 23 years

Baseline GLO surveys: 1855 (Bryan 1928a).

Site: Non-specified tributaries in the Colorado River region (Dellenbaugh (1912)

Dates trapped: 1824 to probably late 1830s (Chittenden 1954; Phillips 1961).

All the major tributaries of the Colorado River were trapped.

Next observation: The Powell expedition of 1871 or 1872 (Gregory and Moore 1931).

“I noted the same characteristics [trenching of stream beds] (and others probably also noted) years ago in places 

where there were no cattle and never had been”..... “I have seen earth-cliffs 30 to 40 feet high with all the 

characteristics of a rock-cliff erosion” (Dellenbaugh 1912).

Estimated time between trapping and next observation: 33 to 47 years.

Baseline GLO surveys: New Mexico: 1869; Arizona: post-1867; Utah: post 1855; Colorado: post 1880 (White 1996).
USDA Forest Service RMRS-GTR-377. 2018 115

1 mile apart. The GLO surveys focused first on those areas that were about to be settled 

or were in the process of being settled by Euro-Americans, leaving large portions of 

each State unsurveyed until later (Clements 1985; White 1996).

An example is the early military expeditions into Arizona and New Mexico. They 

entered the Gila River drainage from southern New Mexico via the Lordsburg Plain 

(Leopold 1951) making their observations of pre-settlement channel incision restricted 

to the middle and lower Gila River Basin (e.g., San Pedro River, the Santa Cruz River). 

The upper tributaries (e.g., San Francisco River, East and West Forks of the Gila River) 

were bypassed and changes unrecorded. Yet references to discontinuous arroyos ending 

at wetlands on the San Pedro River in 1846 (Cooke and Reeves 1976; Hastings and 

Turner 1965) suggest that these features would have existed in the upper watershed as 

well because the entire basin had been trapped between 1826 and 1834 (Pattie 1831; 

Weber 1971).

I am familiar with the upper Gila watershed, having collected stream data on 

several of its tributaries in the 1990s. These tributaries contain the appropriate valley 

characteristics for beaver dam complexes and beavers and I observed beaver sign on the 

East Fork of the Gila in 1993, 1994, and 1995. Therefore, had the GLO surveyors and 

early military and scientific expeditions explored the upper watershed, they likely would 

have observed areas where abundant beaver dam complexes and their relationship to 

wet meadows persisted. These observations would have led future researchers to ask 

different questions and come to different conclusions about the causal mechanisms leading to arroyo development.

The limited and selective exploration of the Southwest prior to EA settlement is an 

example of what Graf (1984) referred to as a “spatial bias.” He saw this bias as “a major 

hazard in geomorphic theory development because of the relatively small size of the 

geomorphic research community.” The limited number of researchers means that “individual scientists can affect the development of theory with relatively few publications, 

and therefore the field origins [emphasis added] of those few publications [or observations] assume disproportionate importance” (Graf 1984). This is what happened in the 

case of the spatially limited but high quality GLO surveys and early expedition records 

in these areas post-trapping. The fragmented but still visible influence of a once-abundant beaver population and the impact of recent trapping on drainage stability appeared 

localized. Thus their significance at a regional scale was missed by later researchers who 

would utilize these records.

Arroyo Formation in the Southwest and Intermountain West

The Role of Beaver Trapping

The presence of pre-historic arroyos and pre-EA settlement arroyos in the 

Southwest and Intermountain West (Balling and Wells 1990; Love 1979) has been 

central to the question about whether EA livestock grazing, climate change, randomfrequency events, or some combination was the dominant causal mechanism that led to 

widespread arroyo development after EA settlement (Cooke and Reeves 1976). Their 

presence led some early researchers to suggest that climate was the dominant causal 

mechanism and EA livestock grazing merely a “trigger pull which timed a change about 
116 USDA Forest Service RMRS-GTR-377. 2018

to take place” (Bryan 1928a). This perspective suggested that streams in these regions 

were inherently sensitive to climatic variability.

The GLO surveys and early expeditions recorded two categories of features: one 

indicative of a stable fluvial system (wetlands, wet meadows), the other indicative 

of a destabilized fluvial system (discontinuous arroyos, actively incising tributaries), 

often within the same drainage (Cooke and Reeves 1976; Hastings and Turner 1965; 

Hendrickson and Minckley 1984). The active nature of the channel incision suggested 

that the destabilization had been fairly recent. Determining whether a long-term 

beaver presence followed by trapping could explain this mix of observations, a direct 

relationship between EA trapping and arroyo formation was considered supported if 

(1) trapping occurred in the area, (2) the time interval between trapping and the next 

observation (i.e., 15 years) was longer than the time needed for substantial channel 

incision to occur (i.e., <10 years), and (3) the magnitude of the observed channelization could have occurred within the intervening time (Fouty 2003). In this case, 

beaver dam failures and non-repair would explain the channelization observed. When 

the interval between trapping and the next observation was shorter than 10 years, 

this suggested that large precipitation events and localized dam failures may have 

interacted synergistically to enhance large floods, thereby accelerating the rate of dam 

failures and channel incision. A literature review found that the features observed by 

the surveys and expeditions in the Southwest and Intermountain West occurred either 

in areas directly mentioned as being trapped or in the general area. The temporal 

separation between trapping and the next observations of stream conditions and site 

characteristics was 9 to 47 years (table 5).

The widespread and rapid removal of beavers throughout these areas led to the 

development of multiple base-level drops within individual drainages as dams failed and 

were not repaired. The impact of multiple base-level drops on channel development and 

drainage network expansion would have been amplified by the period of above-average 

precipitation from 1835 to 1849 observed in the tree-ring data (D’Arrigo and Jacoby 

1991; Meko 1990; Meko et al. 1991). This period occurs shortly after trapping ceased in 

these areas but before the GLO surveys and early expeditions arrived. A second period 

of above-average winter precipitation, also identified in the tree-ring data, occurred from 

1905 to 1920, or 1928 depending on the tree-ring chronology (D’Arrigo and Jacoby 

1991; Meko 1990). This second wet period post-dated not only beaver trapping but 

settlement, widespread livestock grazing, and other land use activities. It is this second 

period of above-average precipitation that contributed to the large floods that widened 

and incised many streams in the West (Burkham 1972; Cooke and Reeves 1976), leading to a further expansion of the channelized network.

The early researchers who sought a causal mechanism to explain the pre- and 

post-settlement arroyos and changes in watershed hydrology did not understand the 

significance of the remnant populations of beaver and their scattered dams. Nor were 

they aware of the period of above-average precipitation that had occurred from 1835 to 

1849. When impacts of the two events are combined, a different, more complex causal 

mechanism for the observations recorded by the GLO surveys and early expeditions and 

the later increase in flood frequencies and magnitudes appears. The causal mechanisms 

expand to include a long-term beaver presence followed by beaver trapping followed by 

above-average precipitation. These series of events explain not only the discontinuous 
USDA Forest Service RMRS-GTR-377. 2018 117

arroyos and actively incising tributaries, but the occurrence of drainages in which the 

arroyos terminated at the base of a wetland.

A long-term beaver presence as an explanation for the wetlands (or cienégas) 

found in the Southwest is supported by Hendrickson and Minckley (1984). They found 

that mid-elevation wetlands characterized by permanently saturated, highly organic reducing soils occurred where (1) groundwater intersected the surface, (2) discharges were 

stable, and (3) flood peaks were low, minimizing the potential for scouring flows and 

channel incision. The features they identified causing groundwater to intersect the surface included upfaulted bedrock, changes in base level of the receiving stream, stream 

impoundments by landslides, and the development of a convex-concave profile. In their 

discussion of the convex-concave profile, they identified two mechanisms leading to 

its development: the deposition of coarse sediments and the placement of beaver dams 

along the stream. However, the influence of beaver dams extends beyond the profile. 

Similar to landslides, though on a smaller scale, beaver dams impound streams, provide 

local base-level control, and raise the base level of the channel by reducing ACC as 

they pond water and trap sediment. Though not as stable as bedrock, as long as beavers 

are present in the system to repair the dams, the dams will operate as a “continuously 

renewed, erosionally resistant substrate” (Parker et al. 1985).

The long-term presence of beaver dam complexes also creates stable wetlands. 

These wetlands are highly resistant to climatic variability and disturbance (Ives 

1942; Naiman et al. 1986, 1988), and can have long residence times on the landscape 

when undisturbed (Hendrickson and Minckley 1984; Ives 1942; Naiman et al. 1988; 

Warren 1926). The dam complexes provide the two other requirements identified 

by Hendrickson and Minckley for wetland development (low flood peaks and stable 

discharge). Large ponds and wetlands in headwater streams have been observed to effectively dampen the effects of both large runoff events and prolonged drought (Grasse 

and Putman 1956; Hillman 1998; Hood and Bayley 2008). Beaver dams decrease flood 

peaks by storing water in the ponds and reducing ACC such that during high-flow 

events there is rapid access to the valley floor where potential flood storage and roughness are greater. In turn, the increased frequency of valley-floor flooding leads to higher 

water tables that stabilize base flows, minimizing the impacts of drought.

Once trapping occurred, the dams were no longer maintained and they ceased to 

act as “continuously renewed, erosionally resistant substrate” akin to bedrock. Instead 

they became points of base-level drop and knickpoint initiation. However, the greater 

resistance of wetlands to incision, compared to the dam sediments, would have effectively halted the headward migration of a knickpoint generated by a base-level drop 

downstream. The result was the development of the observed discontinuous channels 

interspersed with wetlands. It is possible that these drainages may have remained a mix 

of channels, ponds and wetlands without further EA disturbances. The wetlands and lush 

grasses and willows along channels would have maintained channel and stream bank 

stability and thus kept channels narrow, the stream-valley floor hydrologic connection 

high, and downstream flood peaks dampened.

Other areas within the West showed patterns and sequencing of change similar to 

that found in the Southwest. An example is the work done by Buckley (1992) on Camp 

Creek, a tributary to the Crooked River, and located in central Oregon. He found that 

the area had been trapped between 1824 and 1830 by Peter Skeen Ogden and his party 
118 USDA Forest Service RMRS-GTR-377. 2018

and his journals reference plentiful beavers, willows, and aspen (Buckley 1992). Later 

records from military expeditions between 1858 and 1864 note lush grasses, willows, 

swampy areas, and abundant beavers and beaver dams along Camp Creek. By 1876, 

references to beavers are absent in the GLO notes, though they still mentioned the 

presence of large swampy areas along Camp Creek and narrow channel widths (10 to 

33 ft). However, the GLO notes also mention that several homesteaders had been living along Camp Creek and its tributaries since 1871, which may account for the lack 

of beavers. Livestock were in the area by 1876 with numbers increasing into the early 

1900s. By 1905 Camp Creek had incised 25 ft and widths had increased to 60 to 100 ft 

wide (Buckley 1992).

Not all areas had wetlands mixed with arroyos and incising tributaries. Places 

where the GLO notes or early expeditions mention arroyos but not wetlands include the Colorado Plateau (Dellenbaugh 1912), the Zuni River in Arizona (Balling 

and Wells 1990), the Rio Puerco in New Mexico (Bryan 1928a), and the Diablo 

Mountains in California (Bull 1964). F.S. Dellenbaugh, a topographer on the Powell 

expedition of the Colorado River area in the 1870s (Gregory and Moore 1931), mentioned seeing earth-cliffs bordering unnamed tributaries in this area in the 1870s that 

were 30 to 40 ft high. He observed these in areas where EA livestock grazing had not 

yet reached. He suggested that the tributaries were responding to a drop in base level 

on the main stem that had occurred for some unknown reason (Dellenbaugh 1912). 

The Colorado Plateau had been trapped in the 1820s and perhaps as late as the 1840s 

(Phillips 1961). Dellenbaugh’s observations in the 1870s (30 years later) would be 

consistent with beaver-dam failures on tributary streams triggering multiple points 

of base-level lowering. It is possible that wetlands were present, but not observed. It 

is also possible that topography, the presence of Native Peoples and their villages, or 

both, eliminated wetlands prior to the next EA observation.

The Role of Native Peoples

The influence of Native Peoples on arroyo development and the presence or 

absence of wetlands cannot be ignored. The difference between the impacts of beavers 

and beaver trapping versus Native Peoples’ efforts may simply be one of aerial extent. 

Native Peoples would have influenced the areas around their villages creating localized 

changes (Dobyns 1981) while beavers would have had a wider distribution. Reagan 

(1924) observed that “every side-wash, canyon and flat had its village or villages, its 

dams, ditches and reservoirs, as is readily seen by examining the region.” He argued 

that irrigation systems and check dams built by Native Peoples were responsible for the 

development of ponds, wetlands, and aggrading surfaces. Reagan’s (1924) descriptions 

of height, composition and locations of the check dams are similar to beaver dams. The 

check dams were composed of earth, about 5 ft tall and, like beaver dams, would have 

required constant maintenance. Dobyns (1981) also supported the hypothesis that Native 

Peoples played a large role in reducing erosion via their check dams. He felt that their 

reduction in numbers due to contact with EA diseases and conflicts would have led to 

decreased dam maintenance and renewed erosion as the dams failed. The changes in 

the drainage network and hydrology would have been similar to those that occurred due 
USDA Forest Service RMRS-GTR-377. 2018 119

to widespread beaver dam failures. Some of the dams attributed to Native Peoples may 

thus have been beaver dams given their wide distribution prior to EA trapping.

The Zuni River, located in Arizona and a tributary to the Little Colorado River, 

provides possible insight into this question of Native Peoples’ influence because documentation exists of arroyos that pre-date both Spanish and EA activity. These early 

arroyos date from about 1680. Tree-ring dates indicate that the Zuni River had eroded to 

its present level by 1776 when Fray Dominguez observed an arroyo adjacent to the Zuni 

Pueblo as well as arroyos upstream of the pueblo (Balling and Wells 1990). Balling and 

Wells (1990) used modern precipitation records (1897 to 1985) and post-settlement arroyo development to analyze potential causes of early arroyo formation. They suggested 

links between arroyo formation and changes in local precipitation patterns, particularly 

precipitation intensities. However, these very early arroyos were in the vicinity of a 

pueblo, complicating the direct link to precipitation. The Native Peoples may have 

deliberately or unintentionally altered some feature of the landscape that caused arroyos 

to form, such as the digging of an irrigation trench, failure of a check dam, or removing 

beavers from an area. Grazing by Native Peoples’ sheep and horses, which had arrived 

with the Spanish (Love 1916), may have also contributed to arroyo formation by reducing upland vegetation around the pueblo. References from the 1847, 1849, and 1852 

expeditions into this area and elsewhere in eastern Arizona mention the presence of numerous Indian horses and sheep and areas with limited forage for their mules (Leopold 

1951). The link between arroyo expansion in the Zuni River area and EA beaver trapping after 1849 is clearer because the timing of trapping in the area and the period of 

above-average precipitation (1835 to 1849) are known.

The Little Colorado River Basin where the Zuni River occurs was trapped for beavers in the 1820s and 1830s (Gregory and Moore 1931; Phillips 1961). Leopold (1951) 

cites references to the presence of beaver lodges in 1852 on the Little Colorado River 

slightly upstream from the town of Holbrook. This would place the lodges roughly 36 

linear miles downstream of the confluence of the Zuni River and the Little Colorado 

River. In addition, the 1852 expedition observed in one place on the Zuni River below 

the village of Zuni (now a dry wash) “a few populars... and near these trees was a 

beaver-dam” (Leopold 1951). While the observation of populars [poplars] and beaver 

dams in this area may have been a rare sight in 1852, it indicates that beaver were 

present in the area. Their loss would have made streams more susceptible to channelization during the wet period that followed their removal especially in those areas where 

vegetation had been altered by Native Peoples’ livestock or agriculture. Beaver dam 

failures likely combined with human check dam failures, their similar appearance and 

function triggering similar effects. The lack of references to wetlands in the area around 

the Zuni Pueblo by the early military expeditions suggests that the long-term presence 

of Native People may have already modified conditions such that wetlands had been lost 

in response to earlier activities.

Reexamining Changes on the Gila River in Light of Beaver Trapping

This analysis of changes on the Gila River builds on an earlier study by Burkham 

(1972) in which he assessed the likely causes of the channel widening between 1905 

and 1917 on the Gila where it flowed through the Safford Valley of Arizona. This river 
120 USDA Forest Service RMRS-GTR-377. 2018

basin, in southwestern New Mexico and southeastern Arizona, is a watershed scale example of how interpretations of causal mechanisms of change can vary once placed in a 

historical context that includes beaver trapping and more climate information.

The drainage area contributing to the Gila where it flows through the Safford 

Valley is about 7,900 miles2. In 1875, the average channel width was less than 150 ft. 

Between 1905 and 1917 the average width increased to about 2,000 ft during several 

large magnitude winter floods (Burkham 1972). The bulk of the widening occurred in 

1905 and 1906 and the source of the large floods was the mountainous headwaters of the 

Gila River. Examination of stream discharge and precipitation data available after 1910 

shows a strong correlation between large magnitude floods in the Gila River and rainon-snow events or high-intensity, long-duration storm events (Burkham 1970). Based on 

this correlation, Burkham (1972) concluded that the large flood magnitudes in 1905 and 

1906 were the result of precipitation events in the headwaters. His conclusions relied on 

the tree-ring data of Stockton and Fritts (1968) for Arizona, which identified 1907−1926 

as the wettest 20-year period since 1650.

Burkham (1972) also considered whether livestock grazing, rather than precipitation, could have been responsible for the large magnitude floods that occurred between 

1905 and 1917. He concluded that livestock grazing was not a factor based on his understanding that livestock appeared to be largely restricted to the lower watershed while 

the floodwaters had their source in the upper watershed. However, livestock grazing had 

occurred in the upper watershed. Swift (1926) noted livestock grazing on Bonita Creek, 

a tributary to the Gila River upstream of the Safford Valley, as early as 1884. Winn 

(1926) noted livestock grazing on the West Fork of the Gila River in the 1880s. A 1993 

environmental impact statement for a cattle allotment on the East Fork of the Gila River 

stated that the area was severely overgrazed by 1908 (USDA Forest Service 1993). The 

reference to overgrazing indicates that livestock were using the East Fork drainage prior 

to 1908. There is no reason to assume that the West Fork of the Gila and Bonita Creek 

would have been grazed in the 1880s while the East Fork of the Gila, San Francisco 

River, and other tributaries would have been ignored by early livestock owners.

In this arid landscape, the impact of livestock in the tributaries would have been 

high around the streams. Hendrickson and Minckley (1984) concluded that grazing 

in the 1880s would have been concentrated within a 3-mile radius of streams because 

livestock rarely travel more than 3 miles from a water source and only lightly graze 

areas greater than 1.8 miles from water. The result would be concentrated use along the 

tributary stream corridors accelerating the rate at which riparian vegetation and bank 

resistance to erosion decreased. Livestock grazing in the upper watershed would have 

also decreased upland vegetation, decreasing surface roughness. The reduced roughness 

and resistance to erosion would have resulted in gully development in the uplands, thus 

facilitating the increase in storm runoff reaching the stream channels. The changes that 

occurred in Mountain Meadows, Utah, in 1884 is an example of this gully development 

and captures the speed at which it could occur (Cottam and Stewart 1940, table 3).

Increases in runoff leading to elevated streamflows would have enhanced and accelerated changes triggered by the earlier beaver trapping. The Verde, Salt, San Pedro, 

San Francisco Rivers (tributaries to the Gila), as well as the Gila itself are all mentioned 

by name as being trapped between 1826 and 1834 (Pattie 1831; Weber 1971). While not 

specifically mentioned, the smaller tributaries would have also been trapped as trappers 
USDA Forest Service RMRS-GTR-377. 2018 121

systematically moved through the area. Thus, EA livestock grazing in the Gila River 

basin becomes the second, not first, large-scale regional disturbance in the area.

While Burkham (1972) focused on the 20-year period from 1907 to 1926 identified in Stockton and Fritts (1968), they also examined 10-year intervals. In this interval 

set, 1831 to 1840 showed up as having a relatively high probability of above normal 

precipitation. This time interval is close to the 1835−1849 interval identified by later 

researchers (D’Arrigo and Jacoby 1991; Meko 1990; Meko et al. 1991), both intervals 

post-dating intensive trapping. Thus, effects of beaver trapping (1826−1834), followed 

by a wet period (1831−1840 or 1835−1849), and then by intensive grazing in the upper 

watershed (early 1880s) would have interacted synergistically to accelerate drainage 

network development in the headwaters. The result would be more rapid transfer of water from the upper to the lower watershed, increasing flood frequencies, and amplifying 

flood magnitudes for precipitation events of all sizes.

The above scenario of a watershed that had recently undergone changes of such a 

nature that it was now responding to precipitation events differently is supported by the 

oral history of the Gila River Pima Indians and historical newspaper accounts that predate the 1910 installation of stream gages in the watershed. Floods occurred in 1833, 

1869, 1880, 1884, 1889, 1891, 1895, and 1896 (Dobyns 1981; Burkham 1970). The 

floods of 1833, 1895, and 1896 are of particular interest. The 1833 flood was the first 

major downstream flood recorded in Gila River Pima Indian oral history (Dobyns 1981). 

Since trapping had been ongoing in the upper watershed since 1826 and mostly done by 

1831 (Pattie 1831), Dobyns (1981) suggests that this flood may have been the result of 

the abrupt collapse of beaver dams destroyed in the preceding decade. This scenario is 

reasonable given that large discharges have been documented occurring in response to 

abrupt dam failures (Butler 1989; Hillman 1998) and the tree-ring data that recorded a 

period of above-average precipitation around that time (1830−1841 or 1835−1849). The 

floods of 1895 and 1896 were unusual because they occurred during the Southwest’s 

third most severe drought in the last 1,000 years (D’Arrigo and Jacoby 1990). These 

floods are a strong indication that the drainages in the upper watershed had changed.

The timing of the GLO surveys and early expeditions and their spatial geographies influenced the direction that future researchers went as they sought to explain 

arroyo formation and large magnitude flood events post-EA settlement. When the 

surveys and expeditions arrived in the Southwest in the 1850s, beavers had somewhat 

recovered from near extinction. There would have been little awareness of how much 

their numbers have been reduced or of the precipitation events that followed. Beavers 

were observed on the San Carlos River in 1846 (Leopold 1951), on the Little Colorado 

River in 1884 (Colton 1937), and on the San Pedro River in 1858 and as late as 1882 

(Hastings and Turner 1965).

The 1858 observation on the San Pedro River is of interest because it notes the following spacing of beaver dams:

The San pedro river as they Call it—is a stream one foot deep six ft wide and 

runs a mile and half an hour and in ten minutes fishing we Could Catch as 

many fish as we Could use and about every 5 miles is a beaver dam this is a 

great Country for them (Hastings and Turner 1965).
122 USDA Forest Service RMRS-GTR-377. 2018

Comparing the 1858 dam spacing to the numbers of beavers taken by trappers in 

1826 from the San Pedro and to dam spacing found today in basins with relatively unexploited beaver populations makes clear the magnitude of change that had occurred. In 

1826, trappers took 200 beavers from the San Pedro River and found beavers so abundant 

that they named it the Beaver River (Pattie 1831). Examples of average dam frequencies 

include 17 dams/mile in two drainage basins in Quebec (Naiman et al. 1986), 4 dams/

mile in the Kabetogama Peninsula in Northern Minnesota (Naiman et al. 1988), and up to 

6.5 dams/mile in Bridge Creek in eastern Oregon (Demmer and Beschta 2008)—all much 

greater than the 5 mile per dam spacing noted on the San Pedro River in 1858.

Explaining the Absence of Arroyo Formation From 1750 to 1825 Despite 

Large Herds of Spanish and Mexican Livestock

The relative contributions of livestock grazing versus climate as the cause of widespread arroyo formation in the Southwest post-EA settlement has been in question for 

nearly a century. Part of the reason for the uncertainty is that large livestock herds existed 

twice in the Southwest (1750−1825 and 1870−1905) but widespread arroyo formation 

occurred only once, from 1870 to 1905 (Cooke and Reeves 1976; Denevan 1967). Both 

periods of large herds had below-normal seasonal precipitation. Possible explanations for 

this discrepancy in landscape response to livestock grazing have included: (1) a gradual, 

long period of change in climate that altered vegetation to the point that watersheds in the 

late 1800s were more sensitive to livestock grazing than in the late 1700s to early 1800s; 

(2) the combination of overstocking the range and severe summer drought that occurred in 

the late 1800s but not in the prior period; and (3) some combination of the two (Cooke and 

Reeves 1976; Denevan 1967). Another explanation involves the change in beaver numbers 

and distributions between the two periods of intensive grazing.

Beavers were abundant in the late 1700s and early 1800s but greatly reduced by the 

late 1800s due to EA trapping in the 1820s and 1830s. Therefore, Spanish and Mexican 

herds would have grazed on landscapes that pre-date EA beaver trapping while the EA 

livestock herds grazed on landscapes that post-date trapping. The presence of abundant 

beavers during the time of Spanish and Mexican livestock grazing would have minimized 

any impacts that reduced upland vegetation had on runoff and potential flood magnitudes. 

Beaver-created wetlands and ponds would have provided stability to the fluvial systems 

during periods of drought and heavy rainfall (Bailey 1936; Grasse and Putman 1950; 

Hendrickson and Minckley 1984; Hillman 1998; Hood and Bayley 2008; Johnston and 

Naiman 1990). Despite the heavy sheep and cattle grazing, the dams would have kept 

ACC low, causing the additional water to be distributed onto the valley floor where its 

erosive power was less. When dam failures did occur, they would have been repaired or 

only localized, preventing the development of permanent arroyos or a channelized drainage network. In addition, use of the upper watershed for livestock grazing was likely much 

more limited during the earlier period than after the EA heads arrived (Love 1916).

When the large EA livestock herds arrived in the 1870s and began moving into the 

headwaters (Love 1916), the buffering effect of beavers and intact beaver dams had been 

gone for 30 to 40 years. The period of above-average precipitation, post-trapping but prior 

to EA settlement and grazing, would have accelerated dam failures and channel development. While agriculture increased in the lower watershed and other activities such as road 
USDA Forest Service RMRS-GTR-377. 2018 123

building and logging contributed to changes on the land, it would be livestock grazing in 

the upper watershed that made the next large contribution to channel changes because it 

was the most widely dispersed and use would have been concentrated along the stream 

corridors. Thus the extensive arroyo development from 1870 to 1905 was the response of 

an increasingly destabilized system in which one regional EA disturbance (trapping) was 

overlain by another regional disturbance (grazing), along with other localized EA impacts. 

As a result, even small climate events were now capable of increasing the channel network 

(table 3) and flood magnitudes as demonstrated by the frequent pre-1900 floods despite 

periods of drought.

The Question of Stream Sensitivity to Climate Variability

Adding beaver trapping into the story with the addition of newer tree-ring data 

helps answer the question: Are Southwestern streams and riparian zones inherently 

sensitive to climatic variability? The answer to the question is “no,” as long as stabilizing influences, such as abundant beavers and well-vegetated landscapes, are maintained 

because climatic variability is the norm. Using tree-ring data from northern New Mexico, 

D’Arrigo and Jacoby (1991) identified five periods of substantial drought and five periods of above-average winter precipitation over the last 1,000 years. Therefore, neither the 

drought nor the above-average precipitation in the late 1800s and early 1900s is unique, 

though the abruptness of the shift from severe sustained drought to above-average 

precipitation in the early 1900s does not appear to have a match in the tree-ring data 

(D’Arrigo and Jacoby 1991). Regardless, abundant research supports the ability of beaver 

ponds and wetlands to mitigate the effects of both drought and abundant precipitation on 

streamflow and stream systems (Bailey 1936; Beedle 1991; Grasse and Putman 1950; 

Hey and Philippi 1995; Hillman 1998; Hood and Bayley 2008; Schaffer 1940). Thus, 

rather than instability and sensitivity to climatic variability being an inherent characteristic of streams in the Southwest and elsewhere, this responsiveness is the result of EA 

beaver trapping with impacts compounded as subsequent EA disturbances occurred.

From Past to Present: Placing Hydraulic Geometry Relationships in Their 

Historic Disturbance Context

The early GLO surveys and expeditions lacked information on numbers and distributions of beavers and the appearance of stream ecosystems pre-trapping. The speed of 

the changes, the timing, and the spatial geographies of the surveys and early expeditions 

versus trapping, and the continued presence and perceived abundance of beaver post-EA 

trapping, gave the impression that the significance of beaver trapping at the watershed and 

regional scale was minimal. It was into this post-trapping, changing landscape that the 

field of geomorphology emerged in the 1870s. However, it was not until the 1940s that 

fieldwork and quantification of field data took hold in the discipline (Morisawa 1985). By 

then most streams in the lower 48 States had undergone multiple adjustments in channel 

morphology and hydrology in response to various land uses and climatic events. Evidence 

of beavers as a defining influence on stream ecosystems had become lost by the late 1700s 

to early 1800s in the East as other land uses buried beaver-created wetlands beneath feet 

of sediment (Walter and Merritts 2008). It had become fragmented by the 1850s in the 
124 USDA Forest Service RMRS-GTR-377. 2018

Southwest and Intermountain West by the time the GLO surveys and expeditions arrived, 

and invisible everywhere by the 1940s.

Using the limited tools and information at their disposal, the early geomorphologists 

developed relationships that continue to define the discipline and influence stream restoration efforts. However, new information and the development of new tools require that the 

relationships and observations found in early publications—publications covering such 

topics as hydraulic geometry (Leopold and Maddock 1953), formation of river floodplains 

(Wolman and Leopold 1957), and channel patterns (Leopold and Wolman 1957)—be 

reevaluated as to their meaning and appropriateness as guiding principles in stream 

restoration. Leopold and Maddock’s (1953) hydraulic geometry relationships serve as an 

example of how the interpretation of data and observations change once new information 

is incorporated into the analysis.

Leopold and Maddock used data from 112 stream gages to develop their hydraulic geometry relationships. They chose data from a diversity of geographic locations 

(fig. 21a), physiographic and geologic types, and sizes because their intent was to examine 

channel morphology, stream velocity, suspended sediment loads, and discharge information for general trends. The author found gage installation dates for 104 of their 112 gages 

(http://water.usgs.gov). Ninety-four percent of these gages were installed after 1900 (fig. 

21b) and after two waves of regional, large magnitude EA disturbances (table 6). When 

their data are placed within this historic disturbance context, it becomes clear that the hydraulic geometry relationships they developed, and the processes and rates of change they 

observed, reflected highly altered fluvial systems. By the time the gages were installed, the 

drainages had become interconnected systems of channels, the streams had become entrenched, over-wide, and hydrologically disconnected from their valley floors, and beavers 

no longer played their once defining role.

The first regional disturbance that influenced all their stream gages was beaver trapping. The second regional disturbance varied depending on location. For the stream gages 

east of the Mississippi River, the next regional disturbance was the building of thousands 

of water-powered mill dams. These dams trapped the sediment eroding off hillslopes and 

valley bottoms due to logging and agriculture. As time went on, the mill dams failed and 

new channels developed (Walter and Merritts 2008). For the lower Midwest stream gages 

(Kansas, Nebraska, and Missouri), the next regional disturbance was intensive agriculture 

in which abundant wetlands and beaver ponds were drained (Hey and Philippi 1995). The 

magnitude of stream changes in this area was captured by Schumm and Lichty (1963) in 

their work on the Cimarron River in southern Kansas. The Cimarron River increased in 

average width from 50 to 1,200 ft between 1874 and 1939 over roughly 175 miles with the 

variability in magnitude reflecting differences in bank composition and cohesiveness once 

the stabilizing vegetation had been removed (fig. 22).

Finally, for the Northern Rockies/Rockies stream gages (Montana, North Dakota, 

South Dakota, and Wyoming), it is livestock grazing that is the next regional disturbance. 

Leopold and Miller (1954), in their work on alluvial valleys in Wyoming, discuss the 

amount of channel incision that had occurred in this area. They specifically reference some 

of the same streams used by Leopold and Maddock (1953) in their hydraulic geometry 

paper. Thus by the time their stream gages were installed, the drainages conveyed water 

and sediment very differently than they once had.
USDA Forest Service RMRS-GTR-377. 2018 125

Leopold and Maddock’s hydraulic geometry relationships have become one of the 

guiding principles of fluvial geomorphology. While their relationships have been refined 

and made more place-specific since 1953, the conceptual model of drainages being interconnected channels remains. The continued use of these and other relationships developed 

from highly altered stream systems means that restoration efforts end up reinforcing the 

damage and loss of system resiliency that began with the fur trade in the 1600s.

a)

b)

0

5

10

15

20

25

30

35

1600s - 1809

1810 - 18191820 - 18291830 - 18391840 - 18491850 - 18591860 - 18691870 - 18791880 - 18891890 - 18991900 - 19091910 - 19191920 - 19291930 - 19391940 - 19491950 - 19591960 - 19691970 - 1979

Frequency

Time interval (years)

N = 104 1900

Beaver trapping

Figure 21—Spatial and temporal distribution of the stream gages used by Leopold and Maddock (1953) 

and the generalized timing of beaver trapping in the lower 48 States (source: Fouty 2003). (a) Spatial 

distribution of the stream gages. Gages are represented by the RED dots (N = 104), GREEN = beaver 

trapping from 1600s to 1785, ORANGE = beaver trapping from 1810 to 1850, WHITE = no known beaver 

trapping. (b) Installation dates for stream gages (sources: Installation dates U.S. Geological Survey website 

http://water.usgs.gov. Eight of the gages listed in Leopold and Maddock 1953 were not listed on the 

website; beaver trapping dates Phillips 1961). 
126 USDA Forest Service RMRS-GTR-377. 2018

Table 6—Relative temporal relationships of Euro-American disturbances and their impact on watershed hydrology and 

geomorphology (source: Fouty 2003.) 

Beaver establish a long-term presence

Upper watershed: Drainage network pattern a mix of ponds, wetlands and channels. Complex mosaic of riparian 

vegetation. Stream-valley floor hydrologic connection high and the valley floors frequently flooded. Stream 

ecosystem has low sensitivity to climatic variability, high resistance to disturbance and recovers rapidly after a 

disturbance.

Lower watershed: Flood magnitudes and the frequency of large magnitude floods decreases and flood durations 

increase.

First large-scale Euro-American disturbance: Beaver trapping

Widespread, temporally concentrated, and systematic removal of beaver from watersheds.

Drainage network transition I

Upper watershed: Dams fail, ponds drain and stream incises into fine sediments trapped behind the dams. Drainage 

network shifts to an increasingly channel-dominated network. Stream-valley floor hydrologic connection decreases 

as channels incise and widen. Wetland and riparian vegetation patterns begin to change in location and abundance 

in response to localized channelization, dropping water table, decreased valley floor flooding and beaver forage, 

and exposure of pond sediments. Reaches with intact wetlands continue to have low sensitivity to climatic 

variability but other areas have increased sensitivity due to channelization. Decreasing resistance to climatic 

variability and disturbance. Increased channelization in the upper watershed results in more rapid transfer of water 

from the upper to lower watershed. Drainage network a mix of channelized and nonchannelized reaches.

Lower watershed: Channel morphology may remain unchanged as valley floor and stream bank vegetation still 

abundant and dams were located on the floodplains and backwater areas. However, complexity of vegetation 

communities changing as a result of beaver removal. Possible increases in flood peaks and decreases in flood 

durations due to greater channelization in upper watershed and periodic abrupt dam failures.

Second wave of large-scale Euro-American disturbances (i.e. grazing, agriculture, logging, milldams, ditches)

Upper and lower watershed: Vegetation removed from uplands, valley floor and stream banks. Wetlands drained 

deliberately or incise due to land use activities. Creation of points of flow convergence (roads, canals). Result is 

large increases in runoff and decreases the resistance of uplands, valley floors, and stream banks to erosion. 

Upper watershed: Channelization expands and discontinuous channels begin to coalesce.

Drainage network transition II

Upper and lower watershed: Increases in available channel capacity as channels incise and widen. Water routed from 

upper to lower watershed during a storm event more rapidly. Streams and valley floors hydrologically disconnecting. 

Frequency of valley floor flooding in upper watershed decreases while the magnitude and frequency of flooding 

in the lower watershed increases. Stream ecosystem sensitivity to climatic variability increases, resistance to 

disturbance decreases and recovery rates after a disturbance slower

Current conditions (stream gages install during this period)

Upper and lower watershed: Channel-dominated. Streams and valley floors hydrologically disconnected. Reduced 

complexity, abundance and extent of the riparian zone. Loss of wetlands. Stream ecosystem sensitivity to climatic 

variability high, resistance to disturbance low and recovery after disturbance low.

Lower watershed: Increased frequency of higher magnitude floods.
USDA Forest Service RMRS-GTR-377. 2018 127

From Present to Future

Climate change is bringing increased variability. The impacts on wild and human communities are already large and will only increase. As water is essential to all 

communities, we need to shift our landscapes from “water-sheds” to “water-stores” 

by allowing the missing parts to return and take up their places and functions on the 

landscape. Public lands are the ideal place to start beaver-driven stream system restoration. Public lands provide large contiguous areas where beavers can rapidly create vast 

ecologically complex water storage zones and diverse habitat with limited infrastructure 

conflicts. Large portions of these lands occur in the headwaters, making them uniquely 

situated to store water during times of abundance and then release it during times of 

drought. Because many of the streams are first- through fourth-order streams, they are 

the appropriate size for beavers and their dam complexes to rapidly restore stream processes and form in these areas.

The challenges inherent in recovering at least some of the stream-valley floor hydrologic connections and the water storage capability of stream systems without beavers 

become clear when examining the amount of channel incision and/or widening that 

has occurred (table 3) and volume of stream sediment eroded post-EA trapping. Bryan 

(1928a) estimated that the Rio Puerco in New Mexico, with a drainage area of roughly 

6,220 mi2 (USGS stream gage), had lost more than 394,882 acre-ft of sediment over a 

42-year period as a result of channel incision and widening. On a smaller scale, an estimate of 7.2 acre-ft of sediment has been eroded from the 3.4 miles of headwater stream 

reaches studied by Fouty (2003). These drainages are located in southwest Montana and 

east-central Arizona and range from 692 to 18,775 acres.

The channelization of drainages across the North American continent resulted 

in the permanent removal of large volumes of sediment. Therefore, restoration of the 

stream-valley floor hydrologic connection, and the processes that result from that connection, requires abundant beaver dam complexes with their ponds in order to fill the 

0

500

1000

1500

2000

2500

3000

3500

0 20 40 60 80 100 120

Channel width (ft)

Observations along the Cimarron River

N = 120

Figure 22—Changes in 

channel widths and 

variability along the 

Cimarron River in 

southwestern Kansas 

between 1874 and 

1939. Open circles 

= 1874 widths. Solid 

circles = 1939 widths 

(figure generated using 

data found in Schumm 

and Lichty 1963).
128 USDA Forest Service RMRS-GTR-377. 2018

void left by the eroded sediment. Where the amount of erosion prevents recovering the 

original stream-valley floor hydrologic connection, beaver dam complexes are still required to restore stream processes. However, in these cases these new beaver-dominated 

systems will be inset into the larger channels as is the case at Bridge Creek in eastern 

Oregon (Demmer and Beschta 2008; Pollock et al. 2007) and streams near Elko, Nevada 

(Swanson et al. 2015).

Two key factors prevent beavers from expanding in numbers and distribution. 

The first is recreational and commercial beaver trapping (Muller-Schwarze and Sun 

2003), which removes not only existing beavers but their future progeny. An example of 

the cumulative effect of trapping on numbers is found in the data collected by Oregon 

Department of Fish and Wildlife and USDA Wildlife Services for the State of Oregon. 

The data listed 54,034 beavers reported killed between 2000 and 2015. Of this number, 

83 percent (44,784) of the beavers killed were due to hunting and trapping with the vast 

majority from trapping (about 97 percent).

The second factor is insufficient food and building material due to past and current 

land uses, with browse pressure on riparian woody vegetation by livestock and wild 

ungulates being a key contributor on public lands. When livestock are the dominant 

browser, as was the case for streams near Elko, Nevada, changes in livestock management resulted in rapid improvements in the quality of riparian habitat (Swanson et 

al. 2015). The expansion of riparian vegetation and the absence of trapping allowed 

beavers in this area to expand their range such that during an extreme drought in 2012 

the rivers with beavers still had water (Fouty, personal observation, July 15, 2012). 

Where wild ungulates are the browsers, work by Beschta and Ripple (2009, 2010) in 

Yellowstone National Park has shown the role that wolves play in decreasing elk and 

deer use of riparian areas leading to increased willows, aspen and cottonwoods. Here 

too beavers have expanded their range in response to improved habitat and no trapping 

(Smith and Tyers 2008). These studies show that sufficient food and building materials 

and the absence of trapping are required for beaver populations and their water storage 

benefits to expand. On public lands where both wild and domestic ungulates graze, 

changes in livestock management and expansion of wolf populations will be needed to 

reduce the browse pressure on key beaver food and building materials, along with the 

elimination of commercial and recreational trapping.

Conclusions

Separating out cause-and-effect relationships in fluvial systems is challenging because changes to their form and function are the result of many factors interacting over 

time and space. This chapter explored some of those factors in its examination of how 

EA beaver trapping altered the appearance and hydrologic behavior of stream systems 

and why the influence of beavers and beaver trapping were missed in the discipline of 

fluvial geomorphology until recently. It also examined how information gaps led to the 

development of relationships of process and form based on observations and measurements of channelized drainages and altered uplands that created conditions whereby 

water was rapidly shed from the landscape rather than stored and released slowly.

Given the magnitude of the historic changes and their hydrologic consequences, 

the scale of restoration and the rate at which it must occur is enormous if the impact of 
USDA Forest Service RMRS-GTR-377. 2018 129

climate change on water availability, and the systems that depend on water, are to be 

minimized. Partnering with beavers to restore the water-holding capability of our stream 

corridors would rapidly dampen fluctuations in the abundance and scarcity of water and 

leave wild and human communities less vulnerable. Efforts will require broad public 

support and an integrated approach by State and Federal agencies given their respective 

areas of influence and impact. Scientists are in a position to help inform the discussions 

by sharing what we have learned about how past and current land uses affect the ability of the landscape to naturally store water for future use; however, our effectiveness 

will first require that we change the lens we have been looking through. Because the 

discipline of fluvial geomorphology has internalized and codified degraded systems as 

normal, our stream restoration efforts fall short. By placing these fluvial geomorphic 

relationships within their historic disturbance context, one that includes EA beaver 

trapping, new strategies, approaches, and partnerships emerge that are essential for 

restoration to successfully occur. This new lens reveals the essential role beavers play in 

this recovery process.

Acknowledgments

My thanks to R. Roy Johnson and Richard W. Reeves for an early review of 

this paper and to Ken Kingsley for his suggestions and review of figures and tables. 

Thanks also to Katelyn P. Driscoll, R.H. Hamre, Dale A. Jones, Harley Shaw, and Patty 

Woodruff for their later reviews of this manuscript that improved it. Finally, a special 

thanks to Oz, Mariah, and Lily for their field assistance and to R. Roy Johnson (again) 

for helping to shepherd this manuscript through.

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Chapter 8. Arizona as a Watershed—Then and Now: 

Case Studies of Changed Management of Rivers and 

Habitat in the Lower Colorado River System

William E. Werner

Introduction

Prior to human development in the West, rivers flowed freely. Flows in the 

Colorado River varied greatly with season, with snowmelt runoff from the Rocky 

Mountains resulting in annual high flows (Topping et al. 2003). The large sediment 

loads historically found in the Colorado River, estimated to average 160,000,000 tons 

passing Yuma annually (LaRue 1916), have since caused Laguna Dam (fig. 23) near 

Yuma, Arizona, completed in 1905, to silt in almost immediately. Even in a managed 

system, unpredictable unmeasured tributary inflow into the Colorado River below 

Hoover Dam accounts for approximately 96,000 acre-ft of water annually, with about 

62 percent of that flow from Arizona tributaries, including Sacramento Wash, the Bill 

Williams River area, Bouse Wash, Tyson Wash, the Dome Rock-Trigo-Chocolate 

Mountains area, and the Gila River area (Owen-Joyce 1987). Grinnell (1914) provided 

an account of mammals and birds of the Lower Colorado River Valley, which provides 

a benchmark from early in the development of the river. Riparian habitat became established following seed deposition on surfaces where sufficient soil moisture maintained 

the vegetation to maturity. As we see today, large numbers of seeds germinate but few 

plants survive to maturity to provide seed to repeat the cycle.

Human development of the American West included construction of large dams to 

retain water from high flow periods for later use and clearing of floodplains for agriculture and urban development, impacting riparian habitat (table 7). Dam construction not 

only changed the pattern of water flow in rivers, but the pattern of sediment movement 

as well. Fine textured sediment, such as silt, often drops out of water flow in deltaic 

deposits at the upper end of reservoirs so that the fine sediment is no longer deposited 

along rivers to form surfaces where seeds collect and vegetation becomes established as 

described by Brady et al. (1985).

The effect is that, while the processes that perpetuate riparian vegetation persist, 

the pattern and location on the landscape have changed in many areas. Vegetation on 

sediment deposits forming deltas in reservoir pool space is often temporary, depending on water and soil moisture conditions suitable for maintaining the vegetation. 

Conditions change through time, being suitable—following a high flow—with sufficient 

groundwater in streamside alluvium that is supported by baseflow in the stream in the 

upper portions of the delta, or by stable reservoir water levels saturating alluvium in the 

lower portions of the delta. However, more extreme conditions result from either (1) 

complete inundation of vegetation, if a reservoir fills to capacity, or (2) contrastingly, 

drying conditions resulting from inadequate water supply to the vegetation if a reservoir 

is drawn down, resulting in groundwater levels dropping.
USDA Forest Service RMRS-GTR-377. 2018 135

Figure 23—Select Arizona rivers and conservation areas (map by W.E. Werner).
136 USDA Forest Service RMRS-GTR-377. 2018

Table 7—Arizona streams and riparian status.

Stream reach

Dams, irrigation project diversions, 

and flood control Current status of riparian ecosystems

Summary of 

Arizona streams and 

watersheds

Numerous “water projects” throughout 

Arizona (see below).

Much of riparian habitat at time of Euro-American 

settlement lost to agriculture, urbanization, and 

water projects (Dobyns 1981). Cottonwood-willow 

(Populus-Salix) and mesquite (Prosopis spp.) replaced 

by tamarisk (Tamarix spp.) along much of perennial 

reaches. 

Colorado River below 

Glen Canyon Dam 

and tributaries

None, Grand Canyon National Park most 

of river length to Lake Mead National 

Recreation Area.

Prior to Glen Canyon Dam, annual high flows from 

snowmelt in Rocky Mountains scoured vegetation from 

sides of the river. After Dam, tamarisk major woody 

vegetation, with scattered Goodding’s willow (Salix 

gooddingii), coyote willow (S. exigua), seep willow 

(Baccharis spp.). Willow flycatcher habitat occurs in 

lower Grand Canyon/Lake Mead Delta area depending 

on hydrologic conditions and pool elevation at Lake 

Mead.

Bill Williams River Alamo Dam, completed in 1969 to 

reduce flood damage along the lower 

Colorado River.

Large tamarisk stands; dam releases planned to benefit 

native trees, e.g., cottonwood-willows. Beaver dams 

established. Large stands of cottonwood and willow 

habitat following improved water management. Large 

numbers of western yellow-billed cuckoo (Coccyzus 

americanus occidentalis) in new habitat.

Salt River in Salt River 

Valley (Phoenix, AZ)

Corps of Engineers Rio Salado 

Environmental Restoration Project 

for flood control and environmental 

enhancement under authority to address 

effects of previous federal projects.

Original stands of cottonwood-willow gallery forests 

killed when Roosevelt Dam constructed upstream. 

Tamarisk minimal; cottonwood, Goodding’s willow, 

and mesquite. Beavers established; overwintering 

monarch butterflies (Danaus plexippus) in willows. 

Gila River below 

Painted Rock Dam

Storm flows down the Gila River from 

upstream dams contained in Painted 

Rock Reservoir, impounded by Corps of 

Engineers. Wellton-Mohawk Irrigation 

and Drainage District (WMIDD) located 

downstream from Painted Rock Dam. 

Large tamarisk stands, cattail (Typha) marshes; 

following high flows in 1993, seedling cottonwood 

and willow trees were observed in the WMIDD area 

and downstream. Lower reaches of the Gila River 

have supported habitat used by migrant southwestern 

willow flycatcher (McLeod 2008). 

Colorado River below 

Hoover Dam

Davis Dam, Parker Dam, Headgate 

Rock Dam, Palo Verde Weir, Imperial 

Dam, Morelos Dam; numerous irrigation 

districts; Havasu, Bill Williams River, 

Cibola, and Imperial National Wildlife 

Refuges; diversions by Metropolitan 

Water District of Southern California and 

Central Arizona Project.

National Wildlife Refuges created to provide wildlife 

conservation at the time of dam construction but 

were not enough to address effects of river operation 

to habitat. Cottonwood and willow habitat largely 

lost along the Colorado River, replaced by salt cedar. 

Lower Colorado River Multi-Species Conservation 

Program implemented to conserve threatened, 

endangered, and sensitive species and to provide 

ongoing Endangered Species Act compliance for 

operation and maintenance of the river.

If the conditions that support riparian vegetation are temporary at many locations 

in many watersheds without man’s intervention, better understanding of those conditions can help us better manage systems to maintain riparian vegetation in managed 

watersheds. There are opportunities to manage, to a greater or lesser degree, flow of water, sediment, stream morphology, and suitable surfaces for establishment of vegetation 

in flow-regulated systems. Examples include the Colorado River below Glen Canyon 
USDA Forest Service RMRS-GTR-377. 2018 137

Dam, the Bill Williams River, the Salt River in the Salt River Valley, the Gila River 

below Painted Rock Dam, and the Colorado River below Hoover Dam (fig. 23, table 7). 

Success has been variable. One of the challenges managers face is developing appropriate expectations and objectives in systems with variables beyond their control. It may be 

possible to maintain, as an example, a patch of cottonwood trees that can provide a seed 

source downstream when conditions are right to establish cottonwood trees after a large 

flow event. Whether new seedlings persist is constrained by many variables, but without 

the seeds from seed bearing trees there is little or no opportunity. The species composition of riparian vegetation on a managed stream below a dam may be quite different 

from unmanaged streams above the dam (Stromberg et al. 2012).

Colorado River Below Glen Canyon Dam and the Role of Tributaries

Much has been written about riparian vegetation and management of the 

Colorado River downstream from Glen Canyon Dam (fig. 23; e.g., Carothers and 

Dolan 1982 and Sankey et al. 2015). Glen Canyon Dam, forming Lake Powell, not 

only stops and stores snowmelt water from the upper Colorado River watershed, but 

also traps sediment that once flowed downstream. Downstream from Glen Canyon 

Dam, the Paria River discharges water and sediment into the Colorado at Lees Ferry, 

with a mean sediment contribution of approximately 3 million tons/year (Andrews 

1991). While surface water flow in the Paria River watershed is not managed to a large 

degree, the watershed is desert and significant flow of water and sediment depends on 

episodic storm events. Farther downstream, the Little Colorado River discharges water 

and sediment from a larger watershed with a mean sediment contribution of approximately 9 million tons/year (Andrews 1991). Many streams in the Little Colorado River 

watershed have been dammed for consumptive use or recreation, changing the frequency of significant flow events. The Puerco River, a tributary to the Little Colorado 

River, is largely unregulated and can produce a large volume of sediment under intense 

runoff conditions (Wirt et al. 1991).

Sediment inflow into the Colorado River from the Paria and Little Colorado Rivers 

is monitored by the Grand Canyon Monitoring and Research Center as part of the Glen 

Canyon Dam Adaptive Management Program pursuant to the Grand Canyon Protection 

Act. When sediment accumulation in the Colorado River has reached an appropriate 

level, recommendations for higher water releases from Glen Canyon Dam are made to 

the Secretary of the Interior with goals generally to benefit resources along the river by 

re-depositing the sediment at higher elevations to benefit riparian vegetation and to create bars to form backwater aquatic habitats (Salazar 2012).

Research into the effects of efforts to manage water and sediment below Glen 

Canyon Dam conducted by the Grand Canyon Monitoring and Research Center and 

others in association with the Glen Canyon Dam Adaptive Management Workgroup is 

ongoing. As described by Johnson and Carothers (1987) and as a former participant in 

the Workgroup, one of the challenges is identifying management objectives appropriate for the changed system. Water resource development in the watershed above Glen 

Canyon Dam has resulted in large changes to the pattern and magnitude of movement of 

both water and sediment. Yet there are opportunities to do the best we can with circumstances as they are today.
138 USDA Forest Service RMRS-GTR-377. 2018

Sediment from the Colorado River below the Grand Canyon has formed deltaic 

deposits within the reservoir pool area of Lake Mead, providing seed beds suitable for 

native riparian tree species (McLeod et al. 2008). Depending on reservoir operating 

conditions, namely the amount of water entering the system versus releases for downstream use, native trees such as Goodding’s willow (Salix gooddingii) have provided 

habitat for the southwestern willow flycatcher (Empidonax traillii extimus; McLeod 

et al. 2008) as long as moisture remained to support the trees in a stable or falling 

reservoir condition and water did not completely inundate them in a rising reservoir 

condition (1996 condition). Flooded vegetation can apparently provide protective cover 

for young native fish, including razorback sucker (Xyrauchen texanus) increasing survival and recruitment into the adult population (Albrecht et al. 2008). Thus, although 

within the pool of a reservoir, the natural processes facilitating recruitment of native 

trees and birds, as well as native fish, can occur in a managed system, although perhaps 

more accidental than planned.

Alamo Dam and the Bill Williams River

The Bill Williams River is a major tributary of the Colorado River (fig. 23) beginning at the confluence of the Big Sandy and Santa Maria Rivers in west central Arizona. 

Alamo Dam, completed in 1969, was constructed to manage flood flows from the Bill 

Williams River that would otherwise cause flooding along the Lower Colorado River 

in the vicinity of Parker and Yuma, Arizona. Hunter et al. (1987) describe changes in 

habitat below Alamo Dam following flooding from high-water releases. Beginning in 

1989 the Bill Williams River Corridor Steering Committee, a multi-agency effort, began 

planning for improved management of flows from Alamo Dam. We sought to benefit 

native tree species and avoid benefitting nonnative saltcedar (Tamarix spp.) both downstream of the dam and in the reservoir pool, through recommendations to the Corps of 

Engineers (Werner 2010). Saltcedar is well established in the Bill Williams River watershed, and shoreline deposition of seeds at high reservoir conditions in the late 1970s 

and early 1980s resulted in establishment of plants around the shoreline of Alamo Lake 

and where the lake inundated the lower ends of the Santa Maria and Big Sandy Rivers. 

In analyzing an optimal operating range for Alamo Lake, working with all existing flow 

data, we sought to minimize the probability of pushing the “bathtub ring” of saltcedar 

further up the rivers. In addition we attempted to mimic natural timing of downstream 

high releases from Alamo Dam to benefit native trees and tried to avoid releases that 

would favor saltcedar. We also recommended base flow releases that would vary seasonally to sustain trees. This planning was informed by research on the Hassayampa 

River by Stromberg et al. (1991) who found that recruitment of cottonwood and willow 

trees occurred approximately every 7 to 10 years on that unmanaged system. The Bill 

Williams River Corridor Steering Committee was concerned that dam operation and 

flow recommendations not result in a rigid management regime, as had occurred with 

previous recommendations responding to a planned drawdown that resulted in management of the reservoir with a fixed target elevation, and included a recommendation for 

development of a monitoring program and an outline of information and data needs to 

facilitate improvement of recommendations, or adaptive management, in the future. 

During the period of planning by the Bill Williams River Corridor Steering Committee, 
USDA Forest Service RMRS-GTR-377. 2018 139

heavy winter rains occurred in 1993 in the Bill Williams River watershed, resulting in 

a need for high releases from Alamo Dam. The Corps of Engineers was able to incorporate draft recommendations, based on the existing Water Control Manual, into the 

schedule for high releases in an attempt to avoid detrimental effects documented previously by Hunter et al. (1987). The Corps of Engineers funded aerial photography during 

the high releases to document the inundated area along the river.

Operating a reservoir to benefit downstream riparian habitat is like using a debit 

card drawn from a checking account. There is a finite amount of water in the “account” 

from which to make baseflow releases. If there is inflow, “income,” withdrawals can be 

sustained to the extent that the volume released doesn’t exceed the volume available. 

If withdrawals are larger than income, the balance in volume goes down. If inflows 

go down due to reduced precipitation in the watershed, releases must be reduced to 

maintain a “balance” in the system. Reduced baseflow releases may result in a reduction of the total acreage of riparian habitat maintained downstream from the dam, with 

some reaches more susceptible to change than others based on geology and landform 

(Wolcott et al. 1956; House et al. 1999). Maintaining releases to maintain the acreage of 

riparian habitat downstream will reduce the balance available to make future releases, 

to the eventual point that releases cannot be made until inflow occurs. The quandary is 

how much releases should be reduced to be able to make any releases in the future or 

how much downstream habitat acreage we are reluctantly willing to give up, assuming 

that future inflows will occur in the nick of time. There are no right or wrong answers 

to these questions. Decisions can be informed, however, by better understanding the 

variables that control the system and consequences of decisions. Ultimately, we don’t 

control nature, and precipitation may not be as we desire and the acreage of riparian 

habitat will adjust downward.

Similar to Lake Mead, deltaic sediment deposits at the upper end of Alamo Lake 

have provided suitable surfaces and conditions for establishment of native riparian 

vegetation. Goodding’s willow trees in that area have provided nesting habitat for southwestern willow flycatcher (McLeod and Pelligrini 2013). While conditions may remain 

to support this vegetation, suitable nesting habitat conditions may change through time 

as vegetation matures and stands thin. The stands of cottonwood and willow trees that 

became established (fig. 24) following implementation of management recommendations for Alamo Dam downstream from Alamo Dam on the Bill Williams River have 

supported large numbers of western yellow-billed cuckoo (Coccyzus americanus occidentalis) with 139 detections in 2007 (Johnson et al. 2008).

Subsequent to the original recommendations of the Bill Williams River Corridor 

Steering Committee, changes to the authorizing legislation for Alamo Dam, and formal 

revision of the dam Water Control Manual, there has been additional analysis of flow 

recommendations and research and monitoring of flow effects. Monitoring and evaluation of flow prescriptions is reported by Konrad (2010). Flow effects to vegetation have 

been studied by Shafroth et al. (2006, 2010) and Wilcox and Shafroth (2013). Reynolds 

et al. (2014) report changes to vegetation on abandoned floodplain surfaces. Stromberg 

et al. (2012) describe vegetation on the lower Santa Maria River, above Alamo Dam, 

compared to vegetation below Alamo Dam on the Bill Williams River, finding the latter 

less diverse. Vegetation response to water table changes was the subject of research by 

Shafroth et al. (2000). Anderson et al. (2010, 2011) researched beaver dams and effects 
140 USDA Forest Service RMRS-GTR-377. 2018

associated with controlled flooding. Anderson et al. (2016) researched wood decay on 

the Bill Williams River. Ellington and Anderson (2002) researched spatial correlation 

of Apache cicada (Diceroprocta apache), and Tamarix on the Bill Williams River. 

Nelson and Anderson (1999) researched butterfly assemblages in natural and restored 

habitats, including on the Bill Williams River. Wilson and Owen-Joyce (2002) describe 

hydrologic conditions on the lower Bill Williams River. Wiele et al. (2009, 2011) report 

on sediment discharge from the Bill Williams River and sediment concentration in 

downstream Lake Havasu associated with high flow releases designed to benefit riparian 

vegetation.

Salt River in Salt River Valley

Following flooding on the Salt River in 1993, the U.S. Army Corps of Engineers 

conducted studies at the request of local sponsors to determine if a Federal project was 

justified to address flood control, water conservation, recreation, and environmental 

restoration on the Salt River in the vicinity of Tempe and Phoenix (fig. 23). Review of 

environmental restoration was based on authority in the Water Resources Development 

Act of 1986 to address environmental effects of existing projects, in this case upstream 

dams. As a result of these studies, the Congress of the United States authorized the 

Corps of Engineers Rio Salado Environmental Restoration Project on the Salt River between Interstate 10 and 19th Avenue to address flood control, environmental restoration, 

and incidental recreation with the City of Phoenix as the non-Federal sponsor.

In designing the cross section for the Rio Salado Environmental Restoration 

Project, the Corps of Engineers created a “channel within a channel” with dimensions 

such that flows that would be expected every 7 to 10 years would overtop the low-flow 

channel, based on research on the Hassayampa River of the flood frequency maintaining 

Figure 24—Riparian vegetation on Bill Williams River near Narrows on Bill Williams River NWR (photo by W.E. Werner).
USDA Forest Service RMRS-GTR-377. 2018 141

riparian trees (Stromberg et al. 1991), but would be contained within dimensions of a 

flood channel designed to protect adjacent infrastructure. The watershed above the Rio 

Salado Environmental Restoration Project includes the Salt and Verde River watersheds. 

There are four large dams on the Salt River, and flood flows are much less likely to 

completely fill the Salt River reservoir system than to fill the two reservoirs on the Verde 

formed by two relatively smaller dams. Because of this, flood flows on the Salt River 

through Phoenix typically originate in the Verde River watershed.

Within the Rio Salado Environmental Restoration Project water to sustain riparian 

vegetation (fig. 25) is either pumped to higher areas or is drainage water entering the 

Salt River from the local storm drain system. While storm drain flow isn’t predictable 

it has been consistent and supports habitat in the low-flow channel. Apparently sump 

pumps pumping high groundwater from the basement levels of some high rise buildings in downtown Phoenix discharge into the storm drain system, providing some of 

the consistent flow. Efficacy of managing channel size to benefit riparian resources 

through time will depend on the frequency of flood flows and availability of water to 

sustain vegetation. An unanticipated high flow through the Rio Salado Environmental 

Restoration Project occurred in July 2010 when one of the bladders of the inflatable 

dam at the downstream end of Tempe Town Lake failed, releasing a flood down the 

Salt River. The low-flow channel through the Rio Salado reach appeared to perform as 

intended based on high-water marks and flood deposits.

Vegetation within the Rio Salado Environmental Restoration Project includes 

native Fremont cottonwood (Populus fremontii), Goodding’s willow, and mesquite 

(Prosopis spp). Overwintering of monarch butterflies (Danaus plexippus) has been 

Figure 25—Riparian vegetation on Salt River at Central Avenue, Phoenix Arizona (photo by W.E. Werner).
142 USDA Forest Service RMRS-GTR-377. 2018

documented in a dense stand of Goodding’s willow in the Rio Salado Environmental 

Restoration Project area (Morris et al. 2015) although suitability of conditions may depend on water management in areas higher in elevation than the low-flow channel (Gail 

Morris, Southwest Monarch Study, Chandler, AZ, personal communication, October 15, 

2016). Beavers have become established in the area with evidence of tree gnawing and 

dam construction.

Gila River Below Painted Rock Dam

Flow events in central Arizona that exceed the storage capacity of the Salt River 

Project system on the Salt and Verde Rivers—Lake Pleasant on the Agua Fria River, or 

San Carlos Reservoir on the Gila River—flow down the Gila River into Painted Rock 

Reservoir, impounded by the Corps of Engineers Painted Rock Dam. As indicated in 

the discussion of the Salt River, many of these events are associated with flows on the 

Verde River but can also result from storm flows from southeastern Arizona flowing 

into the Gila from the Santa Cruz or San Pedro Rivers. Downstream from Painted Rock 

Dam, the Wellton-Mohawk Irrigation and Drainage District (WMIDD) was developed 

as an element of the Gila Project by the Bureau of Reclamation to irrigate floodplain 

farmland. Irrigation water is diverted from the Colorado River at Imperial Dam and is 

delivered to the WMIDD area along the lower Gila River through a canal system and 

pump plants. Flood flows past Painted Rock Dam exceeded the capacity of the Gila 

River channel through the WMIDD in 1993.

Following high flows in 1993, seedling cottonwood and willow trees were observed in the WMIDD area and downstream. Trees that became established typically 

were on first terrace silty soils that would retain soil moisture more than river bottom 

sandy soils. Efforts to restore a channel after the high flows included plans to maintain 

cutoff meander bends and low-lying areas. In addition, the invert elevation of the channel was raised one foot to increase the probability that future flood flows would overtop 

the low-flow channel every 7 to 10 years, again based on research on the Hassayampa 

(Stromberg et al. 1991) to facilitate regeneration of native riparian trees, including cottonwood and willow. This is an example of attempting to shape the channel such that 

natural processes will maintain riparian habitat.

Groundwater or soil moisture to support riparian habitat along the Gila River 

downstream from Texas Hill, the upstream end of the WMIDD, is either high groundwater resulting from irrigation with imported Colorado River water or from lateral 

movement of irrigation water toward lower elevations in the river channel. The first 

circumstance can change with groundwater pumping patterns and the second can change 

with cropping patterns and irrigation of agricultural fields adjacent to riparian habitat. 

Another complicating variable is that groundwater in the WMIDD is often saline, which 

can vary with mixing with irrigation water from above or less saline water from the 

river channel during flow events.

Following flood flows in 1993 cottonwood and willow seedlings were observed 

downstream from mature trees that were apparently supported by water released at ends 

of the canal system or wasteways. Since a canal system operates on gravity, excess 

water that is not diverted as planned by a farmer must go somewhere and is sometimes 

dumped from the canal system into the river floodplain. Sometimes water is flushed into 
USDA Forest Service RMRS-GTR-377. 2018 143

the river to clean wind borne sand from the canal system. In addition, in some systems, 

leaky gates account for additional water into the river. This results in areas that support 

mature native trees that can provide seeds during high-flow events on the otherwise 

dry river. There were flood flows on the Gila downstream from Painted Rock Dam in 

1993-1995, and smaller high flows in 2005 and 2010. In many areas cottonwood trees 

that became established following the 1993-1995 high flows have died back. Survival of 

cottonwood and willow trees appears to have been better on silty soils of the Gila River 

floodplain than in coarser texture sands in the river channel itself. The silty floodplain 

soils are also desirable for agriculture as water is better retained to provide moisture to 

plants. Lower reaches of the Gila River have supported habitat used by migrant southwestern willow flycatcher (McLeod 2008).

Colorado River Below Hoover Dam

The Lower Colorado River Multi-Species Conservation Program (LCRMSCP) 

is designed to create habitat and support programs for the benefit of native fish and wildlife affected by operation and maintenance of the Lower Colorado River Dam system 

from the upper end of Lake Mead above Hoover Dam to the Southerly International 

Border with Mexico at San Luis, Arizona (Werner 2006). State and Federal water, power, and fish and wildlife agencies worked together to develop the program with a focus 

on ensuring long-term compliance with the Endangered Species Act and to avoid listing 

of additional species (a total of 26 covered species). The program includes creation and 

management of 8,132 acres of habitat, including 5,940 acres of cottonwood-willow 

habitat with characteristics suitable for southwestern willow flycatcher and western 

yellow-billed cuckoo, 1,320 acres of honey mesquite, 512 acres of marsh, and 360 acres 

of backwater, plus specific measures for certain species (LCR MSCP 2004).

Along the Colorado River below Hoover Dam efforts to maintain riparian habitat 

include channel shaping and water management at the Laguna Division Project between Imperial and Laguna Dams (LCR MSCP 2012) and planting on silty floodplain 

soils. Water supporting trees at planting sites is either: (1) surface flow irrigation from 

the Colorado River either by diversion at Palo Verde Diversion Dam for Palo Verde 

Ecological Reserve (fig. 23) in California or (2) pumping from the river at Cibola Valley 

Conservation Area (fig. 23) and Cibola National Wildlife Refuge in Arizona. Because 

of post-dam changes to the Colorado River, the water surface elevation is too low to 

overtop the banks to flood silty floodplain soils.

Opportunities to take advantage of natural flooding of low-lying areas have been 

investigated, but the acreage that can be accomplished through this means is limited, 

often because sediment in the Colorado River is trapped in reservoirs and is no longer 

available for building bars that provide sites for seed germination. In addition, the channel of the Colorado River has been simplified through human development of the area 

such that water does not spread across the floodplain.

Habitat creation targeting western yellow-billed cuckoo and southwestern willow 

flycatcher has included creation of large blocks of intermixed cottonwood, willow, 

and mesquite habitat. Hunter et al. (1987) reported that the yellow-billed cuckoo was 

declining and the willow flycatcher had disappeared from most of its historical range 

in Arizona. There is an extensive monitoring program associated with the LCRMSCP. 
144 USDA Forest Service RMRS-GTR-377. 2018

Western yellow-billed cuckoo began nesting in new stands of vegetation within a few 

years of establishment (Johnson et al. 2008) and there may be bird response and interaction between blocks of mainstem and tributary habitat, including recent habitat on the 

Bill Williams River discussed earlier, as well as response to decrease in irrigation once 

trees are established (Tracy and McNeil 2016), a pattern similar to patterns reported by 

Wallace et al. (2013).

Discussion

Riparian ecosystems are dynamic by their nature and riparian habitat is dynamic 

as stands of vegetation progress through natural succession and mature. Maintaining 

sufficient habitat on the landscape involves more than simply maintaining stands 

of vegetation. Maintaining conditions suitable to support an adequate prey base for 

birds, for example, is an element of maintaining habitat. Doing so may require more 

water than is needed simply for survival of trees. In some circumstances, our ability 

to control variables limits our ability to manage for riparian habitat. In other circumstances, we can set the stage, to the best of our ability, within constraints, such as on 

the Salt River in Phoenix where we can manage the landform and can create baseflows 

but have little control on high flows. In further circumstances, such as the operation 

of Alamo Dam, we can manage flow constraint by managing both high flows and 

baseflows, to a greater or lesser degree, within physical constraints of dam design and 

available water upstream in the system.

In the case of operating Alamo Dam to benefit riparian habitat on the Bill Williams 

River, we can develop contingency plans to exploit opportunities, if and when those 

opportunities arise. On the lower Gila River below Painted Rock Dam, we can set 

the stage by creating suitable landform but have no control on the frequency of high 

flows, which were more frequent during our planning efforts than since. On the Lower 

Colorado River, we can create functional habitat in floodplain settings where it once 

occurred naturally although our ability to manage river flows and sediment movement 

is very low. By doing the best we can in a variety of circumstances in managed systems, 

we can create a patchwork quilt of habitat.

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148 USDA Forest Service RMRS-GTR-377. 2018

Chapter 9. Evaluating Riparian Vegetation Change in 

Canyon-Bound Reaches of the Colorado River Using 

Spatially Extensive Matched Photo Sets

Michael L. Scott, Robert H. Webb, R. Roy Johnson, Raymond M. Turner, 

Jonathan M. Friedman, and Helen C. Fairley

Introduction

Much of what we know about the functional ecology of aquatic and riparian 

ecosystems comes from work on regulated rivers (Johnson et al. 2012). What little 

we know about unregulated conditions on many of our larger rivers is often inferred 

from recollections of individuals, personal diaries, notes, maps, and collections from 

early scientific surveys (Webb et al. 2007) and from repeat photography (Turner and 

Karpiscak 1980; Webb 1996). Later surveys often included photographs typically taken 

at specific and relatively few locations. Because rivers and their related riparian ecosystems vary in both longitudinal and lateral dimensions, such records provide relatively 

general and incomplete information about historical resource conditions, especially at 

large spatial scales and over long spans of time.

Pre-development conditions have been more extensively documented along the 

canyon-bound sections of the Colorado River than along any other river in the western 

United States (Powell 1961; Smith and Crampton 1987; Wheeler 1872). Unique to 

this reach of the Colorado River are the existence of photographs taken systematically 

by the Stanton expedition of 1889-1890 as part of a survey for the route of a railway 

through the canyons of the Colorado and matched originally in the 1990s (Webb 1996). 

This work includes a total of 445 photographs, which were taken between what is now 

Glen Canyon Dam and the upper end of Lake Mead, a distance of 485 river kilometers 

(rkms). The unparalleled spatial coverage of these photos provides a broad-scale and 

comprehensive visual record of pre-development conditions along this reach of the 

Colorado River. As such, this photographic record has and may continue to serve as a 

basis for monitoring change in the condition of desert and riparian resources in these 

remote canyons (Webb et al. 2011).

In this investigation, based on both original matches made in the 1990s and a second set taken between 2010 and 2012, repeat photography in Grand Canyon was used to 

detect and evaluate changes in riparian vegetation over time, relative to distinct changes 

in flow regime downstream of Glen Canyon Dam in Glen, Marble, and Grand Canyons 

(fig. 26). All Stanton original photographs are from the National Archives and are public 

domain. All the matches were done by USGS and are public domain. Our analyses 

are compatible with evaluations based on aerial imagery (Sankey et al. 2015) with the 

additional benefit of species-specific assessments. We evaluated change in riparian 

vegetation relative to earlier descriptions of change and predictions of future condition 

following completion of Glen Canyon Dam (Johnson 1991; Turner and Karpiscak 1980; 

Webb 1996). Finally, we examine on-the-ground response of riparian vegetation in 
USDA Forest Service RMRS-GTR-377. 2018 149

relation to recent work that explicitly links flow and climate variables to riparian vegetation change at a broader scale in the post-dam era (Sankey et al. 2015).

Hydrologic and Climatic Context

Pattern in the occurrence, composition, and productivity of riparian vegetation 

is structured by transverse (lateral) and longitudinal physical gradients related to the 

river. Transverse gradients, particularly in arid regions and along physically constrained 

channels, change sharply over relatively short distances, and riparian vegetation is 

Figure 26—Study area map, including 25 river kilometers of Glen Canyon between Glen Canyon Dam 

and Lees Ferry (cross). Marble Canyon extends from the Paria River to the Little Colorado River and 

Grand Canyon, from the Little Colorado River to approximately 50 river kilometers below Diamond 

Creek (cross). The solid circles are locations of the 160 original Stanton Expedition photos matched 

and analyzed for this study (figure by M.L. Scott).
150 USDA Forest Service RMRS-GTR-377. 2018

primarily arrayed along a hydrologic gradient (Friedman et al. 2006). Variation in riparian vegetation away from the channel correlates with several interrelated factors that 

include decreasing fluvial disturbance and inundation (Auble et al. 1994, 2005), change 

in fluvial landforms (Hupp and Osterkamp 1985), groundwater availability (Webb and 

Leake 2006), and increasing drought stress (Zimmerman 1969). Longitudinal gradients 

range from the scale of local hydrologic controls like debris fans to the watershed, 

where along most western rivers, factors such as temperature, discharge, and floodplain 

area tend to increase predictably in the downstream direction (Ward et al. 2002).

Streamflow and sediment transport in the Grand Canyon were completely altered 

by construction of Glen Canyon Dam. Paleoflood records near Lees Ferry, spanning the 

past 4,500 years, indicate that prior to the dam, the long-term discharge of a common 

flood (2-year recurrence interval) was 2,133 cubic meters per second (cms) or 75,326 

cubic feet per second (cfs) (O’Connor et al. 1994). More recent estimates give a revised value of 2,406 cms (85,000 cfs) for the 2-year return flood (Topping et al. 2003; 

fig. 27). The post-dam period includes two contrasting periods of stream discharge, 

primarily the result of changing reservoir operations. From 1963 to 1993, Glen Canyon 

Dam released water in response to power demands (hydropeaking) with daily highs as 

high as 878 cms (31,005 cfs) and lows averaging 140 cms (4,944 cfs). This period also 

featured three large, long-duration floods in 1983, 1984, and 1986 (fig. 28). From 1993 

to the present, hydropeaking releases were limited to a narrower range of high and low 

flows (US DOI 1996) and six, short duration, high-flow experiments occurred in 1996 

(Collier et al. 1997; Webb et al. 1999), 2004, 2008, 2012, 2013, and 2014 (Grams et al. 

2015; fig. 28).

Riparian Vegetation in the Grand Canyon

Understanding riparian vegetation in the canyons of the Colorado River begins 

with the dynamics and distribution of fine-grained sediment. The most conspicuous 

pre-dam riparian vegetation, termed the old high-water zone (Johnson 1991), consisted 

of discontinuous bands of woody plants associated with alluvial deposits emplaced by 

Figure 27—Annual peak 

discharge, in cubic meters 

per second, for the period of 

record at the Lees Ferry gage 

(U.S. Geological Survey gage 

09380000). Horizontal lines 

and associated values indicate 

the average 2-year return peak 

flow for the pre-dam and two 

post-dam periods. Discharge 

of the 1884 flood was 

estimated. Relatively wet or 

dry periods, based on regional 

precipitation, are indicated.
USDA Forest Service RMRS-GTR-377. 2018 151

relatively common pre-dam floods up to 2,800 cms. The vegetation in this zone consists 

of an ensemble of species whose composition shifts upstream to downstream. Netleaf 

hackberry (Celtis reticulata) is a common tree of this zone through Cataract, Glen, 

and Marble Canyons. The old high-water zone in Marble Canyon (fig. 26) includes 

Redbud (Cercis orbiculata) and Apache plume (Fallugia paradoxa), a shrub, present to 

about river kilometer (rkm) 65. Below rkm 65 and where the canyon widens, catclaw 

acacia (Acacia greggii) and honey mesquite (Prosopis glandulosa) dominate this zone. 

Throughout most of their range, neither of these warm desert species is restricted to 

riparian habitats, but in the Grand Canyon, mesquite is an obligate riparian species and 

is limited to alluvial terraces and talus slopes adjacent to the river (Warren et al. 1982). 

Honey mesquite is largely absent in the narrow inner gorges (~rkm 120-270) where 

alluvial terraces are rare and the old high-water zone is absent or discontinuous. Here, 

catclaw grows along the canyon walls and on talus slopes. Below rkm 270, mesquite 

reappears and becomes abundant on alluvial deposits (Anderson and Ruffner 1987).

The species in the old high-water zone are largely native and have been termed 

facultative riparian species (Johnson and Lowe 1985) and with the exception of honey 

mesquite, they do not strictly require riverine disturbance or moisture subsidies to 

establish and persist. However, species like catclaw reach their greatest densities and 

sizes along tributaries or in the main canyons, within influence of the river (Anderson 

and Ruffner 1987). In the case of honey mesquite in Grand Canyon, and other riparian 

settings, flooding likely serves as an important agent in seed dispersal and germination (Stromberg 1993).

Figure 28—Average daily discharge at Lees Ferry for the post-dam period. The effects of two different hydropower 

peaking release patterns on peak and base flows are evident, as are the series of large, long-duration floods in the 

mid-1980s. Three short duration, high flow experiments occurred in 1996, 2004, and 2008, followed by three high 

flow experiments in 2012, 2013, and 2014.
152 USDA Forest Service RMRS-GTR-377. 2018

To establish and persist, early successional, woody riparian plants require bare, 

moist, and typically fine-grained alluvium that remains relatively safe from subsequent fluvial disturbance (McBride and Strahan 1984; Scott et al. 1997). Pre-dam 

photographs (Turner and Karpiscak 1980; Webb 1996) and systematic plant collections (Clover and Jotter 1944) in the canyons of the Colorado River suggest that these 

conditions were infrequently met. Canyon width and the occurrence of debris fans 

provide the primary structural controls on fine sediment deposition in these canyons 

(Howard and Dolan 1981; Schmidt 1990). Most fine-grained alluvial surfaces below 

the high flood line occurred as eddy sand bars and discontinuous channel margin 

deposits associated with debris-fan channel constrictions. These surfaces, particularly 

in narrow canyon settings, were apparently reworked frequently enough by pre-dam 

flows that they were largely devoid of vegetation.

Species, including sandbar willow (Salix exigua), seep-willows (Baccharis 

emoryi and B. salicifolia), Fremont cottonwood (Populus fremontii), and arrow-weed 

(Pluchea sericea), are riparian obligates, requiring connection with the river for 

moisture throughout the growing season. These and other obligate riparian species 

were found at the mouths of or upstream on some of the perennial tributaries or in 

wider canyon reaches like Lees Ferry (Clover and Jotter 1944). In western Grand 

Canyon, where wider canyon reaches allow for reduced flood power and more extensive deposition of fine sediment (Schmidt and Grams 2011), riparian trees survived 

in some locations. A Fremont cottonwood tree can be seen below the high-water line 

at rkm 313.6 in a 1923 photo (Turner and Karpiscak 1980). Clover and Jotter (1944) 

note the occurrence of large tree willows (Salix gooddingii), which also can be seen in 

pre-dam photos from Granite Park (Turner and Karpiscak 1980) and other locations in 

western Grand Canyon as documented in written records (Webb 2005) and photos. As 

with species in the old high-water zone, some obligate riparian species show longitudinal trends in occurrence. For example, desert broom (Baccharis sarothroides) first 

appears around rkm 200 and becomes very common below National Canyon. In the 

late 1930s, the nonnative riparian tree tamarisk (Tamarix ramosissima) was reported 

to be relatively common from Green River, Utah, to the confluence of the Colorado 

River. It also was common in the approaches to Cataract Canyon, in wide reaches of 

Glen Canyon, and downstream from Grand Canyon. It was found above and at Lees 

Ferry but reported only below Vaseys Paradise and at the mouth of Saddle Canyon 

in Marble Canyon. In lower Grand Canyon, it was found above Lava Falls, near the 

mouth of Spring Canyon and at Separation Rapid (Clover and Jotter 1944; Webb 

2005).

The closing of Glen Canyon Dam in March 1963 fundamentally altered the 

downstream aquatic and riparian ecosystems in Glen, Marble, and Grand Canyons 

(Dolan et al. 1974; Johnson et al. 1977; Johnson 1987). The pre-dam floods that 

reworked alluvial sediments were abruptly reduced (fig. 27), as was the annual pulse 

of fine sediment, by approximately 80 percent, which came from the Colorado River 

watershed upstream of Lees Ferry (Topping et al. 2000; Rubin et al. 2002). Riparian 

vegetation along the former high flood line was disconnected from the river and 

formerly active channel bedforms suddenly became ideal sites for successful establishment of riparian vegetation, both native and nonnative.
USDA Forest Service RMRS-GTR-377. 2018 153

Accounts in the late 1960s by P.S. Martin (as cited in Johnson 1991 and Webb 

1996) attest to the “explosive spread” of tamarisk through Grand Canyon. No doubt 

there was competition among tamarisk and various native riparian species including 

sandbar willow, Goodding’s willow, Fremont cottonwood, and several shrub species 

that rapidly colonized newly available establishment sites (Webb 1996). Photos from 

the mid-1970s show the encroachment of dense stands of riparian vegetation onto 

formerly scoured alluvial surfaces (Turner and Karpiscak 1980; Webb 1996). Although 

many newly established stands of riparian vegetation were dominated by tamarisk, 

mixed stands of tamarisk, sandbar willow, and seep-willow also were present. In 

certain settings, riverine marsh vegetation became established (Stevens et al. 1995). 

As has been documented downstream of dams on other rivers (Williams and Wolman 

1984), the channel of the Colorado River canyons below Glen Canyon Dam narrowed. The rapid establishment of riparian vegetation created new habitat for a variety 

of wildlife species, including birds (Willson and Carothers 1979). These dramatic 

transformations in downstream aquatic and riparian ecosystems, coupled with rapidly 

escalating human use of the river corridor, raised concerns (Dolan et al. 1974). Efforts 

were soon underway to understand the nature of these changes with the hope that 

improved understanding would lead to anticipation of future change and improved resource management in Grand Canyon National Park immediately downstream of Glen 

Canyon Dam (Johnson 1991; Schmidt et al. 1998).

First Predictions of Change

After completion of Glen Canyon Dam, Turner and Karpiscak (1980) matched 

photos taken before the dam with those taken in the early to mid-1970s. The most 

striking change was the widespread and rapid establishment of new riparian plant assemblages. Dense stands were often dominated by tamarisk but included many native 

riparian and riverine marsh species. They speculated that the new riparian community 

was still assembling and, based on the recent spread of camelthorn (Alhagi maurorum) from the Little Colorado River and observed occurrences of both Russian olive 

(Elaeagnus angustifolia) and Siberian elm (Ulmus pumila), predicted an increased 

importance for these nonnative species in the riparian vegetation community.

Turner and Karpiscak (1980) observed local declines in the density of plants 

in the old high-water zone. Based on the fact that flooding would no longer directly 

inundate these zones, they predicted a more generalized decline of species from the old 

high-water zone. At the same time, they noted species like catclaw and honey mesquite 

establishing with tamarisk and sandbar willow in the newly established riparian communities and predicted that they would become a minor but important component of 

the post-dam riparian community known as the “new high-water zone.” Finally, they 

concluded that prediction of future conditions was not possible given the apparent rapid 

flux in the new riparian communities and key unknowns like streamflow, which was 

now determined by power and water-supply demands, as well as water availability. 

Acknowledging the unpredictability of regional climate, they pointed to the possible effects of drought and excess moisture on release patterns from the dam and the effects of 

prolonged low flows or inundation on the newly formed communities. Only the passage 

of time would refine our understanding of these relationships.
154 USDA Forest Service RMRS-GTR-377. 2018

Following the largest post-dam floods on record in the mid-1980s (figs. 27 and 28), 

Stevens and Waring (1985) recorded high mortality in the new high-water zone riparian 

vegetation. They also noted significant differences in post-flood establishment. Tamarisk 

was five times more abundant than any other species on reworked flood surfaces, and 

clonal species like arrow-weed, sandbar willow, and common reed (Phragmites australis) rapidly colonized disturbed surfaces. They suggested that given different mortality 

rates and colonization patterns across species, longer-term compositional shifts in riparian vegetation were likely to result from large, long-duration post-dam floods. These 

results illustrate the points made earlier by Turner and Karpiscak (1980) that only by 

understanding the response of the new riparian ecosystem to the range of possible postdam flows over time can we begin to predict future conditions.

From 1989-1992, Webb (1996) matched all of the 445 original Stanton expedition 

photos from 1889-1890 in what remained of Glen Canyon as well as Marble and Grand 

Canyons. Among other things, Webb’s work with his colleagues provides a spatially 

comprehensive and quantitative assessment of change in riparian vegetation against 

pre-dam conditions and serves to evaluate trends detected by Turner and Karpiscak 

(1980) 20 years earlier. Despite high mortality recorded for riparian vegetation in 1983-

1984 below the 1,400 cms (49,440 cfs) stage, these photos document the pervasive 

presence of post-dam riparian vegetation assemblages in the new high-water zone, 

including stands of riverine marsh habitat in wider canyon settings (Stevens et al. 1995). 

Vegetation was sparse only in settings where currents were swift and there was limited 

availability of fine sediment (Webb 1996). He argues that the rapid spread of tamarisk 

following dam closure was facilitated by source populations in major tributaries where 

it had long been established (Clover and Jotter 1944; Dodge 1936). However, Stevens 

(1989) observed that sandbar willow’s ability to spread from root sprouts could ultimately lead to a gradual replacement of tamarisk by willow in the post-dam setting.

Establishment of old high-water zone species like honey mesquite and catclaw 

in the new high-water zone, noted by Turner and Karpiscak (1980), was confirmed by 

Webb (1996). He also documented the loss of desert broom from the old high-water 

community and new occurrences in the new high-water zone in the lower canyon. 

Because of relatively slow establishment and growth rates of species like honey mesquite and catclaw, mature stands of these trees in the new riparian assemblage could 

take decades to develop (Webb 1996). Finally, Webb found no clear evidence suggesting 

a decrease in honey mesquite densities in the old high-water zone over the preceding 

century but noted crown dieback in many individuals.

Methods

Repeat photography is the art and science of using ground-based photography to 

match historical images of landscapes. This approach has long been used to qualitatively 

and quantitatively assess landscape change (Webb et al. 2010), including in Grand 

Canyon (Turner and Karpiscak 1980; Webb 1996). Repeat photographs where views are 

a century or even a few decades apart are not suitable for resolving features or processes 

that respond on shorter time steps. However, the intervals of our photos are appropriate 

for examining increases in the number or size of long-lived individual plants in the riparian zone like catclaw acacia or changes in the distribution and size of riparian species 
USDA Forest Service RMRS-GTR-377. 2018 155

that tend to grow in dense clumps like sandbar willow where individuals are difficult 

to identify. Our goal was to quantify change in the abundance of key riparian species 

and the overall cover or biomass of riparian vegetation resulting from long-term, nonperiodic change in streamflow and precipitation.

Photographs produced by the Stanton railroad survey expeditions of 1889 and 

1890 through Marble and Grand Canyons were matched in 1989-1992 (Webb 1996). In 

2010-2012, all photographs were again matched approximately 20 years following the 

first match using the same cameras and techniques described in Boyer et al. (2010). At 

each camera station, plants were identified and changes in riparian vegetation between 

the original and first matched photos, compared to the present, were interpreted in 

the field and recorded as annotations on hard copies of those images. High-resolution 

digital scans of these images were used in conjunction with field annotations to quantify 

changes in riparian vegetation.

We examined in detail 160 sets of matches for a total of 492 individual photos 

(an online Appendix table and all photos are available at: http://dx.doi.org/10.5066/

F76W988R). These included eight matched photo sets with additional dates (e.g., 1975 

and 1989) that bracketed the high, post-dam flows (1983 to 1986), allowing us to make 

inferences about riparian vegetation dynamics over shorter time steps and in response 

to relatively large post-dam flows. We selected this subset of matched photos to analyze 

because the images generally featured clear views of riparian vegetation along canyonbound portions of the Colorado River through Glen, Marble, and Grand Canyons. Our 

selected photo points provide a broad representation of physical settings in the canyons 

examined (fig. 26). We focused our analysis on changes in the number of individuals 

and cover of key riparian species, which could be reliably identified in the imagery as 

well as general changes in the cover or biomass of riparian vegetation in the field of 

view. Change in number of individuals and biomass were assessed visually as a categorical increase, no change, or a decrease. Because the photo match comparisons between 

1989-1992 and 2010-2012 were made between leaf-off and leaf-on conditions, changes 

in state had to be unambiguous and typically involved the loss or addition of individual 

plants or groups of plants. We include photo matches as figures that illustrate changes in 

riparian vegetation across a wide range of fluvial geomorphic settings.

Results and Discussion

Our photographic images fall within three distinct regimes relative to flow regulation in Grand Canyon (fig. 28). The original Stanton photos were from the period 

of largely unregulated flow prior to construction of Glen Canyon Dam. Photos from 

the 1970s and 1980s and the first full set of repeat photos (Webb 1996) were taken 

during the period of unrestricted hydropeaking (figs. 27 and 28). The second set of 

repeat photos was taken during the period of restricted hydropeaking and prescribed 

experimental floods (fig. 28). Of note is the fact that the original Stanton photos were 

taken 5 to 6 years following the occurrence in 1884 of the largest historical flood 

known on the Colorado River, with an estimated discharge of about 5,946 ± 850 cms 

(210,000 ± 30,000 cfs) at Lees Ferry (Topping et al. 2003). The views in 1990 occurred at the end of a regionally wet period and the 1983 flood with a peak discharge 

of 2,621 cms (92,560 cfs), whereas the views in 2010-2012 followed a period referred 
156 USDA Forest Service RMRS-GTR-377. 2018

to as the early 21st century drought (Woodhouse et al. 2010). Finally, the 2010-2012 

views were taken following a controlled flood with a discharge of 1,175 cms (41,500 

cfs). Results of our photo matches summarizing overall change in riparian vegetation 

over the past 20 years are presented in table 8. Corresponding changes for key riparian species are included in table 9. An online Appendix (http://dx.doi.org/10.5066/

F76W988R) provides a table listing information and the interpretation results for each 

photo station along with all the photos examined.

Turner and Karpiscak (1980) predicted increases in the invasive species Russian 

olive and Siberian elm along the Lower Colorado River. Such increases have been 

large elsewhere, and Russian olive is now the fourth-most frequently occurring 

woody riparian species in the western United States (Friedman et al. 2005). However, 

neither of these species appeared in our photo matches and only a few Russian olive 

stems were observed upstream of Lees Ferry in 2011. These species do not appear to 

have spread. In the case of Russian olive, the National Park Service (NPS) in Grand 

Canyon began removing Russian olive in the 1990s and continues to remove them 

when observed (Hahn 2011). The removal efforts by NPS appear to have been successful given the rapid spread of Russian olive along smaller, regional streams like 

the Escalante River and at Canyon de Chelly, coinciding with a wet episode beginning 

about 1980 and extending through the 1990s (Reynolds et al. 2014; Webb et al. 2007).

Table 8—The percentage and number of photos, for the periods 1989–1992 and 2010–2012, in which tamarisk was present 

in the matched view. A total of 160 original photos were examined. Categorical state changes in cover (percentage 

increase, decrease, no change) for all woody riparian vegetation; all native woody riparian species; and all old high waterzone (OHWZ) species, are compared over the period 1989–1992 to 2010–2012, based on the number of views in which 

the vegetation category occurred (n). 

Tamarisk 

presence in 

1989–1992 

views 

(n = 160)

Tamarisk 

presence in 

2010–2012 

views 

(n = 160)

Changes in matched views for the interval: 1989–1992 to 2010–2012

(number of views in which the vegetation category occurred)

Increase 

in woody 

riparian 

vegetation

(n = 159)

No change 

in woody 

riparian 

vegetation

(n = 159)

Increase 

in woody 

riparian 

natives

(n = 157)

No change 

in woody 

riparian 

natives

(n = 157)

No change 

in OHWZ 

species

(n = 48)

Decrease 

in OHWZ 

species

(n = 48)

% 79.5 96.9 89.3 9.4 87.3 11.5 43.8a 18.8a

n 128 156 142 15 137 18 21 9

a State change for OHWZ species could not be determined in 38% of the views in which they occurred (n = 18).

Table 9—The number and percentage of views in which key riparian species increased in cover during the 

interval: 1989–1992 to 2010–2012. A total of 160 original photos were examined.

Increases in riparian species during the interval: 1989–1992 to 2010–2012

Baccharis spp. Salix exigua Pluchea sericea Acacia gregii Prosopis glandulosa

% 53.4 1.9 13 10.6 6.8

n 86 3 21 17 11
USDA Forest Service RMRS-GTR-377. 2018 157

New High-Water Zone

The original Stanton photos show little or no vegetation below the old high-water 

zone. These surfaces, particularly in narrow canyon settings, were apparently reworked 

frequently enough by pre-dam flows that they were largely devoid of vegetation, except 

in a few protected settings (see figs. 29-33). Turner and Karpiscak (1980) qualitatively 

described widespread establishment of riparian vegetation on formerly active channel 

deposits in the mid-1970s below Glen Canyon Dam. By 1973, the greatest increase in 

vegetated area (up to 20 percent) occurred in those areas of the channel above the 878 cms 

(~31,000 cfs) stage, or the maximum hydropower peaking flows, and remained 10 percent or less below this stage (Sankey et al. 2015). As predicted by Turner and Karpiscak 

(1980), the vegetation established below the old high-water zone following flow regulation was susceptible to removal by subsequent high flows. It would not take long for the 

effects of flooding and prolonged inundation on the newly formed riparian ecosystem to 

be realized.

The largest flood in the post-dam record occurred in 1983, resulting from a series 

of late snow and rainstorms and rapid runoff in the basin. The flood peaked at 2,621 cms 

(92,560 cfs) and was of unusually long duration. Comparatively large, long-duration 

floods followed in 1984 and 1986 (fig. 28). Mortality, resulting from physical removal 

or prolonged inundation, exceeded 45 percent at some locations for tamarisk, honey 

mesquite, and catclaw, which had established below the 1,700 cms (~60,000 cfs) stage 

in the new high-water zone (fig. 29). Clonal species like sandbar willow and arrow-weed 

had lower mortality rates of 7 percent and 33 percent respectively. Shallowly rooted species like the seep-willow (Baccharis spp.) suffered 90 percent mortality (Johnson 1991; 

Stevens and Waring 1985).

The flooding in 1983-1986 killed or removed up to 10 percent of all riparian vegetation (Waring 1996), particularly at lower river stages (fig. 29), and vegetated cover 

below 1,274 cms (~45,000 cfs) showed little or no increase over the interval from 1973 to 

1984. Presumably, any increases prior to the flood were offset by mortality from the flood 

(Sankey et al. 2015). Because of species-specific differences in flood mortality and early 

differences in post-flood reestablishment patterns, Stevens and Waring (1985) suggested 

that post-dam flooding from 1983 to 1985 had the potential to shift composition in favor 

of species like tamarisk and catclaw acacia and away from species like seep-willows and 

longleaf brickellbush (Brickellia longifolia). By the 1990s, tamarisk remained a prominent 

component of the new riparian assemblage and it appeared in 71 percent of the views of 

the river corridor taken by Webb (1996).

In August 1991, operating rules for power generation at Glen Canyon Dam changed 

and maximum releases were limited to 708 cms (~25,000 cfs) (US DOI 1996; fig. 28) and 

by 1992, vegetated cover values above this stage and below the maximum stage of the 

1983 flood increased to between 25 and 30 percent. Values remained at about 10 percent 

below the new hydropower maximum (Sankey et al. 2015). Three short-duration, highflow experiments (HFEs) occurred in 1996, 2004, and 2008 (fig. 28). By 2002, cover had 

declined in the zone between the maximum HFE flood and the 1983 flood and remained 

low through 2009. In contrast, cover increased from 5 to 15 percent below the peak stage 

of the HFE floods, particularly between 2005 and 2009 (Sankey et al. 2015). Our photo 

matches in 2010–2012 are consistent with the aerial imagery analysis results of Sankey et 
158 USDA Forest Service RMRS-GTR-377. 2018

al. (2015) and show increases in riparian vegetation in 90 percent of the views compared 

with 1989–1992. Tamarisk was present in 79.5 percent of the 1989–1992 views we examined (compared with 71 percent of the total matched population examined by Webb 1996) 

and increased to a presence of 96.9 percent by 2010–2012 (table 8).

There was no indication that sandbar willow was replacing tamarisk, as it increased 

in only 1.9 percent of the views (table 9). This contrasts with earlier speculation that 

sandbar willow might gradually replace tamarisk under the post-dam flow regime (Stevens 

1989). Conversely, our data suggest that the recent increases in riparian vegetation at lower stages reported by Sankey et al. (2015) involve a number of other native species such as 

seep-willow. Overall, native woody species increased in 87.3 percent of our views, with 

seep-willow increasing in 53.4 percent of the views over the past 20 years (tables 8 and 9). 

Seep-willow suffered high flood mortality in the mid-1980s and had lower representation 

in 1990 matched photos. Much of the increase in tamarisk, seep-willow, and arrow-weed 

Figure 29—Tamarisk in Glen Canyon 

(photo 751; online Appendix, http://

dx.doi.org/10.5066/F76W988R). 

(a) (top) (1889) This downstream 

view, 11 river kilometers above Lees 

Ferry, shows a remarkably clear view 

of fine-grained alluvium draped 

over talus. Except for a few scattered 

plants near the top of the bank, the 

deposit is largely free of vegetation. 

(b) (bottom) (1975) Within 12 years 

following closure of Glen Canyon 

dam, the bank is heavily vegetated 

with primarily tamarisk and sand-bar 

willow. The large tamarisk stem, high 

on the bank, may have established 

during high flows before the dam. 

(c) (next page) (1989) Near-shore 

stands of tamarisk and willow appear 

to have been removed or killed-back 

by flooding a few years preceding 

this view. (d) (next page) (2011) 

Woody riparian vegetation, including 

tamarisk, seep-willow, and sand-bar 

willow, have reestablished near shore 

and a diversity of grasses, herbs, and 

small shrubs fill spaces in between. 

The large tamarisk on the high bank is 

still alive but has many dead branches 

in the crown.
USDA Forest Service RMRS-GTR-377. 2018 159

occurred in near-shore locations (figs. 30 and 31). Sankey et al. (2015) found that vegetation expansion into lower elevations of the riparian zone occurred during periods when 

peak flows were lower and base flows higher and when inundation duration fell below 

about 5 percent.

Any role that HFEs (fig. 28) may have played in the recent encroachment of 

riparian vegetation is unclear. A more explicit linkage between the increase in native 

and nonnative woody shrubs and specific flow events is difficult to make without much 

shorter time steps between photo matches that tightly bracket the events. Data presented 

by Sankey et al. (2015) suggest an increase in near-shore vegetation sometime between 

2005 and 2009 (see his figure 6). However, whether this was an event-specific response 
160 USDA Forest Service RMRS-GTR-377. 2018

to the 2008 HFE or simply subtle shifts in inundation duration during the period can 

only be answered by detailed vegetation monitoring in real time or by aging encroaching vegetation retrospectively.

The Glen Canyon Dam Adaptive Management Program (www.gcdamp.gov) 

conducted three HFEs in 1996, 2004, and 2008. Results from the 1996 and 2008 

HFEs indicate that some near-shore wetland plants were removed by these flows and 

woody riparian plants were buried by sediment, which temporarily reduced vegetation cover. Losses, however, were short-lived (12 months or less). Recovery favored 

Figure 30—Channel narrowing 

in Marble Canyon (photo 

2526; online Appendix, 

http://dx.doi.org/10.5066/

F76W988R). 

(a) (top) (1889) View 

upstream into still water 

from the upstream edge of 

the South Canyon debris 

fan (river kilometer 51). A 

channel margin deposit of 

fine alluvium along the near 

wall is devoid of vegetation 

and is marked by a series 

of flow recession lines. 

(b) (bottom) (1992) The 

channel margin deposit 

has expanded and is now 

stabilized by tamarisk, sandbar willows and grasses. 

Leafless, small stems of what 

may be sand-bar willow 

or tamarisk grow along the 

edge of a secondary channel 

adjacent to a cobble bar 

with small tamarisk lining 

the stream ward edge. 

(c) (next page) (2010) A 

dense stand of tamarisk, 

sand-bar willow, and 

seep-willow now flank 

the narrowing secondary 

channel and the cobble 

bar is largely covered 

by vegetation and fine 

sediment.
USDA Forest Service RMRS-GTR-377. 2018 161

woody and herbaceous species that were able to reproduce vegetatively, like sandbar 

willow. Because the timing of these flows occurred predominantly before the seed 

dispersal window for tamarisk, the establishment of this species following the releases 

was comparatively low (Kennedy and Ralston 2011). Strategies for subsequent HFEs 

have focused on building and maintaining sandbars from tributary-derived sand and 

balancing the effects of releases in order to limit competition between nonnative trout 

(Oncorhynchus mykiss) and native humpback chub (Gila cypha) populations (Wright 

and Kennedy 2011). The timing of HFEs, which is largely driven by sediment inputs 

from tributary sources and competitive interactions between fish, could have important 

implications for riparian plant species that reproduce by seed. This is especially true for 

woody riparian pioneer species like native cottonwood, willow, and nonnative tamarisk. 

Because these species have partially non-overlapping seed dispersal periods, subtle 

differences in flood timing within these seed release periods can have profound influences on the composition of newly established seedling cohorts (Shafroth et al. 2010). 

Continued monitoring of wetland and riparian vegetation relative to HFEs is important 

to understanding ecosystem level responses to flow management in the Grand Canyon.

Only ~1 percent of the views showed a decrease in woody native riparian species, 

whereas 11.5 percent of the views showed no change in cover and density of these species (table 8). The images reveal that in certain narrow canyon settings, shear stresses 

are likely to remain high enough under the current flow regime to preclude colonization 

and persistence of native riparian vegetation (fig. 32). Thus, unless the flow regime 

changes in the future, certain channel locations should remain free of persistent vegetation. Finally, catclaw acacia and honey mesquite increased in the new high-water zone 

in 10.6 percent and 6.8 percent of our views, respectively (table 9). For both species, 

this pattern was especially pronounced beginning at about rkm 105 and some stands 

are now several meters tall (fig. 33). The widespread occurrence of tamarisk, honey 

mesquite, catclaw acacia, and a number of other native and nonnative riparian plants in 
162 USDA Forest Service RMRS-GTR-377. 2018
USDA Forest Service RMRS-GTR-377. 2018 163

Figure 31—Vegetation encroachment, upper Grand Canyon (photo 1749 b; online Appendix, http://dx.doi.org/10.5066/

F76W988R). 

(a) (1890) Looking downstream to a sand and gravel bar on river right, upstream of the Papago Creek debris fan, upper 

Grand Canyon (river kilometer 123). Both fan and bar features are low and devoid of vegetation. (b) (1991) Tamarisk 

has colonized the higher portions of the sand bar and the Papago Creek fan, which appears to have increased slightly in 

elevation. (c) (2011) Tamarisk has increased in size and number on both the bar and debris fan. On the bar, dense stands 

of tamarisk, sand-bar willow, arrow-weed, and seep-willow have filled in between the older, higher tamarisk and the 

channel bank. Tamarisk and seep-willow also line the channel margin upstream of the fan as well as the upstream edge of 

the debris fan.

the new high-water zone predominantly reflect natural establishment and survival in the 

post-dam period. However, nonnative Russian olive and tamarisk have and are being 

removed within Grand Canyon National Park.

At the same time, native plants have been planted in various locations, including 

cottonwoods at Lees Ferry (rkm 0) (Stevens and Burke 2001), various native plants at 

Soap Creek Camp (rkm 18.5) (Theobald 2001), and honey mesquite and catclaw acacia 

at Nankoweap (rkm 85) and Granite Park (rkm 336) (Anderson and Ruffner 1987). 

Additionally, in 2012–2013, the NPS initiated a riparian restoration project at Granite 

Camp, upstream of Monument Creek (rkm 151) (McMaster 2012). This effort included 

removing tamarisk and planting honey mesquite, along with other woody native species (https://www.nps.gov/grca/learn/nature/granite.htm). Similar restoration efforts 

are being planned as part of the Long-term Experimental Management Plan for the 

operation of Glen Canyon Dam (https://ltempeis.anl.gov). The presence of sites in the 

new high-water zone, where vegetation has been established and maintained artificially, 

or removed, could potentially muddle future interpretations of vegetation responses to 

flow management, particularly if the increasing number of restoration activities are not 

systematically tracked into the future.
164 USDA Forest Service RMRS-GTR-377. 2018
USDA Forest Service RMRS-GTR-377. 2018 165

Figure 32—Stasis in upper Granite Gorge (photo 2319; online Appendix, http://dx.doi.org/10.5066/F76W988R). 

(a) (1890) Upstream view of the Granite Gorge from Clear Creek (river kilometer 136). The walls of Vishnu schist and 

Zoroaster granite are too steep and narrow to support any riparian vegetation. (b) (1991) After a century and the closing 

of Glen Canyon dam, the inner gorge appears little changed. (c) (2011) In spite of large increases in riparian vegetation 

in wider canyon reaches, some steep, narrow canyon settings are not likely to provide conditions for establishment and 

persistence, even under current regulated flow conditions.

Old High-Water Zone

In 1973, Turner and Karpiscak (1980) saw limited evidence for declines in trees in 

the old high-water zone. Aside from crown dieback and increases in density for honey 

mesquite, Webb (1996) found little compelling evidence suggesting widespread declines 

in the old high-water stands; catclaw acacia, for example, seemed little affected. Similarly, 

results of Sankey et al. (2015) show little directional change in vegetated cover in this 

zone from 1965 through 1992 but sustained decreases starting in 2002. They point to the 

early 21st century drought, beginning in the late 1990s, as a likely cause. This is consistent 

with Webb’s predictions that following flow regulation species like honey mesquite would 

be susceptible to drought and undergo slow structural decline (Webb 1996).

In 72 percent of the initial matched views by Webb (1996) containing honey mesquite (n = 58), the trees either increased or showed no change in density. In 10 percent 

of the views with density increases, there were corresponding decreases in plant stature, 

suggesting possible climate-related factors like drought and frost. Decreases in severe 

frosts after 1890 may have spurred new establishment in some settings. Increases in 

precipitation and streamflow in the early 20th century (Hereford et al. 2006; fig. 27) could 
166 USDA Forest Service RMRS-GTR-377. 2018
USDA Forest Service RMRS-GTR-377. 2018 167

Figure 33—(left and above) A new high water line in lower Grand Canyon (photo 1772 a; online Appendix, http://dx.doi.

org/10.5066/F76W988R). 

(a) (1890) Looking upstream and across the upper portion of the fan at Parashant Wash (river kilometer 320). A depression 

filled with fine sediment at the lower right of the image has a patch of sizable vegetation, which is not identifiable. 

Upstream, and on both sides of the river, a distinct and largely continuous band of woody vegetation can be seen above 

the old high water line. (b) (1991) Bermuda grass dominates that portion of the fan, which contained fine sediment and 

vegetation in the previous image. Scattered, taller tamarisk and desert broom occur on the upper portion of the fan. 

Stunted and dead tamarisk occupy lower portions of the fan, possibly reflecting flood mortality from the mid-1980s floods. 

Across the river, a dense line of tamarisk occurs below the band of old high water vegetation, which appears to have 

thinned and some dead stems are apparent. (c) (2011) Tamarisk higher on the bar appears stressed and is dying back and 

the desert broom is gone. Closer to the river, a dense band of woody vegetation, dominated by mesquite, is tall enough to 

obscure the river. 

have contributed to height growth, which might then have been lost to drought-pruning 

following dam closure (Webb 1996). Although Webb thought it unlikely that honey mesquite would be lost from this relict riparian assemblage over the short-term, he suggested 

disconnection from the river in the post-dam era might lead to long-term reproductive failure and that future meteorological drought could reduce existing honey mesquite stands to 

scattered individuals.

In the 2010–2012 matches that we examined, the old high-water zone occurred 

in 30 percent (n = 48) of our views (table 8). In 18 of those views, an evaluation of 

change in this pre-dam riparian plant assemblage was not possible, primarily because 

plant growth in the new high-water zone over the past 20 years obscured the view 

of the old high-water zone (see fig. 33). In spite of a relatively small number of unobstructed views of the old high-water zone (n = 30), our results generally support 

the observations of Sankey et al. (2015) that this zone showed declines in overall 

cover beginning in 2002. In our 2010–2012 matches, none showed increase in cover 
168 USDA Forest Service RMRS-GTR-377. 2018

of this vegetation. Rather, cover declined in 19 percent of our views and showed no 

discernable change in 44 percent of our views (table 8). This suggests that in spite of 

generalized regional drought over the past decade (Hereford et al. 2002), factors such 

as aspect, microclimate, and side-slope drainage in high-relief canyons may buffer 

some stands of old high-water zone vegetation from the effects of drought. Some of 

these stands may in fact have become established because of the presence of shallow 

groundwater of local origin (e.g., the stands at Ledges Camp, rkm 245) and may therefore be largely unaffected by river flows.

Plant pathogens and unintended introductions of insect pests have had unexpected and profound influences on long-term compositional changes in riparian forest 

ecosystems (Johnson et al. 2012). A new factor potentially influencing the long-term 

distribution and importance of tamarisk in riparian systems throughout the western 

United States is the purposefully introduced tamarisk leaf beetle (Diorhabda spp.; see 

Chapter 5). Monitoring data from the Colorado River Basin show that the beetle was 

present throughout Marble and middle and lower Grand Canyon by 2011 and 2012 

(http://www.tamariskcoalition.org/sites/default/files/files/2014 percent20Distribution 

percent20Map percent20CRB.pdf). We observed tamarisk defoliation and premature 

leaf senescence related to beetle activity in 3.7 percent of our views. Anticipating 

the effects of the beetle on the structure and composition of riparian vegetation assemblages in the Grand Canyon is not yet possible and, as stated earlier by Turner and 

Karpiscak (1980), will require the passage of time.

Our recent work provides another point-in-time assessment of riparian vegetation 

change in canyons of the Colorado River influenced by Glen Canyon Dam. This is an 

effective way to monitor changes that unfold slowly but is inadequate for capturing 

the details of processes that occur more rapidly, such as event-specific mortality or 

establishment and removal processes on dynamic geomorphic surfaces. Such refinements require a more mechanistic understanding that explicitly links vegetation with 

hydrology and climate. Relationships between vegetation and streamflow have been 

used to predict vegetation response to flow alterations (Auble et al. 1994; Primack 

2000). Using surveyed vegetation plots and a hydrologic model, Auble et al. (2005) 

modeled individualistic plant responses to inundation duration. This relation was used 

to predict changes in vegetation patterns away from the channel following changes in 

flow. Because of limits to the amount of topographic data that Auble et al. (2005) could 

collect, their work was limited to a short reach and could not account for longitudinal 

variation in the flow-vegetation relationship. However, Sankey et al. (2015) have recently developed fine-resolution spatial and temporal relationships between variations 

in riparian vegetation cover relative to streamflow and climate. They accomplished this 

using aerial imagery, instantaneous flow and topographic data, and a spatially extensive 

hydrologic model of Marble and Grand Canyons. These data provide the basis for 

predicting future vegetation change in this system. Thus, we examine our findings and 

earlier predictions in the context of these new results. We used photographic data at a 

scale fine enough to identify changes in numbers of individuals of particular species. 

This complements the analysis of Sankey et al. (2015), which was based largely on 

aerial imagery that was spatially more extensive but lower in resolution.
USDA Forest Service RMRS-GTR-377. 2018 169

Management Implications

1. Tamarisk was present in Glen, Marble, and Grand Canyons in the pre-dam period 

(Clover and Jotter 1944). With other native riparian species, it spread rapidly, likely 

from both tributary sources and existing stands along the Colorado, throughout the 

Grand Canyon following dam closure. Despite extensive mortality following flooding 

from 1983–1986, tamarisk is still a dominant riparian species. Based on photographic 

evidence, it was widespread by 1989–1992 and was found in 71 percent of the views 

examined by Webb (1996) and in 79.5 percent of the views we examined for the same 

time period.

2. Woody riparian vegetation has increased throughout the river corridor in Grand Canyon 

between the 1990s and 2012. In general, total woody riparian vegetation, including 

tamarisk, showed increased cover and density during this period in 89.3 percent of 

the matched images in 2010–2012. Recent work shows that vegetation expansion into 

lower topographic position in the riparian zone occurred during periods when peak 

flows were lower and base flows higher and when inundation duration fell below about 

5 percent (Sankey et al. 2015). There was no evidence that brief, pulsed inundation 

during HFEs limited vegetation expansion. Gains in riparian vegetation cover over the 

past two decades were primarily below the maximum stage of post-dam controlled 

floods (1,274 cms [~45,000 cfs]), especially in near-shore locations for species like 

tamarisk, seep-willow, and arrow-weed.

3. Whether HFEs have been a factor in the recent expansion of riparian vegetation is 

unknown and will require careful, real time monitoring of vegetation following HFEs 

or retrospective analyses examining the age structure of encroaching vegetation. The 

seasonal timing of HFEs also warrants further study as they could preferentially shift 

the structure and composition of riparian vegetation.

4. In contrast to observed vegetation increases, 9.4 percent of the rematches in 2010–2012 

show no apparent change, and <2 percent of the views show a decrease in woody 

riparian vegetation cover and density. Under the current flow regime, some narrow 

canyon settings are likely to remain free of persistent riparian vegetation.

5. Some of the increase in riparian vegetation results from continued spread and growth 

of tamarisk, but an important component of the increase includes the establishment of 

native woody riparian species, including old high-water line species, such as catclaw 

acacia, honey mesquite, and desert broom, in the new high-water zone or below the 

stage of the 1983 flood (2,747 cms [~97,000 cfs]).

6. In addition to the natural establishment of riparian vegetation in the new high-water 

zone, nonnative species like tamarisk and Russian olive have been removed and native 

vegetation planted as part of experiments or restoration efforts in Marble and Grand 

Canyons. Artificial manipulation of vegetation should be tracked, actions and locations 

clearly documented, and the information easily accessible, so that cause-and-effect 

relationships between flow management and vegetation response can be assessed as 

accurately as possible.

7. Disconnection of the old high-water assemblage from river flows has left species, 

like honey mesquite, vulnerable to drought stress, and recent evidence suggests that 
170 USDA Forest Service RMRS-GTR-377. 2018

vegetated cover has declined during the period of early 21st century drought, beginning 

in the early 1990s. The work of Webb (1996) suggests species like catclaw acacia may 

be less affected. Some stands may be buffered against drought by factors such as aspect 

and local ground-water sources.

8. The effects of the recently introduced tamarisk leaf beetle on long-term structure and 

composition of riparian assemblages in Glen, Marble, and Grand Canyons remain 

uncertain.

Acknowledgments

We thank the many individuals who helped with the field work on this project. 

Diane Boyer and Helen Raichle both supervised student workers and scanned negatives 

themselves to create a fully digital photographic archive used in our analyses. Grand 

Canyon National Park provided logistical support for our river trip in 2010, and Grand 

Canyon Monitoring and Research Center provided additional funding for logistics and 

salaries. Additional funding was provided by the U.S. Geological Survey. Ryan Morrison, 

USGS, provided helpful review comments. Any use of trade, firm, or product names is for 

descriptive purposes only and does not imply endorsement by the U.S. Government.

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174 USDA Forest Service RMRS-GTR-377. 2018

Chapter 10. Breeding Waterbirds of the Mexican Portion 

of the Colorado River Delta

Eric Mellink and Osvel Hinojosa-Huerta

Introduction

Once a mighty and wild river with abundant wetlands, the section of the Colorado 

River flowing through Mexico has become a trickle ... whenever it flows. Most of the 

time since the 1960s, until recently, it did not and was completely dry. This brought tremendous changes in the “original” constitution and biological processes of the region, 

although they have not been fully investigated.

Although any fine-scale reconstruction of the historic characteristics of the 

Colorado River Delta is impossible, wetlands were widespread over an area of 

500,000 ha in what in part is today the Mexicali Valley in Baja California and Sonora. 

Luxurious riparian forests existed in dozens of braided channels in the deltaic alluvial 

floodplain and the influence of the river extended 65 km into the Gulf of California 

(Felger 2000; Leopold 1953; Rodriguez et al. 2001; Sykes 1937).

However, this image must be seen with a note of caution. What is usually considered the natural conditions of the Lower Colorado River is not accurate, as the 

picture of large amounts of sediment flowing into the sea through the Colorado River 

documented by Sykes (1937) was flawed. These large sediment loads were atypical, 

because they resulted from extensive loss of beaver (Castor canadensis) dams from the 

river’s tributaries due to trapping and from overgrazed watersheds. This began with the 

arrival of the Spaniards during the 18th century (Dobyns 1978, 1981). These increased 

sediments must have changed the physiognomy of the lower delta, although the details 

and consequences remain unknown.

In addition, in the last 100 years, the reduction of freshwater flows and other hydrological changes in the basin have caused biotic degradation of the delta, to the point 

where wetlands now encompass only about 10 percent of the previous area (Glenn et 

al. 2001; Zamora-Arroyo et al. 2005). Although some diversions and dams were constructed in the early 1900s, the major changes in the delta began with the construction 

of Hoover Dam in 1936 and Glen Canyon Dam in 1963 (Zamora-Arroyo et al. 2005). 

These reservoirs were filling up until 1979, capturing flows that would otherwise reach 

the delta, driving the habitat degradation of the region (Glenn et al. 2001).

However, even in their current reduced and altered state, these wetlands continue 

to provide critical habitat for endangered species and for thousands of migratory 

waterbirds (Gomez-Sapiens et al. 2013; Hinojosa-Huerta et al. 2007). The area has 

been recognized as a Wetland of International Importance by the Ramsar Convention 

and a Site of Hemispheric Importance by the Western Hemisphere Shorebird Reserve 

Network (Hinojosa-Huerta and Carrillo-Guerrero 2010). This biological richness has 

been maintained by the persistence of a diversity of habitat types in the area, ranging 

from riparian forest and freshwater streams to brackish marshes, tidal mudflats, and 
USDA Forest Service RMRS-GTR-377. 2018 175

the estuarine/marine environment of the Upper Gulf of California (Hinojosa-Huerta et 

al. 2007).

Since the late 1970s, periodic river releases, agricultural return flows, and treated 

effluent of wastewater plants have revitalized portions of the delta, and the populations 

of some waterbird species have responded accordingly (Glenn et al. 2001; HinojosaHuerta et al. 2004). This resiliency in the delta has been attributed to a number of key 

elements, including (1) the maintenance of undeveloped land in the river corridor and 

lower delta, (2) the conservation of a geomorphologically active floodplain, (3) the 

influence of the large tides of the Upper Gulf of California, and (4) the maintenance 

of some flows of water as river releases, agricultural drainage, and groundwater flows 

(Glenn et al. 2013; Nagler et al. 2005; Nelson et al. 2013).

In this chapter, we present information on the current status of breeding waterbirds 

in the Colorado River Delta in each of the habitat types in the region, including information on historical changes, current threats, and conservation recommendations.

Fresh Water, Riparian, and Marsh Habitat

Riparian and riverine habitats have been affected the most in the lower Colorado 

Delta, as they depend on freshwater flows. Nevertheless, approximately 1,200 hectares 

of cottonwood (Populus fremontii) and Goodding’s willow (Salix gooddingii) forests, 

with about 47 km of wet channel, remain in the riparian corridor between Morelos Dam 

and the Hardy River, maintained by groundwater flows and sporadic releases from the 

dam (Glenn et al. 2013).

The brackish marshes, dominated by southern cattail (Typha domingensis), are 

in better condition because they benefit from agricultural return flows and can tolerate 

some influence of tidal seawater. Major brackish marshes include the Ciénega de Santa 

Clara, with 5,800 ha of emergent vegetation and shallow lagoons, the Hardy River, the 

El Doctor wetlands, and Laguna del Indio (Zamora-Arroyo et al. 2005). In addition, a 

100-ha open water and marsh habitat in the northern portion of the Hardy River subbasin was recently re-created with treated effluents from Las Arenitas Wastewater Plant, 

30 km south of Mexicali.

Freshwater, riparian, and marsh habitats in the delta support four groups of waterbirds: waterfowl (mostly wintering and migrating), divers (mostly grebes), waders, and 

marshbirds (Patten et al. 2001). Waterbirds were particularly conspicuous before the 

1930s, when major reductions in water flow in the river began (Bancroft 1922; Kramer 

and Migoya 1989; A. Leopold 1949, 1953; A.S. Leopold 1959; Payne et al. 1991; Price 

1899). From the descriptions by Aldo Leopold (1949, 1953) and others it seems that 

the current numbers of between 20,000 and 50,000 ducks and geese, especially in the 

Ciénega de Santa Clara and the Rio Hardy region (Hinojosa-Huerta et al. 2013a), are 

just a pale reminiscence of what the area used to harbor. Four waterfowl species breed 

here: cinnamon teal, redhead, mallard, and ruddy duck, while the fulvous whistlingduck was a common breeder before the habitat changes in the delta (Hinojosa-Huerta et 

al. 2007; Patten et al. 2001; scientific names are in table 10).

Three species of shorebirds, black-necked stilt, American avocet, and killdeer, and 

seven species of waders breed along waterways and marshlands in the Colorado Delta: 

great blue heron, snowy egret, great egret, tricolored heron, cattle egret, green heron, 
176 USDA Forest Service RMRS-GTR-377. 2018

Table 10—Waterbirds that breed, or have bred recently, in the Mexican portion of the Colorado River Delta. The list follows 

the taxonomic nomenclature and order of the seventh edition of the Checklist of North American Birds (American 

Ornithologists’ Union 1998), including the changes made in the 56th supplement to the checklist. Risk categories are: 

Endangered (E, En peligro de extinción), Threatened (T, Amenazado) and under special protection (Pr, sujetos a protección 

especial), according to the Mexican list of species at risk (NOM-059-SEMARNAT-2010).

SpeciesFreshwater /marshUpper estuaryMontague IslandRisk category

Anseriformes

Anatidae: ducks and geese

Mallard (Anas platyrhynchos) x

Cinnamon teal (Anas cyanoptera) x

Redhead (Aythya americana) x

Ruddy duck (Oxyura jamaicensis) x

Podicipediformes

Podicipedidae: grebes

Pied-billed grebe (Podilymbus podiceps) x

Eared grebe (Podiceps nigricollis) x

Pelecaniformes

Phalacrocoracidae: cormorants

Double-crested cormorant (Phalacrocorax auritus) x

Ciconiiformes

Ardeidae: herons, egrets, and bitterns

American bittern (Botaurus lentiginosus) x T

Least bittern (Ixobrychus exilis) x x Pr

Great blue heron (Ardea herodias) x x x

Great egret (Ardea alba) x x

Snowy egret (Egretta thula) x x x

Tricolored heron (Egretta tricolor) x

Reddish egret (Egretta rufescens) x Pr

Cattle egret (Bubulcus ibis) x

Green heron (Butorides virescens) x x

Black-crowned night-heron (Nycticorax nycticorax) x x x

Threskiornithidae: ibises 

White-faced ibis (Plegadis chihi) x

Gruiformes

Rallidae: rails and coots

Black rail (Laterallus jamaicensis coturniculus) x E

Yuma Ridgeway rail (Rallus obsoletus yumanensis) x x T

Virginia rail (Rallus limicola) x Pr

Common gallinule (Gallinula galeata) x

American coot (Fulica americana) x
USDA Forest Service RMRS-GTR-377. 2018 177

and black-crowned night-heron, while little blue heron might do so too (Mellink et al. 

2002; Patten et al. 2001). After a rebound from early 20th century hunting for their 

nuptial plumes (Funcke in Mellink 2000), most breeding species are presumed to have 

reduced their colonies and their numbers, while some wintering species also have reduced their numbers, although recently some new colonies have formed (Mellink et al. 

2002). Conversely, the white-faced ibis has increased its numbers as a result of agriculture (Anderson and Ohmart 1982; Mellink et al. 1997), and the cattle egret, which has 

bred in the Mexicali Valley since the early 1970s, is increasing in numbers in the region 

(Garret and Dunn 1981; Rosenberg et al. 1991).

Pied-billed grebes nest in the marshes of the delta, especially in the Ciénega de 

Santa Clara and the Hardy River, where they are abundant (Hinojosa-Huerta et al. 

2013a). Eared grebes are common winter visitors and were documented nesting in the 

delta for the first time in 2010, at Las Arenitas wetlands, when 75 adults and at least 

seven nests were observed (Erickson et al. 2011). Since then, pairs, nests with eggs, 

and adults feeding juveniles have been recorded every year (Edith Santiago, Sonoran 

Institute, and Eduardo Soto-Montoya, Reserva de la Biosfera Alto Golfo de California y 

Delta del Río Colorado, personal communication).

SpeciesFreshwater /marshUpper estuaryMontague IslandRisk category

Charadriiformes

Recurvirostridae: stilts and avocets

Black-necked stilt (Himantopus mexicanus) x x

American avocet (Recurvirostra americana) x x

Haematopodidae: oystercatchers

American oystercatcher (Haematopus palliatus) x E

Charadriidae: plovers

Snowy plover (Charadrius alexandrinus) x x

Wilson’s plover (Charadrius wilsonia) x x

Killdeer (Charadrius vociferus) x x

Laridae: gulls and terns

Laughing gull (Leucophaeus atricilla) x x

Least tern (Sternula antillarum) x x Pr

Gull-billed tern (Gelochelidon nilotica) x x

Caspian tern (Hydroprogne caspia) x

Forster’s tern (Sterna forsteri) x

Royal tern (Thalasseus maximus) x

Elegant tern (Thalasseus elegans) x Pr

Black skimmer (Rynchops niger) x x

Table 10—Continued.
178 USDA Forest Service RMRS-GTR-377. 2018

Western and Clark’s grebes breed in nearby regions in California and Arizona, and 

they were common breeding visitors in the delta during the 1980s (Patten et al. 2001) 

but have not been recorded breeding in the delta since then.

Seven species of marshbirds breed among the emergent vegetation of the wetlands in the area, and some of them have large populations in the Ciénega de Santa 

Clara (Hinojosa-Huerta et al. 2013b). One of the species of major interest is the Yuma 

Ridgeway rail, protected both in the United States and Mexico, and whose population in 

the Ciénega de Santa Clara is almost 75 percent of the known population of the subspecies, with estimates ranging between 6,000 and 8,000 individuals (Hinojosa-Huerta et 

al. 2001, 2013b). Some minor populations of this rail occur along the marshes of the 

Colorado and the Hardy Rivers and in vegetated agricultural drains throughout the valley (Hinojosa-Huerta et al. 2013b).

The most sensitive marshbird in the delta is the California black rail, which is 

protected as Endangered in Mexico and has fewer than 200 individuals in the region. 

The largest populations of this species in the delta are located at El Doctor wetlands and 

the Ciénega de Santa Clara (Hinojosa-Huerta et al. 2013b). The same Ciénega de Santa 

Clara holds important populations of other marshbirds, including Virginia rail, least bittern, and American bittern, as well as common gallinule and American coot.

The habitat changes in the delta and Lower Colorado reduced the populations of 

marshbirds in this region (Eddleman et al. 1988), but a modest contribution of agricultural return flows (about 140 million m3 per year, or less than 1 percent of the average 

annual flow of the Colorado) has created and maintained nearly 10,000 ha of marsh 

habitat for these birds, sustaining the largest population of these species in Northwestern 

Mexico and the Sonoran Desert ecoregion (Hinojosa-Huerta et al. 2008).

As a product of the Cerro Prieto Geothermal facility, several ponds were created. 

These have small islets on which at least 11 species of waterbirds nest: great blue heron, 

great egret, snowy egret, black-crowned night-heron, laughing gull, gull-billed, Caspian 

and Forster’s terns, black skimmer, snowy plover, American avocet and, possibly, blacknecked stilt (Mellink et al. 2002; Molina and Garret 2001). The nearby 100 ha wetland 

of Las Arenitas, associated with the wastewater treatment plant, also provides habitat 

for several breeding waterbirds, including redhead, Yuma Ridgeway rail, Virginia rail, 

American coot, common gallinule, and eared grebe.

Upper Estuary

The tides of the Upper Gulf have one of the largest amplitudes in the world and 

extend 44 km upstream from the river mouth during the highest tides of the year, connecting with the flows of the Hardy River and other agricultural drains (Nelson et al. 

2013). This area, known as the Upper Estuary, covers nearly 12,000 ha and is dominated 

by mudflats, tidal channels, Palmer’s salt grass (Distichlis palmeri) patches, saltcedar 

(Tamarix spp.) trees, and few patches of emergent vegetation. The mudflats have been 

traditionally used for nesting by at least five species: American avocet, black-necked 

stilt, snowy plover, killdeer, and least tern (Mellink et al. 1996).

This area has experienced large-scale changes recently. Most of it was above the 

highest tidal line, but coseismic subsidence caused by the April 4, 2010, earthquake 

dropped the elevation of the area by up to 1.5 m (Nelson et al. 2013), creating this large 
USDA Forest Service RMRS-GTR-377. 2018 179

area of mudflats. The decrease in elevation also facilitated the interaction between tides 

and drains, especially in sites like Laguna del Indio, where the surface of open water 

increased. These recently formed habitats are being used by a diversity of waterbirds, 

in particular migratory shorebirds. Prior to the 2010 earthquake, the area had very few 

birds (Hinojosa-Huerta et al. 2004), and recent aerial counts estimate between 13,000 

and 20,000 shorebirds in the area (Gomez-Sapiens et al. 2013).

The increase of open water and flooding of pre-existing saltcedar thickets favored 

the establishment of new colonies of waterbirds, especially at Laguna del Indio, where 

a colony started forming in 2012, and, in 2014, had nesting and juveniles of great blue 

heron (21 nests), black-crowned night-heron (12 nests), snowy egret (9 nests), great 

egret (3 nests), green heron (2 nests), and double-crested cormorant (7 nests) (HinojosaHuerta et al. 2015). We have also detected vocalizing pairs of Yuma Ridgeway rail, 

Virginia rail, and least bittern during the breeding season.

Montague Island

Montague Island (22 km long and 7 km in maximum width, surface area = 

47 km2), is a low lying, silt island at the mouth of the Colorado River (31° 43’ 45” N 

and 114° 45’ 05” W). The island is nearly featureless, but in its southeastern end there 

are a series of tidal channels, with banks supporting Palmer’s salt grass, the only vascular vegetation of the island, and several sub-fossil shell banks of a nearly-extinct clam 

(Mulinia coloradoensis; Karl Flessa, University of Arizona, personal communication).

Montague Island existed early in the 20th century, but its existence before the 19th 

century has not been ascertained. It may have been formed by increased sediment supplies produced by the removal of beavers and by cattle grazing in the Colorado River 

watershed. Its birds, as well as its other biota, have been studied scantly. Of all known 

early visitors to the Colorado River Delta, only L.J. Goldman is known to have visited 

Montague Island. He failed to find any nesting birds there, but considered that “... some 

of the smaller species might have had nests in the grass,” possibly referring to the largebilled savannah sparrows (Passerculus sandwichensis rostratus), which he found to be 

abundant on the island (Smithsonian Institution Archives, Record Unit 7176).

Montague Island remained marginal to biological reconnaissance for nearly eight 

decades, until 1991, when we began limited and intermittent research work (Mellink 

2003; Mellink et al. 2002; Palacios and Mellink 1992, 1993; Peresbarbosa-Rojas and 

Mellink 1994, 2001). At least 12 species of waterbirds have been documented to nest on 

Montague Island (table 10). Additionally, one landbird, the large-billed savannah sparrow, also nests there. The data suggest that Montague Island has been used by nesting 

birds for no more than a century, as we elaborate below.

At the turn of the 20th century the Colorado River carried large, albeit variable, volumes of water all the way to the Gulf of California. At this time, even under 

conditions of an “unusually low river,” as in 1902, Montague and Gore Islands were 

“frequently submerged at periods of high tide and river ... [and] at high tide a semisubmerged fringe of salt grass is all that is visible” (Sykes 1937). Similar conditions 

were recorded in 1915 (L.J. Goldman in 1915, Smithsonian Institution Archives, Record 

Unit 7176), also a relatively dry year (U.S. Claims Court 1988). Conditions in “normal” 
180 USDA Forest Service RMRS-GTR-377. 2018

or wet years could only be less suitable. Hence, no successful nesting seems to have 

been possible at that time.

By 1935, water storing and diversion projects, including Boulder Dam, had been 

profusely established on the Colorado River and caused a reduction in water flow in the 

lower delta (Sykes 1937; U.S. Claims Court 1988). In addition to several biological impacts (Mellink and Ferreira-Bartrina 2000), the reduction in flow might have allowed for 

extended periods of dry ground and the development of channels with vegetated banks 

on Montague Island, making it suitable for nesting by waterbirds for the first time.

The ardeids nested widely in the region and would have easily colonized 

Montague Island when nesting habitat became available. The larids present a more 

intriguing question. On his 1915 trip to Montague Island, Goldman collected 12 specimens (held in the National Museum of Natural History), among them two gull-billed 

terns (James Dean, National Museum of Natural History, personal communication). 

This was the first documentation of this species in the delta of the Colorado River, 

although it was overlooked by both Nelson (1921) and Grinnell (1928). In the Salton 

Sea, there were at least 500 pairs of gull-billed terns nesting by 1927 (Pemberton 

1927). For such a colony to have formed since the creation of the Sea, or in 6 years 

as Pemberton thought, the founding group must have been large, or gull-billed terns 

must have been already nesting somewhere in the delta. The two specimens during the 

breeding season from Montague Island before conditions were adequate suggest this 

later possibility, in which case gull-billed terns might have nested in some southern, 

but unsurveyed, sections of the delta.

Not only were laughing gulls and black skimmers not nesting on Montague Island 

at the time, but they had not been recorded in the delta in the early 20th century at all 

(Dawson 1923; Grinnell 1928; Murphy 1917). Laughing gulls began to nest in the 

Salton Sea in the late 1920s (Miller and van Rossem 1929). This species is highly visible, and individuals can feed several kilometers from their breeding sites. They would 

not have been overlooked by ornithologists visiting the delta if they had nested in the 

area. This species is now found in the area even in the non-breeding season (Patten et 

al. 2001). Its scarcity in the Salton Sea in 1928 (Miller and van Rossem 1929) suggest 

that it might have been a newcomer to the area, perhaps from the southern Sonoranorthern Sinaloa colonies.

Black skimmers are seemingly recent arrivals to the delta, as they were recorded 

for the first time in the Salton Sea in 1968 (Collins and Garrett 1996; McCaskie et al. 

1974). Given the species’ post-Salton Sea history, it is possible that black skimmers 

used Montague Island as a stepping stone to reach it (Peresbarbosa-Rojas and Mellink 

2001), as is suggested by banding records. A likely source for the Montague Island 

colony would have been the southern Sonora-northern Sinaloa wetlands.

Wilson’s plovers were detected to nest there only recently (Eduardo Soto, the 

head of monitoring, Upper Gulf of California and Delta del Río Colorado Biosphere 

Reserve, personal communication). These would have colonized from nearby colonies 

in Sonora, where they nest along the entire northern coast (Mellink and Palacios 1993).

One of the most striking characteristics of Montague Island is frequent tidal inundations (Peresbarbosa-Rojas and Mellink 2001). Such inundations result in very low 

reproductive output of birds in many, or perhaps most, years. However, in some years 

there is a good crop of chicks. For example, on 28 June 2001, there were abundant 
USDA Forest Service RMRS-GTR-377. 2018 181

chicks of snowy egret and laughing gull that were old enough to prove that the colonies 

had not been completely inundated, and giving the prospect of a season with many 

fledglings of these two species for that year. These inundations also cause within-year 

dynamics as colonies totally or partially move between the Estero and the Lighthouse 

or from lower tidal channels of the Estero to higher ones (Peresbarbosa-Rojas and 

Mellink 2001).

There are also inter-year differences, as exhibited during 1993 and 1994. In these 

two non-El Niño years there were some differences, at least in the number of snowy 

egrets, laughing gulls, and least terns (table 11). Not only were there fewer nests of 

snowy egrets and laughing gulls in 1994, but these species did not begin nesting until 

June, much later than in 1993 (Peresbarbosa-Rojas and Mellink 2001). During 1993 

there was much runoff of fresh water through the Colorado River, and although difference in flow was the only evident difference, it was not clear how it could cause the 

observed differences in nest numbers.

At a larger scale, during El Niño (Southern Oscillation) years the colonies on 

Montague Island exhibit remarkable changes (table 11; Mellink 2003). In 1992 and 

1998, both El Niño years, there was an influx of nesting elegant and royal terns, and in 

1992 there were many more gull-billed tern nests than in 1993 and, in 1998, many more 

least tern and black skimmer nests. Such increases in Montague Island bird populations 

do likely not reflect a major increase in available food or favorable breeding conditions. 

Rather, it might reflect poor conditions at other breeding grounds (Mellink 2003), at 

least for those species that nest far away and depend on marine resources. However, our 

knowledge is inadequate to speculate on the particular origin of individuals reaching 

Montague Island.

Table 11—Number of nesting pairs of waterbirds recorded on Montague Island. El Niño conditions prevailed during 

the breeding seasons of 1992 and 1998.

Species 1991 1992 1993 1994 1998 1999 2001

Great blue heron ? - 1 2 - - 1?

Snowy egret >100 + 87 23 129 + +

Black-crowned night-heron 20? 2? - 10 - 5 -

American oystercatcher n.r. - ? - 1 2 -

Laughing gull 60 + 170 127 102 + +

Caspian tern - - - def. - - -

Least tern n.r. 20 <10 27 ±110 8 -

Gull-billed tern - ±175 92 94 >77 + few

Elegant tern n.r ±275 - - 160 2 1

Royal tern n.r. ±275 - - 135 16 12

Black skimmer - ? 14 21 ±178 20 low 100s

? = possibly nesting

- = not found nesting

n.r. = not reviewed

def. = nest defense behavior, but nest not searched for or not found; 

+ = the species was nesting in numbers comparable to previous years 

few = that there were notably fewer pairs nesting than in previous years.
182 USDA Forest Service RMRS-GTR-377. 2018

Conservation Problems

As in other delta ecosystems worldwide, insufficient and unreliable water supplies and poor water quality are the ultimate causes of environmental degradation in 

the Colorado River Delta, intensified in this particular case by its location in an arid 

region. The remnant wetlands in the delta have survived with accidental releases or unintentional flows, and until very recently, there was no water secured for environmental 

purposes. This is a major challenge for all the different wetland areas, including the 

Ciénega de Santa Clara, that rely almost completely on agricultural return flows coming 

from the United States. Such flows have been planned to be diverted for desalinization 

and further consumptive uses, which will likely cause the degradation of these wetlands.

Most of the present delta vegetation has been enhanced and maintained by 

flood releases during the 1980s and 1990s (Glenn et al. 2001), restoring in part the 

cottonwood-willow forests. However, lack of a proper flow regime increases the depth 

to groundwater, reduces the recruitment of native trees, and increases soil salinity, which 

favors the establishment of exotic plants, and thereby decreases habitat value. This is 

coupled with a generalized increase of groundwater extractions, which causes drastic 

declines of the water table in certain sections of the river corridor, drying up remaining 

wetlands and stressing riparian vegetation (Ramirez et al. 2013).

Pollution is a concern in the delta that has been prevalent for several decades (see 

García-Hernández et al. 2000, 2013; Mora 1991; Mora and Anderson 1995). The farming areas in Baja California, Sonora, and Arizona that drain toward the delta are under 

intense, highly mechanized cultivation. This type of farming involves large quantities of 

agrochemicals, some of which drain into the wetlands of the delta and flow to the sea in 

the areas with tidal exchange (Daesslé et al. 2008). Major pollutants of concern in these 

wetlands and their biota include metals such as mercury, copper, arsenic, and selenium, 

as well as residue of DDT, PCBs and other organic agrochemicals, especially in the 

Ciénega de Santa Clara, Hardy River, and the Upper Estuary (García-Hernández et al. 

2006, 2013; Guardado-Puentes 1975; Mora 1991).

Human disturbance to the colonies has not been a problem yet. However, as extractive activities become more restricted due to the protected status of the area, larger 

numbers of tourists, especially at the seashore, may have negative impacts on bird communities. Indeed, some fishermen already have taken tourists to Montague Island during 

the breeding season. If this activity increases without controls to prevent disturbance, 

the impacts might aggravate natural impacts (inundating tides, for example), and eventually discourage birds from nesting here.

Predation can impact waterbird colonies (e.g., Pratt and Winkler 1985) but has 

not been documented on the lower Colorado Delta in general. However, on Montague 

Island nesting waterbirds are subject to predation by coyotes, which sometimes take a 

heavy toll on the colonies (Peresbarbosa-Rojas and Mellink 2001). It is not possible to 

determine the real effect of such predation, especially in the presence of other events 

that destroy eggs such as high tides or possibly ants (Peresbarbosa-Rojas and Mellink 

2001). However, prevention of coyote predation might be of benefit and would be easily 

accomplished, if desired.

Two other potential threats, indicated for seabirds in Baja California (Everett and 

Anderson, 1991), seem not to be a problem for Montague Island nesting waterbirds: 
USDA Forest Service RMRS-GTR-377. 2018 183

(1) human exploitation of birds or eggs, and (2) impacts of fisheries. In the delta, fisheries have been greatly restricted, involve artisanal gill-nets that do not entangle birds, and 

extract only fish that are much larger than those consumed by the nesting birds.

Restoration and Conservation Actions

One of the key strategies to restore waterbird habitat has included securing water 

sources for the different wetland areas in the delta. This has involved securing the effluent of wastewater plants as instream flows, maintaining the agricultural return flows to 

the wetlands, and the dedication of irrigation water rights to restore key areas (HinojosaHuerta and Carrillo-Guerrero 2010; Zamora-Arroyo et al. 2008). The Colorado River 

corridor has recently received allocated flows for ecological restoration under Minute 

319 of the International Water Treaty, securing 195 million cubic meters of water on a 

5-year period (2012–2017), with one-third of the water provided by Mexico, one-third 

by the United States, and one-third by a binational coalition of environmental organizations (Gerlak 2015). Part of this water (130 million cubic meters) was delivered as a 

“pulse flow” during 2014, to revitalize the delta ecosystem, to promote the greening-up 

of the existing vegetation and the germination of new native plants, and to learn more 

about the hydrological and biological response of the system (Flessa et al. 2013). Since 

then, and until 2017, the rest of the water (65 million cubic meters) is being delivered 

as “base flows” at critical areas of the corridor, to maintain the habitat that has been 

restored (Flessa et al. 2013).

These efforts have been coupled with intensive restoration actions in specific 

locations along the riparian corridor, where the intervention includes clearing exotic 

vegetation, grading the land, establishing native plants, and adding water to sustain the 

new vegetation and create open water zones (Hinojosa-Huerta et al. 2005). Between 

2008 and 2014, a total of 337 ha has been restored and maintained.

Part of the efforts in the delta also includes the implementation of land conservation 

strategies. One of the earliest achievements was the establishment of the Upper Gulf of 

California and Colorado River Delta Biosphere Reserve, which protects the Ciénega de 

Santa Clara, El Doctor Wetlands, Laguna del Indio, and the mudflats near the mouth of 

the river. To complement the public conservation efforts implemented by the Mexican 

government with the Biosphere Reserve, environmental organizations have established 

conservation easement agreements with landowners in key wetlands (Ciénega de Santa 

Clara and El Doctor) to protect priority habitat for endangered and migratory species. A 

total of 1,200 ha has been protected under this mechanism in the delta.

Concluding Remarks

Amazingly, despite the intensive and extensive modification of the wetlands of the 

Mexican portion of the Colorado River Delta, 37 species of waterbirds nest here. Most 

of those that nested here in the early 20th century, before the large-scale environmental 

changes, do so still today, albeit perhaps in much lower numbers. Only one species has 

ceased breeding in the delta (fulvous whistling-duck), but several have been gained, 

mostly thanks to the creation of new habitat, including Montague Island (as a nesting 

locale), the Cerro Prieto islets and, on the U.S. side but interconnected, the Salton Sea.
184 USDA Forest Service RMRS-GTR-377. 2018

At the same time, the rise of public and government environmental awareness, 

public pressure, and careful negotiation by NGOs have opened unprecedented opportunities for the conservation of wetlands and their associated biota in this area. Fulfillment 

of such opportunities with serious, knowledgeable, science-based wetland restoration 

and creation will increase the role of the Mexican portion of the Colorado River Delta as 

an important area for the conservation of regional biodiversity.

Acknowledgments

We are very thankful to Edith Santiago and Eduardo Soto-Montoya for kindly 

sharing unpublished information with us, and Karl Flessa and James Dean, for providing important information.

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Chapter 11. Terrestrial Vertebrates of Mesquite Bosques 

in Southwestern North America

R. Roy Johnson, Elaine E. Johnson, and Steven W. Carothers

Introduction

The major emphasis of this chapter is to address the species richness and population densities of land vertebrates in riparian mesquite bosques (woodlands). We find no 

single publication that lists vertebrates—amphibians, reptiles, birds, and mammals—of 

riparian mesquites of the Southwest lowlands. These vertebrates are listed for a few 

river valleys, such as the Santa Cruz River (Webb et al. 2014) and San Pedro River 

(Stromberg and Tellman 2009), but not for the Southwest as a whole. This chapter presents a summary of the vertebrate fauna of mesquite bosques. Our lists of vertebrates of 

bosques and/or associated cottonwood-willow forests and riparian deciduous woodlands 

have been constructed from literature, historic records and specimens, and our first-hand 

knowledge.

Mesquite Bottomlands as Wildlife Habitat

A large number of the vertebrate species occurring as obligate and facultative 

riparian species in cottonwood-willow gallery forests also occur in mesquite bosques. 

These species inhabit riparian mesquites not only along streams but also along dry 

watercourses (tables 12–15). An amphibian or reptile species may often occur in cottonwood-willow or mixed deciduous forests, in mesquite bosques, or in upland ecosystems. 

A mammal species—except larger mammals and bats—may also often occur in one 

of these three watershed zones. Although an avian species may use one of those three 

zones, several species occur in all three zones. Some birds, especially larger species, 

roost and nest at distances from riparian ecosystems but spend parts of the day flying 

over and/or foraging in bosques and adjacent zones.

Some vertebrates are so ubiquitous in bosques that two of them are often called the 

“mesquite warbler” (Lucy’s Warbler; Gilman 1909; Johnson et al. 1997) and “mesquite 

mouse” (Merriam’s mouse; Hoffmeister 1986; IUCN 2015; see tables for scientific 

names). Several tropical species whose ranges extend no farther north than the extreme 

southwestern United States are now relatively rare in this region, having experienced 

drastic reductions in their U.S. populations, paralleling the losses of most mesquite 

bosques. Notable are the ferruginous pygmy-owl (Johnson et al. 2003; Phillips et al. 

1964), gray hawk (Bibles et al. 2002; Glinski 1998), and brown vinesnake (Stebbins 

1985). Although the species we list (tables 10 to 13) are mainly associated with perennial and larger intermittent streams, some species also maintain populations within 

xeroriparian ecosystems associated with desert washes.

Birds of mesquite bosques and xeroriparian ecosystems are the best studied of vertebrates. Recent surveys of Sonoran and Chihuahuan desert ecosystems have compared 

breeding birds of xeroriparian ecosystems along desert washes—most of them lined 
USDA Forest Service RMRS-GTR-377. 2018 189

with mesquites—with the surrounding uplands. Numbers of species and/or numbers of 

individual birds of these xeroriparian ecosystems are commonly 10 times or more compared to those of the surrounding uplands (Johnson and Haight 1985, 1988). Some of 

the larger drainages, such as Aguajita, Growler, and Vekol washes in southern Arizona, 

have small bosques with large mesquite trees that support nesting by numerous species of smaller birds and even larger birds, such as Cooper’s hawks, red-tailed hawks, 

and great horned owls (Groschupf et al. 1988) and rare species such as the ferruginous 

pygmy-owl (Johnson et al. 2003). This small tropical owl was first discovered for the 

United States in a mesquite bosque along the Rillito River in 1872 by Major Charles 

Bendire (Bendire 1892; Proudfoot and Johnson 2000). In addition, large mammals often 

rely on xeroriparian habitat along these washes for food, shelter, and movement corridors, including species such as mule deer (Krausman et al. 1985), javelina, and coyotes 

(R.R. Johnson, numerous observations, Tucson, Arizona, during several decades).

Of particular interest is a group of amphibians in Vekol Valley in southwestern Arizona, just north of the United States-Mexico international boundary. The arid 

Vekol Valley is populated by three amphibian species that occur no farther north than 

southwestern Arizona—Sonoran green toad, lowland burrowing treefrog, and Sinaloan 

narrow-mouthed toad (Enderson and Bezy 2007; R. Babb, Biologist, Arizona Game and 

Fish Department, personal communication). These three tropical amphibians are among 

species that occur in mesquite grasslands of this region (Stebbins 1985). The vegetation 

of Vekol Valley is largely Sonoran desertscrub (Brown 1982) but has remains of more 

extensive semidesert grasslands that existed prior to heavy grazing by cattle. The valley is also transected by the large, mesquite-lined Vekol Wash that contains numerous 

earthen cattle tanks. Amphibians are generally considered inhabitants of at least moist 

conditions, often perennial water, but these three, as well as several other species, have 

adapted to more arid conditions. Like several other species of frogs and toads, these 

three amphibians aestivate during drier and colder parts of the year, emerging during 

summer rains to breed and lay eggs in resulting pools of water.

The Mesquite-Invertebrate-Vertebrate Foodweb

It is not surprising to find that avian and mammalian faunas of a region show a 

preference for mesquites and other riparian habitats because of access to food, shelter, 

and for some—water—associated with riparian ecosystems (see tables 13, 14, and 

15). Most insects are herbivorous; thus insect populations are elevated in the increased 

vegetation in riparian ecosystems, compared to upland ecosystems of the Southwest 

lowlands. More than 700 species of insects associated with mesquites in the new 

world were reported by Ward et al. (1977). This is reflected in the large percentage of 

insectivorous riparian birds found in mesquite bosques and cottonwood-willow riparian 

forests (tables 11 and 12).

Lizards are also largely insectivorous (Jones and Lovich 2009; Stebbins 1985) 

and snakes that prey on lizards and mammals (Stebbins 1985) also occur along with 

their prey in riparian ecosystems. Some snakes, such as gartersnakes, are aquatic 

or semi-aquatic and often inhabit hydroriparian or mesoriparian ecosystems along 

perennial or intermittent streams, feeding on frogs, toads, tadpoles, and other aquatic 

organisms (Stebbins 1985). Amphibians also commonly occur in riparian habitats 
190 USDA Forest Service RMRS-GTR-377. 2018 Table 12—Amphibians and reptiles (herpetofauna) occurring in lowland riparian ecosystems of the Southwest.a Common name Scientific name Santa Cruzb R. near Tucson San Pedro R.c LRd MBe CWf MEg MRh SDi Sonoran tiger salamanderj Ambystoma mavortium stebbinsi X L,S Couch’s spadefoot [toad] Scaphiopus couchii A X X X X X New Mexico (Chihuahuan Desert) spadefoot [toad] Spea [Scaphiopus] multiplicata X X Sonoran Desert (Colorado River) toad Ollotis alvaria A X X X X X Southwestern Woodhouse’s toad Anaxyrus woodhousei australis A X X X X Southwestern toad A. microscaphus X X X Red-spotted toad A. punctatus X X X X X X X Great Plains toad A. cognatus X X X Green toad A. debilis X X Texas toad A. speciosus B&B Sonoran green toad A. retiformis E&B Canyon treefrog Hyla arenicolor X X X X Lowland burrowing treefrog Smylisca fodiens E&B Tarahumara frog Lithobates [Rana] tarahumarae L, Ex Lowland leopard frogk L. yavapaiensis A, Exl X ??Leopard frog??m L.?? pipiens?? ? X X X X X American bullfrogn L. catesbeianus X X X Sinaloan narrow-mouthed toado Gastrophryne mazatlanensis X E&B Amphibians 18 AMPHIBIA 9 11+ 4 6 8 6 6 Yellow mud turtle Kinosternon flavescens L,S Sonoran mud turtle K. sonoriense XEp X X X X Spiny softshell (turtle) Trionyx spiniferus (introduced) Apalone spinifera X X Pond slider Pseudemys scripta Trachemys scripta (introduced) X X Desert box turtle Terrapene ornata A X Morafka’s desert tortoise *Can Gopherus [agassizii] morafkai A X X X X X Turtles 6 TESTUDINES 3 5 1 2 4 1 1 Tucson banded gecko Coleonyx variegatus A X X X X X Desert iguana Dipsosaurus dorsalis X X 
USDA Forest Service RMRS-GTR-377. 2018 191 Table 12—Continued.

2

Common name Scientific name 

Santa Cruzb

R. near Tucson 

San Pedro R.c

LRd MBe CWf MEg MRh SDi

Common chuckwalla Sauromalus ater X X X Common lesser earless lizardq Holbrookia maculata X X LrGreater earless lizard Cophosaurus texanus X J+ X X X X X Eastern zebra-tailed Lizard Callisaurus draconoides A X X X X X Common collared lizard Crotophytus collaris X X X X Long-nosed leopard lizard Gambelia wislizenii AEs X X X X X Desert spiny lizard Sceloporus magister A X X X X X X Twin-spotted spiny lizard S. bimaculosus J&LtSonoran spiny lizard S. clarkii A X X X X X Southwestern fence lizard S. undulatus cowlesi 

S. cowlesi 

XuEv X J+ X X 

Western side-blotched lizard Uta stansburiana A X X X X X Long-tailed brush lizard Urosaurus graciosus X X X Ornate tree lizard U. ornatus A X X X X X X Desert horned lizard Phrynosoma (Anota) platyrhinos J&Lw J X X Regal horned lizard P. (A.) solare A X X X Great plains skink Plestiodon obsoletus X X Many-lined skink P. multivirgatus J,J&L, 

Sx

Gilbert’s skink P. gilberti J,J&L, 

Sy

 J 

Madrean alligator lizard Elgaria kingiiz J&L, 

Saa

Giant spotted whiptailbb Aspidoscelis strictogramma X X J+ 

J&Lcc

New Mexican whiptail A. neomexicana J&L,SddLittle striped whiptail A. inornata B&BeeDesert grassland whiptail A. uniparens X J&LffPlateau striped whiptail A. velox J+ X X Chihuahuan spotted whiptail A. exsanquis J&L,LggSonoran spotted whiptailhh A. sonoraeii X X J&LjjGila spotted whiptail A. flagellicaudatus J+ X X Marbled whiptail A. marmorata J&LkkTiger whiptail A. tigris X X X X X X X Checkered whiptail A. tesselata J&LllGray checkered whiptail A dixoni B&Bmm
192 USDA Forest Service RMRS-GTR-377. 2018 Table 12—Continued.

3

Common name Scientific name 

Santa Cruzb

R. near Tucson 

San Pedro R.c

LRd MBe CWf MEg MRh SDi

Reticulate Gila monster Heloderma suspectum A X X X X X Lizards 34 REPTILIA: LARCERTILIA 14 17 9 18 10 16 16 Southwestern threadsnake (blind snake) 

Rena humilis A X X X X X X 

New Mexico threadsnake R. dissectus X SnnRosy boa Lichanura trivirgata X X Ring-necked snake Diadophus punctatus A X J X X Mexican hog-nosed snake Heterodon kennerlyi X Soo X Spotted leaf-nosed snake Phyllorhynchus decurtatus A LppSaddled (Pima) leaf-nosed snake P. browni A X X X Red racer (coachwhip) Coluber flagellum A X X X X X Striped whipsnake C. taeniatus J+ X X Sonoran whipsnake C. bilineatus A X J+ X X Western patch-nosed snake Salvadora hexalepis A X X X X X Trans-Pecos rat snake Bogertophis subocularis B&BqqGreat Plaines rat snake Pantherophis emoryi B&BrrArizona glossy snake Arizona elegans A X X X Sonoran gopher snake Pituophis catenifer affinis A X X X X X X Desert kingsnake Lampropeltis getula, L. splendida A X X X X X X New Mexican milk snake L. triangulum B&BssWestern long-nosed snake Rhinocheilus lecontei A X X X X X X New Mexico gartersnake Thamnophis sirtalis dorsalis B&B,S Black-necked gartersnake T. cyrtopsis X X X X X X (Northern) Mexican [brown] gartersnake *Can 

T. eques Ett X Luu

Marcy’s checkered gartersnake T. marcianus AvvEww X Lxx X Ground snake Sonora semiannulata X X X Western shovel-nosed snake Chionactis occipitalis X X X Chihuahuan hook-nosed snake Gyalopion canum IUCNyySmith’s black-headed snake Tantilla hobartsmithi X X J+ X X X X Brown vine snakezz Oxybelis aeneus X SaaaLyre snake Trimorphodon biscutatus X X X X Banded sandsnake C. stramineus A X X X X X Sonoran nightsnake Hypsiglena torquata 

Hypsiglena chlorophaea

A X X X X X X 
USDA Forest Service RMRS-GTR-377. 2018 193 Table 12—Continued.

4

Common name Scientific name 

Santa Cruzb

R. near Tucson 

San Pedro R.c

LRd MBe CWf MEg MRh SDi

Sonoran coral snake Micruroides euryxanthus Xbbb X X X Western diamond-backed rattlesnake 

Crotalus atrox A X X X X X 

Rock rattlesnake C. lepidus S, 

B&Bccc

Speckled rattlesnake C. mitchellii X X X Sidewinder C. cerastes X X X Black-tailed rattlesnake C. molossus X X X X Tiger rattlesnake C. tigris TCB

ddd

Northern Mojave rattlesnake C. scutulatus A X X X X Snakes 38 REPTILIA: SERPENTES 19 22 10 20 17 18 24 TOTAL 96 ALL HERPS 45 55+ 24 46 39 41 47 CW = Cottonwood-Willow, LR = Lowland Riparian, ME = Mesquite Bosque, MB = Mixed Broadleaf, MR = Mixed Riparian scrub (xeroriparian), SD = Sonoran 

Desertscrub; abbreviations—except LR—after Jones (1988). 

A = Arnold (1940), B&B = Bartlett and Bartlett (2010), E&B = Enderson and Bezy (2007), J-Jones (1981), J+-Jones et al. (1985), J&L = Jones and Lovich (2009), 

Lowe (1964b), Stebbins (1985) 

Ex = Extirpated *Can = candidate species for Federally Threatened or Endangered 

a Nomenclature after SSAR (2012) and Jones and Lovich (2009); systematic order after Stebbins (1985); species may be obligate, preferential, or 

 facultative riparian. 

b After Webb et al. (2014). c After Rosen (2009). d Lowland riparian records other than Arnold (1940), Jones (1988), Rosen (2009), or Webb et al. (2014). e Mixed Broadleaf (trees). f Cottonwood-Willow (trees). g Mesquite bosque (woodland). h Xeroriparian (Johnson et al. 1984; mixed riparian scrub along usually dry washes). i Sonoran Desertscrub (Arizona Upland; Brown 1982). 
194 USDA Forest Service RMRS-GTR-377. 2018 Table 12—Continued.

5

j Throughout much of the SW lowlands (Stebbins 1985), often introduced accidentally as fishing bait (Lowe 1964b); formerly occurred in San Pedro River (Rosen 

2009); an endangered species (USFWS) listed in 1996 by Arizona Game and Fish Dept. in “Wildlife of Special Concern.” k Special Status Species, Arizona Game and Fish Department. 

l Extinct by 1975 at latest (Rosen 2001, 2003). m Rana pipens has been split into several species and it is uncertain which of those the older records represent. n Introduced species from the eastern Unites States (Stebbins 1985). o Although ITIS (no date) still uses Gastrophryne olivacea (Great Plains Narrow-mouthed Toad) our information suggests that G. mazatlanensis extends northward 

into extreme SW Arizona from Mexico (see text). p Widely inhabited perennial water along the Santa Cruz; extirpated (Rosen 2001). q Found on a “Tucson bajada adjoining West Branch floodplain” of Santa Cruz River (Rosen 2003:40). r “Bottomlands” in grasslands (Lowe 1964b:160) among other habitats. s Extirpated in Tucson area by 1950s or 1960s (Rosen 2008). 

t “In dry canyons and arroyos or along rivers and streams” (Jones and Lovich (2009:204) among other habitats. u “Downtown Tucson, 1995” (Rosen 2003:40). v Extirpated in Tucson area (Rosen 2008). w “Alluvial fans, washes” (Jones and Lovich (2009:192) among other habitats. x “Dense streamside growth” (Stebbins 1985) among other habitats. y “Riparian corridors” (Jones and Lovich 2009:450), “cottonwood, willow and mesquite bottomlands” (Stebbins 1985:150) among other habitats. z Formerly Gerrhonotus kingii. aa Along streams, such as “broad leaf stream-border habitats along major drainageways in desert and grassland” (Stebbins 1985:166) among other habitats. bb Formerly A. burti stictogramma now considered a separate species from A. burti (SSAR 2012). cc “Preferred thickets include Velvet Mesquite . . . Riparian woodland is also used” (Jones and Lovich 2009:332). dd A unisexual species in “floodplain habitats” (Stebbins 1985:154), and the “Rio Grande corridor” (Jones and Lovich 2009:367) among other habitats. ee “Along dry washes” (Bartlett and Bartlett 2013:200) among other habitats. ff “Common along lower-elevation riparian corridors” (Jones and Lovich 2009:399) among other habitats. gg A unisexual species in “riparian corridors” (Jones and Lovich 2009:339), “riparian situations” (Lowe 1964b:165). hh A triploid unisexual species; females reproduce parthenogenetically (Jones and Lovich 2009). ii Rosen (2003) discusses these parthenogenetically reproducing, single sex species and uses the Tiger Whiptail because of ongoing research into this 

complicated group of lizards. 

jj “Especially abundant along riparian corridors” (Jones and Lovich 2009:387). kk Chihuahuan Desert . . . [in] Desert willow, mesquites” (Jones and Lovich 2009:363) among other habitats. ll A unisexual species in “riparian corridors” (Jones and Lovich 2009:391) among other habitats. mmA unisexual species in “Dry washes, canyon bottoms” (Bartlett and Bartlett 2013:195) among other habitats. nn “Often frequents . . . canyon bottoms or washes near permanent or intermittent streams” (Stebbins 1985:172). oo “Occasionally in mountain canyon bottoms or on the floodplains of streams (Stebbins 1985:176-177). pp “Fossorial in sandy soil” (Lowe 1964b:170). qq “Often . . . along intermittent streams, or in riparian situations” (Bartlett and Bartlett 2013:109). rr “Within its range there is usually a spring, stock tank, stream, pond, or lake” (Bartlett and Bartlett 2013:111). ss “Found in . . . riparian areas” (Bartlett and Bartlett 2013:119) among others. tt Northern Mexican Gartersnake known from Rillito River, possibly also along Santa Cruz River; extinct in Tucson Basin by1960s (Lowe 1985). 
USDA Forest Service RMRS-GTR-377. 2018 195 6 uu “In or near water along streams” (Lowe 1964b:167). vv Apparently a tentative identification by Rosen (2003). ww Reported from several locations on Santa Cruz floodplain; extirpated by 1970s (Lowe 1985, Webb et al. 2014). xx “Near to, rivers, streams and ponds in arid and semi-arid habitats” (Lowe 1964b:167). yy “Occurs in lower desert communities (e.g., mesquite . . .)” (IUCN 2015), “slightly sandy habitats, near a permanent water source” Wikipedia online. zz First discovered for the U.S. in mesquites along the Santa Cruz River south of Tucson (Vorhies 1926), “also among grasses and reeds at the sides of streams and ponds” (Lowe 1964:171); now restricted in the United States to the Nogales, AZ, area (Van Devender et al. 1994). aaa “Stream bottoms grown to sycamore, oak, walnut and wild grape” (Stebbins 1985:220) among other habitats. bbb “Most abundant . . . along arroyos and river bottoms. A secretive species, abroad chiefly at night” (Stebbins 1985:223). ccc “Streambeds . . . may occur in the lowlands” (Stebbins 1985:228), “near rocky intermittent streams and rivers” (Bartlett and Bartlett 2013:144) among other habitats. ddd “Usually on rocky slopes or in washes” (Brennan 2008) among other habitats, “on the lower slopes” (Stebbins 1985:231). Table 12—Continued.
196 USDA Forest Service RMRS-GTR-377. 2018

Table 13—Riparian breeding birdsa of mesquite bosques and mixed deciduous woodlands.

Mesquite Mixed deciduous

Location Lower 

Verde R.

San 

Pedro R.

San 

Pedro R.

Dry 

Beaver 

Crk.b

West 

Clear 

Crk.c

Years of records 1975 1972-

1973

2006-

2015

1969-

1974

1969-

1974

Source NES 

(1978)

TAG

(1975)

IBA

(2015)d SWC(1974)

SWC

(1974)

Common name Scientific namee

#Gambel's quail Callipepla gambelii X XO X

Great blue heron Ardea herodias X

Turkey vulture Cathartes aura F X

#Cooper's hawk Accipiter cooperii ?? X

#Gray hawk Buteo plagiatus X

#Swainson's hawk Buteo swainsoni F X

Red-tailed hawk B. jamaicensis X

#White-winged dove Zenaida asiatica X XO X

#Mourning dove Z. macroura X X X X X

Inca dove Columbina inca ??

#Common ground-dove C. passerina X X

#Yellow-billed cuckoof Coccyzus americanus X

#Greater roadrunner Geococcyx californianus XO X

Barn owl Tyto alba ??

#Western screech-owl Megascops kennicottii ??

Great horned owl Bubo virginianus ?? ??

#Elf owl Micrathene whitneyi ??

Lesser nighthawk Chordeiles acutipennis F X

#Black-chinned hummingbird Archilochus alexandri X ?? X X

Anna’s hummingbird Calypte anna X

Hummingbird sp.g ????????? XO

#Gila woodpecker Melanerpes uropygialis X XO X X X

#Ladder-backed woodpecker Picoides scalaris X X X X X

Northern flicker Colaptes auratus X

#Gilded flicker C. chrysoides X

#Northern beardlesstyrannuletCamptostoma imberbe ?? X

Western wood pewee Contopus sordidulus X X

#Willow flycatcher

[gray flycatcher]

Empidonax traillii 

[E. wrightii]

Xh

#Black phoebe Sayornis nigricans ?? X

Say's phoebe S. saya ??

#Vermilion flycatcher Pyrocephalus rubinus X X

#Ash-throated flycatcher Myiarchus cinerascens X X X X X

#Brown-crested flycatcher M. tyrannulus X X X X

Western kingbird T. verticalis XO X X

#Bell's vireo Vireo bellii X X X X

#Chihuahuan raven Corvus cryptoleucus F

Common raven C. corax F X

Violet-green swallow Tachycineta thalassina X

#Northern rough-winged 

swallow

Stelgidopteryx serripennis X

Bridled titmouse Baeolophus wollweberi X X

#Verdin Auriparus flaviceps X X X X X

#Cactus wren Campylorhynchus 

brunneicapillus 

??

#Bewick's wren Thryomanes bewickii X X X X

#Black-tailed gnatcatcher Polioptila melanura X

#Curve-billed thrasher T. curvirostre ?? ??

#Crissal thrasher T. crissale X

#Northern mockingbird Mimus polyglottos ?? X

European starlingi Sturnus vulgaris X

#Phainopepla Phainopepla nitens XO X X

#Lucy's warbler Oreothlypis luciae X X X X X

#Common yellowthroat Geothlypis trichas ??
USDA Forest Service RMRS-GTR-377. 2018 197

Table 13—Continued. 2

Mesquite Mixed deciduous

Location Lower 

Verde R.

San 

Pedro R.

San 

Pedro R.

Dry 

Beaver 

Crk.b

West 

Clear 

Crk.c

Years of records 1975 1972-

1973

2006-

2015

1969-

1974

1969-

1974

Source NES 

(1978)

TAG

(1975)

IBA

(2015)d SWC(1974)

SWC

(1974)

Common name Scientific namee

#Yellow warbler Setophaga petechia XO X X X

#Yellow-breasted chat Icteria virens X X X X

#Canyon towhee Melozone fusca X

#Abert's towhee M. aberti X X

#Rufous-winged sparrow Peucaea carpalis X

#Black-throated sparrow Amphispiza bilineata X X X

#Song sparrow Melospiza melodia X X

#Summer tanager Piranga rubra X X X X

#Northern cardinal Cardinalis cardinalis X X X X

#Pyrrhuloxia C. sinuatus 

#Blue grosbeak Passerina caerulea X X X

#Brown-headed cowbird M. ater X X X

#Hooded oriole Icterus cucullatus X X X

#Bullock's oriole I. bullockii X ??

#Lesser goldfinch Spinus psaltria XO X X X

#House finch Haemorhous mexicanus X XO X X X

Breeding species 19 30+ 43+ 19 27j

Pairs/40 ha (100 A) 244 277+k ??? 332 312l

# = Mesquites found in our studies or specifically mentioned under breeding habitat in literature, especially by Corman and WiseGervias. (2005).

X = Known or presumed to be nesting on the plot.

XO = Recorded during breeding census in small numbers, presumed nesting in the general area but possibly not on plot.

F = Flying over the area during the breeding season; often large birds foraging.

?? = Species breeding in the region in riparian habitat and recorded on the study plot during the summer but status uncertain, 

e.g., records early and/or late but no or few mid-summer records.

Sources: IBA = Important Bird Areas, Arizona and Tucson Audubon Society (2015); NES = Stamp (1978; SWC = Carothers et al. 

(1974); TAG = Gavin and Sowls (1975).

 

a Species may be obligate, preferential, or facultative riparian.

b After Carothers et al. 1974.

c After Carothers et al. 1974.

d Website—http://aziba.org/?page_id=30 (accessed 20 August 2015). e After AOU (1998) and AOU Check-list (2015). f

Yellow-billed cuckoo, a candidate for listing as a threatened or endangered species (USFWS online 2012, is the last breeding bird 

to arrive annually, generally in early June) (Phillips et al. 1964, Johnson and Simpson MS b). g Probably black-chinned hummingbird. h Misidentified in Carothers et al. (1974) as gray flycatcher, actually willow flycatcher. 

i

European starling introduced into United States from Europe, first reported for Arizona in 1946 (Phillips et al. 1964; Monson and 

Phillips 1981). 

j

Includes all species on two sites but number of pairs only for site with greatest number.

k Of the 36 species recorded during the breeding season, 16 were in numbers too small to be counted or numbers were not 

otherwise recorded (designated XO) thus are not included in the total. 

l

A second mixed deciduous site along West Clear Creek with only 193 pairs was highly disturbed, largely by a large U.S. Forest 

Service campground.
198 USDA Forest Service RMRS-GTR-377. 2018

Table 14—Riparian breeding birds of cottonwood-willow riparian gallery forests or cottonwood-willow in association with 

mesquitea. 

Mesquite

and c-w

Cottonwoodwillow (c-w)

 Location

GMF

Santa

Cruzb

BPC

Salt

R.c 

Verded

Valley

Lower 

Verde R.

Lower

Rio

Grande

Years of records

1902-

1963 1930s-1980s

1969-

1974 1975 1977

Source

RHW

(2014)

RRJ

(2000)

SWC

(1974)

NES 

(1978)

RE-W

(1979)

Common name Scientific namee

#Gambel’s quail Callipepla gambelii XO X X X

Pied-billed grebef Podylimbus podiceps XO X

Least bittern Ixobrychus exilis X

Great blue heron Ardea herodias XO X Xg

Green heronh Butorides virescens XO X

#Black-crowned night-heron Nycticorax nycticorax XO

Black vulture**i Coragyps atratus F

Turkey vulture Cathartes aura F F

Ospreyj Pandion haliaetus F

Bald eagle Haliaeetus leucocephalus X

#Cooper’s hawk Accipiter cooperii XO X X

#Common black hawk** Buteogallus anthracinus XO X X

#Harris’s hawk Parabuteo unicinctus XO X

#Gray hawk** Buteo plagiatus XO

#Swainson’s hawk Buteo swainsoni XO

Zone-tailed hawk** B. albonotatus XO

Red-tailed hawk B. jamaicensis XO X

Sora Porzana carolina X

Common moorhen Gallinula chloropus X

American coot Fulica americana X

Killdeer Charadrius vociferus XO X X

Spotted sandpiper Actitis macularia X

#White-winged dovek Zenaida asiatica XO X X X

#Mourning dove Z. macroura XO X X X X

Inca dove Columbina inca XO

#Common ground-dove C. passerina XO

#Yellow-billed cuckoo**l Coccyzus americanus XO X X X X

#Greater roadrunner Geococcyx californianus XO X X

Barn owl Tyto alba XO X

#Western screech-owl Megascops kennicottii XO X

Great horned owl Bubo virginianus XO X X

#Ferruginous pygmy-owl**m Glaucidium brasilianum X

#Elf owl Micrathene whitneyi XO X

Burrowing owln Athene cunicularia XO

Lesser nighthawk Chordeiles acutipennis XO X

#Common poorwill Phalaenoptilus nuttallii XO X
USDA Forest Service RMRS-GTR-377. 2018 199

Mesquite

and c-w

Cottonwoodwillow (c-w)

 Location

GMF

Santa

Cruzb

BPC

Salt

R.c 

Verded

Valley

Lower 

Verde R.

Lower

Rio

Grande

Years of records

1902-

1963 1930s-1980s

1969-

1974 1975 1977

Source

RHW

(2014)

RRJ

(2000)

SWC

(1974)

NES 

(1978)

RE-W

(1979)

Common name Scientific namee

#Black-chinned hummingbird Archilochus alexandri XO X X X X

Anna’s hummingbird Calypte anna X

#Costa’s hummingbird C. costae X

#Gila woodpecker Melanerpes uropygialis XO X X X

#Ladder-backed woodpecker Picoides scalaris XO X X X X

Northern flicker Colaptes auratus X

#Gilded flicker C. chrysoides XO X X

#Crested caracara Caracara cheriway XO

American kestrel Falco sparverius XO X X

#Northern beardless- tyrannulet Camptostoma imberbe XO

Western wood pewee Contopus sordidulus X

#Willow flycatcher

[gray flycatcher]

Empidonax traillii 

E. wrightii

XO

#Black phoebe Sayornis nigricans XO X X

Say’s phoebe S. saya XO

#Vermilion flycatcher Pyrocephalus rubinus XO X X X

#Ash-throated flycatcher Myiarchus cinerascens XO X X X X

#Brown-crested flycatcher M. tyrannulus XO X X X

#Tropical kingbird**o Tyrannus melancholicus XO ?p

Cassin’s kingbird T. vociferans XO X

Western kingbird T. verticalis XO X X

#Rose-throated becardq Pachyramphus aglaiae XO

#Loggerhead shrike Lanius ludovicianus XO

#Bell’s vireo Vireo bellii XO X X

#Chihuahuan raven Corvus cryptoleucus

Common raven C. corax XO X

Purple martin Progne subis XO

Violet-green swallow Tachycineta thalassina

#Northern rough-winged swallow Stelgidopteryx serripennis XO X X

Cliff swallow Petrochelidon pyrrhonota X X

Bridled titmouse Baeolophus wollweberi X

#Verdin Auriparus flaviceps XO X X

#Cactus wren Campylorhynchus 

brunneicapillus 

XO X

Rock wren Salpinctes obsoletus XO X

Canyon wren Catherpes mexicanus XO

#Bewick’s wren Thryomanes bewickii XO X X X X

#Black-tailed gnatcatcher Polioptila melanura XO X

Table 14—Continued.
200 USDA Forest Service RMRS-GTR-377. 2018

Mesquite

and c-w

Cottonwoodwillow (c-w)

 Location

GMF

Santa

Cruzb

BPC

Salt

R.c 

Verded

Valley

Lower 

Verde R.

Lower

Rio

Grande

Years of records

1902-

1963 1930s-1980s

1969-

1974 1975 1977

Source

RHW

(2014)

RRJ

(2000)

SWC

(1974)

NES 

(1978)

RE-W

(1979)

Common name Scientific namee

American robin Turdus migratorius X

#Curve-billed thrasher T. curvirostre XO X

#Bendire’s thrasher Toxostoma bendirei XO

#Crissal thrasher T. crissale XO X

#Northern mockingbird Mimus polyglottos XO X X

European starlingr Sturnus vulgaris XO X X X

#Phainopepla Phainopepla nitens XO X

#Lucy’s warbler Oreothlypis luciaes XO X X X X

#Common yellowthroat Geothlypis trichas XO X X

#Yellow warbler Setophaga petechia XO X X X

#Yellow-breasted chat Icteria virens XO X X X

#Canyon towhee Melozone fusca XO X

#Abert’s towhee M. aberti XO X X X

#Rufous-winged sparrow Peucaea carpalis XO

#Black-throated sparrow Amphispiza bilineata XO X

#Song sparrow**t Melospiza melodia XO X

#Summer tanager Piranga rubra XO X X X X

#Northern cardinal Cardinalis cardinalis XO X X X

#Pyrrhuloxia C. sinuatus XO

#Blue grosbeak Passerina caerulea XO X X X

Painted bunting P. ciris X

Red-winged blackbird Agelaius phoeniceus XO X X X

Great-tailed grackle Quiscalus mexicanus X

#Bronzed cowbirdu Molothrus aeneus XO X

#Brown-headed cowbird M. ater XO X X X X

#Hooded oriole Icterus cucullatus XO X X

Orchard oriole Icterus spurius X

#Bullock’s oriole I. bullockii XO X X X X

House finch Haemorhous mexicanus XO X X X X

#Lesser goldfinch Spinus psaltriaw XO X X X

#House finch Haemorhous mexicanusv XO X X X

House sparrowx Passer domesticus XO X X

Breeding species 83y 71 34z 28 27aa

Pairs/40 ha (100 A) ???bb 611+cc 847dd 684 354

# = Mesquites found in our studies or specifically mentioned under breeding habitat in literature, especially by Corman and Wise-Gervias 

(2005).

X = Recorded on area.

XO = Recorded on area but numbers insufficient to count or numbers not otherwise recorded.

Table 14—Continued.
USDA Forest Service RMRS-GTR-377. 2018 201

F = Flying over the area during the breeding season; often large birds foraging.

** = Locally extirpated after earlier observations. 

Sources: NES = Stamp (1978); RE-W = Engel-Wilson and Ohmart (1979); RHW = Webb et al. (2014); RRJ = Johnson et al. (2000); SWC = 

Carothers et al. (1974).

a Species may be obligate, preferential, or facultative riparian.

b Great Mesquite Forest or San Xavier Bosque, Tucson mesquite forest (Phillips et al. 1964), or Grand Mesquite Forest (Brandt 1951), located 

on the Santa Cruz River on the San Xavier Indian Reservation, 10 miles s. of Tucson. This best known mesquite forest of the Southwest was 

destroyed by the 1970s through anthropogenic causes, including agricultural clearing by Native Americans, wood-cutting, and excessive 

groundwater withdrawal (Johnson and Carothers 1982, Webb et al. 2014). 

c Blue Point Cottonwoods, located on the Salt River approximately 2 mi upstream from its confluence with the Verde River. 

d The first quantitative riparian bird study for the Southwest, documenting some of the highest avian population densities for the United States 

and even the world.

e After AOU (1998) and AOU Check-list (2015).

f

 Vorhies saw pied-billed grebes at Indian Dam, GMF, frequently in the 1930s, including three calling on open water, April 15, 1934 (Vorhies 

et al. 1935). They may have been nesting here since the species begins nesting in December and January in Arizona lowlands (Corman 

and Wise-Gervais 2005). 

g Great blue heron rookery in cottonwoods off of plots but adults foraged on plots.

h Vorhies saw green heron nests at Indian Dam, GMF, for “three years in a row” in the mid 1930s (Vorhies et al. 1935:244). 

i

 “At least a dozen” black vultures were seen “in the bottom of a ditch” on 7 May 1922, along the Santa Cruz River, 12 miles south of Tucson 

(Kimball 1923:109).

j

 Osprey was extirpated as a breeding species after being reported breeding nearby and foraging over BPC (Phillips et al.1964).

k The GMF was one of the most noted sites in Arizona for hunting white-winged doves in the early 1900s (Arnold 1941, 1943; Brown 1989). 

After visiting the GMF Bent (1937:259) wrote, “White-winged doves fairly swarmed through the thickets, and their tiresome notes were 

the dominant sounds.”

l

 Yellow-billed cuckoo, a candidate for listing as a threatened or endangered species (USFWS 2015), is the last breeding bird to arrive 

annually, generally in early June (Phillips et al. 1964; Johnson and Simpson MS b).

m Herbert Brown took a ferruginous pygmy-owl south of Tucson in 1884, the only Santa Cruz River record (Johnson et al. 2003). The Arizona 

population was listed as federally endangered in 1997 but delisted in 2007 and extirpated in the Tucson Basin by 2009. 

n Western burrowing owl (Athene cunicularia hypugea) is a “Bird of Conservation Concern” (Klute et al. 2003).

o Tropical kingbird was first reported as a breeding species for the United States along the Santa Cruz River between Tucson and GMF 

(Phillips 1940; Phillips et al. 1964). Brandt thought he saw one in the GMF but was uncertain. The species was first recorded in Arizona 

(and the United States) by a specimen taken by H.H. Kimball near Fort Lowell on May 12, 1905 (Peters 1936; Phillips et al. 1964). It is 

unknown if it was a solitary bird or a member of a breeding pair or colony.

p One of a pair collected at BPC.

q Rose-throated becard nest found in 1958 and calling male recorded in 1959 (Phillips et al. 1964).

r

 European starling introduced into United States from Europe, first reported for Arizona in 1946 (Phillips et al. 1964; Monson and Phillips 

1981).

s

 From Vermivora luciae to Oreothlypis luciae. 

t

 Song sparrow, formerly a common riparian nesting species in the Tucson region (Bendire 1872a, 1872b), was extirpated from the region 

shortly after 1895 (Phillips et al. 1964). Records by Arnold and Marshall were probably migrating individuals.

u Bronzed cowbird first reported arriving in Arizona from Mexico in 1909 (Visher 1909, Gilman 1914, Phillips et al. 1964).

v From Carpodacus mexicanus to Haemorhous mexicanus.

w From Carduelis psaltria to Spinus psaltria.

x House sparrow introduced into United States from Europe, first reported in Tucson in 1902-1903 (Howard 1906, Phillips et al. 1964).

y Includes total species breeding, foraging, and flying over area, during breeding season from 1902 to 1960s. It also includes species, 

especially those nesting, not recorded by listed ornithologists but for whom other records exist, e.g., Vorhies (1935) saw pied-billed grebes 

frequently at Indian Dam, GMF, in the 1930s and three calling on open water, April 15, 1934, as well as nesting green herons. 

z Includes all species on four sites but number of pairs only for site with greatest number.

aa Excludes a non-breeding species, white-crowned sparrow (Zonotrichia leucophrys), observed during the census period.

bb During the more than 60 years this plot was studied no one conducted a census to determine numbers of estimated breeding pairs.

cc Censusing on this 250 acre plot was not as rigorous as for other plots included in this table.

dd Of the four cottonwood plots censused in the Verde Valley, only one had not been modified either by cutting cottonwood trees 

(“phreatophyte control”) or other human intervention that impacted the number of trees/acre.

Table 14—Continued.
202 USDA Forest Service RMRS-GTR-377. 2018

Table 15—Mammals of mesquite bosques and associated riparian vegetation in the Southwest.a

Common name Scientific name Santab

Cruz R.c SandPedro R.e

Virginia oppossum Diadelphus virginiana X

Desert shrew Notiosorex crawfordi X

California leaf-nosed bat Macrotus californicus X

Yuma myotis Myotis yumanensis Xf

Cave myotis M. velifer X X

Long-legged myotisg M. volansh X

California myotis M. californicus X

Red bati Lasiurus borealis X

Hoary bat L. cinereus Xj

Western pipistrelle Parastrellus hesperus Xk

Big brown bat Eptesicus fuscus X X

Townsend’s big-eared bat Corynorhinus townsendii X

Pallid bat Antrozous pallidus X X

Brazilian free-tailed bat Tadarida brasiliensis X X

Pocketed free-tailed bat Nyctinomops femorosaccus Xl

Big free-tailed bat N. macrotis Xm

Western bonneted batn Eumops perotis X

Desert cottontail Sylvilagus auduboni X X

Antelope jackrabbit Lepus alleni X X

Black-tailed jackrabbit L. californicus X X

Harris’ antelope squirrel Ammospermophilus harrisii X X

Rock squirrel Otospermophilus variegatus X X

Round-tailed ground squirrel Xerospermophilus

tereticaudus

X X

Botta’s pocket gopher Thomomys bottae X X

Silky pocket mouse Perognathus flavus X

Bailey’s pocket mouse Chaetodipus baileyi X X

Hispid pocket mouse C. hispidus X

Desert pocket mouse C. penicillatus X X

Rock pocket mouse C. intermedius X

Merriam’s kangaroo rat Dipodomys merriami X X

Ord’s kangaroo rat D. ordii X X

American beaver Castor canadensis X

Plains harvest mouse Reithrodontomys montanus X

Fulvous harvest mouse R. fulvescens X X

Western harvest mouse R. megalotis X X

Cactus mouse Peromyscus eremicus Xo X

White-footed mouse P. leucopus X X

Deer mouse P. maniculatus X

Merriam’s mouse (mesquite mouse)p P. merriami X

Northern grasshopper mouse Onychomys leucogaster X

Southern grasshopper mouse O. torridus X X

Arizona cotton ratq Sigmodon arizonae X X

Tawny-bellied cotton rat S. fulviventer X

Yellow-nosed cotton rat S. ochrognathus X

White-throated woodrat Neotoma albigula X X

Common muskrat Ondatra zibethicus E

Common porcupine Erethizon dorsatum X

Coyote Canis latrans X X

Gray wolf r C. lupus E E

Kit fox Vulpes macrotis X X

Common gray fox Urocyon cinereoargenteus X X

American black bear Ursus americanus X

Grizzly bear U. arctos E

Ringtail Bassariscus astutus X X
USDA Forest Service RMRS-GTR-377. 2018 203

2

Common name Scientific name Santab

Cruz R.c SandPedro R.e

Northern raccoon Procyon lotor X X

White-nosed coati Nasua narica X

Long-tailed weasel Mustela frenata E

American badger Taxidea taxus X X

Northern river otter Lontra canadensis E

Western spotted skunk Spilogale gracilis X X

Hooded skunk Mephitis macroura X X

Striped skunk M. mephitis X X

White-backed hog-nosed skunk Conepatus leuconotus X X

Mountain lion (puma) Puma concolor X

Bobcat Lynx rufus X X

Jaguar Panthera onca E

Collared peccary (javelina) Pecari tajacu X X

Mule deer Odocoileus hemionus X

White-tailed deer O. virginianus X

Total species 44 60

E = Extirpated

 a List of mammals after Duncan (1988) and Soykan et al. (2007); systematic order after Baker et al. (2003) 

and Cockrum (1960); nomenclature after Baker et al. (2003) and ITIS (no date); species may be obligate, 

preferential, or facultative riparian. 

b Elevation along Santa Cruz River from 2,450+ ft to 2,900+ ft. c Records from the San Xavier bosque (Webb et al. 2014) unless otherwise noted. 

d Elevation along San Pedro River from 1,900+ ft to 4,200+ ft. e Mesquite bosques and cottonwood-willow riparian forests not separated by Soykan et al. (2007). f “Gravel Pit waterhole, S end Cottonwood Lane, Tucson” (Cockrum (1960:37). g This and some other species of bats that breed at higher elevations or in other geographic regions may be 

recorded during migration.

h Not listed by Hoffmeister (1986), perhaps either a misidentification or a migrant, since breeding localities 

in Arizona are generally 5,000 ft elevation and higher (Cockrum 1960).

i

“Probably statewide in riparian communities of the Upper Sonoran and Transitional Life Zones” 

(Cockrum 1960:53).

j

“Valencia Rd. and Santa Cruz R., Tucson” (Cockrum 1960:55), Hoffmeister (1986). k “Florida Canyon, 35 mi. S Tucson” (Cockrum 1960:50; Hoffmeister (1986). l

“Gravel Pit, S end Cottonwood Lane, Tucson” (Cockrum (1960:64), Hoffmeister (1986).

m “Mission Manor, 10 mi. S Tucson” (Cockrum (1960:65). n Mesquite specifically mentioned in habitat for Eumops perotis (Best et al. 1996). o Specimen from Contentinal, on the Santa Cruz (Cockrum 1960).

p Peromyscus merriami is so closely associated with mesquite bosques it is often called mesquite mouse. q Originally considered a subspecies of Sigmodon hispidus, S. arizonae is more likely the species here. r Extirpated; in 1850, Judge Benjamin Hayes of California wrote in his journal of the San Xavier bosque 

south of Tucson “they [wolves] were howling all around us, and one of very large size, was seen” (Davis 

1982:51).

Table 14—Continued.
204 USDA Forest Service RMRS-GTR-377. 2018

and lay their eggs in the adjacent water. In addition, numerous species of lizards and 

snakes inhabit sandy bottoms of desert washes and/or xeroriparian trees along their 

banks (Jones and Lovich 2009; Stebbins 1985). Mesquites are the first trees listed as 

habitat for the tree-dwelling ornate tree lizard (Hasse 2009) and Clark’s spiny lizard 

(Schwalbe and Rosen 2009).

A group of lowland, broadleaf, winter deciduous obligate riparian trees termed the 

riparian “big five” by Lowe (1961, 1964a) consists of cottonwood (Populus fremontii), 

willow (Salix spp.), sycamore (Platanus wrightii), ash (Fraxinus velutina), and walnut 

(Juglans major). The importance of most of the riparian trees, especially to birds, is discussed in other parts of this paper. Lowe’s riparian big five all produce flowers that are 

insignificant and generally not attractive to insects, other invertebrates, or vertebrates. 

Contrastingly, mesquite trees produce large, fragrant inflorescences in the spring and 

summer that attract insects and, in turn, insectivorous birds and amphibians and reptiles.

Numerous species of insects feed on wood, flowers, and fruits of mesquites. 

Mesquites also produce beans that are fed upon by a large number of animals while 

the only tree of the “big five” that produces large, edible fruit is the walnut with an 

outer shell so hard that few animals feed on it. By contrast, mesquite beans are fed on 

by insects, rodents (Phillips and Comus 2000), javelinas (Eddy 1961), coyotes (Lantz 

1905), and even wild turkey (Bendire 1872). Mesquite beans are fed on so extensively 

by insects that suggestions have been made to use some species of insects in biological 

control of “undesirable” mesquites on rangelands (Smith and Ueckert 1974). At times, 

coyote scat (feces) consists almost entirely of mesquite beans (R.R. Johnson, numerous 

observations, Tucson, Arizona during several decades).

A number of vertebrate species extend northward from Mexico into extreme 

southwestern United States, often occurring in mesquite bosques, or narrow xeroriparian patches of mesquites and other small trees along washes. Some reptiles, such as 

the brown vinesnake and several amphibians, discussed earlier, extend northward 

from Mexico, occurring in the United States only in southern Arizona. Another previously mentioned species, the gray hawk, is an uncommon tropical species occurring 

in the United States largely in southern Arizona. It preferentially utilizes mesquites 

and “mesquite woodland occupies large proportion of home ranges (about 54 percent), 

and majority of foraging locations (about 80 percent) occur within mesquite woodland” (Bibles et al. 2002). The range of the previously mentioned mesquite mouse, or 

Merriam’s mouse, extends northward from Mexico, occurring in the United States only 

in southern Arizona (Hall and Kelson 1959).

An arboreal food chain associated with mesquite bosques, illustrative of the 

importance of riparian mesquites, consists of arthropods, the ornate tree lizard, brown 

vinesnake, and gray hawk. This food chain extends northward from Mexican thornscrub 

(Brown 1982) into the mesquite bosques of the Santa Cruz Valley of extreme southern 

Arizona and has been documented through scientific records and by online photos. The 

ornate tree lizard, a resident of this region, forages preferentially in mesquite trees on 

“insects and other small arthropods” (Jones and Lovich 2009) and, in turn, this lizard is 

prey for the brown vinesnake. Recent photos show brown vinesnakes in mesquite trees, 

one eating an ornate tree lizard (Dancing Snake Nature Photography 2013).

Finally, completing the food chain, the gray hawk feeds “almost entirely on 

vertebrates, especially reptiles” (Bibles et al. 2002) and an online photo shows a 
USDA Forest Service RMRS-GTR-377. 2018 205

gray hawk grasping a brown vinesnake (Tucson Herpetological Society, no date). 

Both tropical species, the brown vinesnake and gray hawk, occur no farther north 

than southern Arizona. The gray hawk occurs in extreme southern Arizona and Texas 

(Bibles et al. 2002) and the brown vinesnake was first discovered for the United States 

in mesquites along the Santa Cruz River south of Tucson (Brown et al. 2015; Vorhies 

1926; Webb et al. 2014). This small snake has now been pushed farther south and, due 

to riparian habitat loss, is now restricted in the United States to near Nogales, Arizona 

(Stebbins 1985; Webb et al. 2014).

Mesquite Bosques and Cottonwood-Willow Riparian Gallery Forests

The preferential use of riparian habitats by birds is well documented. Studies in 

both mesquite bosques and cottonwood-willow forests have reported some of the highest avian population numbers and greatest biodiversity in North America (Carothers 

et al. 1974; Johnson et al. 1977; Szaro 1980). Johnson et al. (1987) list 69 percent of 

the Southwest lowlands avifauna either as obligate riparian-nesting birds or showing 

a preference for riparian and other wetland habitats. Early in-depth studies in riparian 

ecosystems of the desert Southwest concentrated largely on birds of cottonwood-willow 

(Populus-Salix) forests with later studies conducted in bosques. The first quantitative 

analysis of avian populations in southwestern riparian habitats was by Carothers et al. 

(1974). That study and a following publication (Carothers and Johnson 1975) found 

that no other habitat in North America is more important to non-colonial nesting birds 

than Fremont cottonwood (Populus fremontii) forests along the Verde River in central 

Arizona. This same finding was later echoed by Szaro (1980) in an extensive survey of 

studies in southwestern riparian forests. Studies in both bosques and cottonwood-willow forests were later conducted by others, including Gavin and Sowls (1975), Stamp 

(1978), and Engel-Wilson and Ohmart (1979) (see tables 13 and 14).

Especially important to birds are combinations of riparian mesquites and 

cottonwood-willow forests. Mesquite bosques constitute prime avian habitat (Johnson 

and Simpson 1971; Johnson et al. 2000). Cottonwood trees are generally taller than 

mesquites and provide additional structure for nesting and other avian activities. Bird 

species diversity (BSD) is directly related to foliage height diversity (FHD). FHD is 

determined by the arrangement of vegetation in three-dimensional space. A lack of 

taller trees (and thus reduction in the value of the dimension of height) also results 

in a reduction of BSD (Carothers et al. 1974; MacArthur et al. 1962). A combination 

of bosques and cottonwood-willow forests tend to support greater avian biodiversity 

than either alone (Johnson et al. 2000; Webb et al. 2014; table 14). In addition, the soft 

wood of cottonwood trees provides ideal nesting cavities in contrast to very dense mesquite wood, which is difficult for cavity-nesting birds to excavate.

Along larger rivers, notably the Lower Colorado River Valley and Rio Grande 

Valley, bosques and/or cottonwood-willow forests have been largely replaced by 

dense Tamarix (saltcedar or tamarisk) thickets. Some vertebrates, especially several 

avian species, have adapted to saltcedar, but many have not (Webb et al. 2014). 

Extensive changes to riparian ecosystems in general have also occurred with the advent of agricultural fields and rural and suburban areas that have developed in major 

river valleys, often replacing riparian ecosystems. Again, some species have adapted 
206 USDA Forest Service RMRS-GTR-377. 2018

to trees and shrubs planted in city parks and cemeteries, or to shade trees around housing areas, many have not.

Grazing has also been identified as a major impact to riparian ecosystems (Knopf 

et al. 1988). Krueper et al. (2003) found that populations of 40 species of riparian birds 

showed detectable increases within 3 years after cattle were removed from a section of 

the San Pedro River, Arizona.

In addition to birds, a large percentage of amphibians and reptiles and mammals 

have also been affected by deterioration of riparian vegetation. Most of these are ground 

dwellers, many of them living in burrows, and resulting changes on the ground include 

loss of shade, increasing temperatures, and even modification of chemical composition, 

such as increased soil salinity (Shafroth et al. 1995).

Summary

The importance of mesquite bosques to the riparian fauna and flora of lowlands 

of the western United States cannot be overemphasized and in recent times there has 

been an ever-increasing effort toward stream restoration (Appendix B; Carothers et al. 

1974; Stanley et al., in prep.) and reestablishment of cottonwood-willow riparian gallery forests. There has, however, been a general lack of effort toward reestablishment 

of riparian mesquite bosques. An online search found several references to planting 

cottonwoods and willows to reestablish riparian forests, including a 22-page manual 

published by the Natural Resources Conservation Service (Hoag 2007). Contrastingly, 

the only online references to planting of mesquites was as ornamentals or as urban 

shade trees. Until there is equal emphasis on reestablishment of mesquite bosques as 

well as cottonwood-willow riparian gallery forests, there will be a continuing, and possibly increasing, loss in biodiversity of riparian vertebrate faunas of the lowlands of the 

western United States.

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212 USDA Forest Service RMRS-GTR-377. 2018

Appendix A: Western Pioneers of Riparian Study Through the 

1980s

Compiled by: Kenneth J. Kingsley, John T. Stanley, Steven W. Carothers, R. Roy 

Johnson, Deborah M. Finch, Duncan T. Patten, Anne Sands, Bo Shelby, and Frederick 

J. Swanson

Table A.1 is a list of people who had significant early roles in the development of 

the study of riparian resources in the western United States during the period of time 

prior to and through the 1980s. Nominees were selected based on their meeting one or 

more of the following criteria involving developing the science of riparian ecology and/

or advancing the “riparian environmental movement” through the 1980s.

1. Conducted early riparian studies and published or presented results that advanced the 

establishment of riparian ecology as an emerging science.

2. Published at least three papers on some aspect of riparian science; or if fewer than 

three, a landmark or state of the art paper that would advance the establishment of 

the science. These publications are usually listed online, especially in Google Scholar 

and Web of Science.

3. Assumed leading roles in encouraging resource management agencies to support and 

set policy for management and protection of riparian ecosystems.

4. Organized conferences, symposia, riparian councils, or similar undertakings.

5. Published manuals on riparian restoration and/or completed a project that serves as a 

demonstration site that is an outstanding illustration of important restoration factors.

6. Attracted regional or national attention by advocating the study and conservation of 

riparian ecosystems.

Persons listed are ordered alphabetically within the decade of first known involvement in riparian ecology, based upon authors’ recollections and/or first known 

publication or formal presentation. Undoubtedly, some names have been inadvertently 

omitted and some aspects of the work of some individuals included below have been 

overlooked. Contributions to the science are briefly summarized based on best-known 

work, and some important contributions may have been omitted. The conferences mentioned in the table are those listed in a separate appendix in this general technical report.
USDA Forest Service RMRS-GTR-377. 2018 213

Table A.1—Listing of people who had a significant role in the development of the study of riparian resources in the western 

United States, prior to and through the 1980s. 

Person Affiliation Riparian activities and results

Pre-1960s

Willis L. Jepson University of California, 

Berkeley

Published paper in 1893 on the riparian botany of the lower 

Sacramento River. 

Lloyd G. Ingles Fresno State University First known quantified population studies of riparian nesting 

birds for California or the West.

Luna B. Leopold University of California, 

Berkeley

One of nation’s leading hydrologists; published several books 

and scientific papers on streams.

Richard G. Beidleman University of Colorado; 

Colorado College

Cottonwood riparian thesis, 1948 and similar dissertation, 1954 

at University of Colorado; numerous publications but none of 

this riparian work. 

1960s 

Robert L. Beschta Oregon State University Prolific author and co-author on watershed management, forest 

and range hydrology, sediment transport, channel morphology, 

water quality, wetlands and riparian areas, effects of logging, 

grazing, and predator presence or absence on watersheds and 

riparian areas.

Bud Bristow Arizona Game and Fish 

Department 

Secured funding for Carothers’ and Johnson’s earliest Verde 

River avian projects; actively fought “phreatophyte control” 

through public presentations and writings.

C. J. Campbell USDA Forest Service Published paper in 1968 on “pseudoriparian” plants; 

coauthored 1964 “phreatophyte” publication with Dick-Peddie.

Steven W. Carothers Museum of Northern 

Arizona; SWCA 

Environmental Consultants

Co-principal on earliest avian riparian studies along Verde R., 

first in-depth riparian avian population study for SW; several 

publications on riparian birds and riparian management and 

restoration; several publications on ecology of riparian area 

along Colorado River in Grand Canyon.

William A. Dick-Peddie New Mexico State 

University, Las Cruces

Publications on classification and ecology of riparian 

ecosystems, especially in New Mexico.

R. Roy Johnson Prescott College; National 

Park Service; University of 

Arizona 

Co-principal with Carothers on early avian studies along Verde 

River, first in depth riparian avian population study for the 

Southwest; coordinator of research on Colorado River corridor 

in Grand Canyon; numerous publications on riparian birds and 

riparian management; organized three regional and national 

riparian conferences and presented numerous papers. 

Dale A. Jones USDA Forest Service Obtained funding for Carothers and Johnson’s Verde Valley 

avian projects; organized early riparian field training sessions; 

assisted in organizing first regional riparian conference for the 

United States (1977).

Charles H. Lowe University of Arizona Developed early definition and classification of riparian 

ecosystems; first published what would become the Brown and 

Lowe classification system for Biotic Communities. 

Wendell L. Minckley Arizona State University Primarily an ichthyologist also contributed to publications on 

cienegas, riparian classification systems, and other riparian 

studies. 

Douglas Morrison USDA Forest Service The 1977 Tucson riparian conference was dedicated to Doug 

for his role in halting destruction of mature cottonwood trees as 

“phreatophyte control” projects along the Verde River in central 

Arizona. 
214 USDA Forest Service RMRS-GTR-377. 2018

Person Affiliation Riparian activities and results

Amadeo M. Rea Prescott College; San Diego 

Natural History Museum; 

University of San Diego

Studied ecological history of the Gila River Indian Community. 

Published book on riparian biota of the Gila River documenting 

losses after construction of upstream Coolidge Dam.

Kenneth Thompson University of California, 

Davis 

In 1961, published article titled “Riparian Forests of the 

Sacramento Valley, CA.” Presented papers at the 1976 and 1977 

CA conferences. 

Raymond M. Turner University of Arizona; U.S. 

Geological Survey

Pioneered repeat photography showing environmental changes 

in riparian vegetation along rivers of the southwest and mapped 

Arizona streams and wetlands. 

1970s 

Dana L. Abell University of California, 

Davis 

Conducted research on river systems in Sequoia NP. Member of 

steering committee for 1981 California conference; Technical 

Coordinator of 1988 California conference. 

Bertin W. Anderson Arizona State University; 

Revegetation and Wildlife 

Management Center, Blythe, 

CA

Conducted research on experimental revegetation of riparian 

trees and their use by birds along the lower Colorado River 

and Kern River Preserve; published papers in journals and 

conference proceedings and co-author of book on birds of the 

lower Colorado River. 

George T. Austin University of Nevada, Las 

Vegas 

Conducted early studies of birds in desert riparian ecosystems, 

as well as butterflies and birds of Nevada.

W. James Barry California Department of 

Parks and Recreation 

Leader in protection and management of riparian ecosystems 

in California’s State Park System; member of California Riparian 

Revegetation Study Group; presented papers at 1981, 1984 

and 1987 California conferences; advisor for 1988 California 

conference. 

Mark M. Brinson East Carolina University An exceptionally versatile expert on water based ecosystems, 

including not only riparian wetlands of the West but also coastal 

and Eastern wetlands. 

David E. Brown Arizona Game and Fish 

Department; Arizona State 

University.

Published classification system for Biotic Communities of North 

America with Charles H. Lowe, map of Arizona’s perennial 

streams and wetlands, and book on wetlands and waterfowl in 

Arizona. 

James W. Burns Office of California Secretary 

of State

Led Sacramento River Task Force. Presented papers at 1976, 

1977, and 1981 California conferences and paper at 1978 

Georgia conference.

Kenneth W. Cummins OR State University; 

Humboldt State University, 

California

Numerous publications on interaction between aquatic and 

riparian ecosystems; co-author of River Continuum Concept.

Robert B. Ditton Texas A&M University Studied riparian mammals, recreation, and livestock impacts in 

Big Bend National Park.

Jerry F. Franklin USDA Forest Service; 

University of Washington; 

National Science Foundation 

(NSF)

Large number of publications on riparian and related forest 

issues; recipient of numerous awards for conservation and 

related environmental issues; chaired and served on many 

related boards and commissions; Program Officer at NSF that 

provided funding for River Continuum Concept.

David A. Gaines Davis Audubon Society, CA; 

Mono Lake Committee, CA 

Conducted breeding bird censuses in Sacramento Valley 

riparian woodland in early 1970s; organized 1976 California 

conference; member of 1977 California conference committee; 

led Mono Lake Committee; published and presented papers on 

riparian birds and destruction of riparian habitat at 1977 and 

1981 California conferences. 

Richard L. Glinski Arizona State University; 

Arizona Game and Fish

Conducted studies on riparian raptors; published several papers 

and books. 

Table A.1—Continued.
USDA Forest Service RMRS-GTR-377. 2018 215

Person Affiliation Riparian activities and results

Bernard H. Goldner Santa Clara Valley Water 

District, San Jose, CA 

Managed contracts for riparian revegetation projects in 

Santa Clara County, CA; member and host of CA Riparian 

Revegetation Study Group; presented papers at 1981, 1987, and 

1988 California conferences.

Lois T. Haight National Park Service Co-author of 25 publications, mostly on riparian birds; assisted 

with organization of 1977 Tucson riparian conference. 

Robert H. Hamre USDA Forest Service Editor and leading advocate of publications of several riparian 

conference proceedings. 

Burchard H. Heede USDA Forest Service Studies and publications on stream dynamics and riparian 

vegetation.

Robert F. Holland California Department of 

Fish and Game

Developed system for the classification of natural communities 

in California. Served on Advisory Committee for 1988 CA 

conference and presented papers at 1977 and 1988 California 

conferences. Co-author of field guide to common riparian plants 

of California.

J. Greg Howe University of California, 

Davis

Member of Conference Committee and assistant editor of 

proceedings for 1977 California conference. Co-author of 

several conference papers.

John P. Hubbard Delaware Museum of 

Natural History; New 

Mexico Game & Fish 

Early inventories of the Gila River Valley in New Mexico. 

Publications on birds of New Mexico and classification and 

inventory of riparian ecosystems in New Mexico.

Edward A. Keller University of California, 

Santa Barbara

Presented papers at 1977 and 1988 California conferences 

on fluvial systems and processes to minimize adverse effects 

of stream channelization. Co-author of 1989 USFWS report 

profiling riparian habitats of southern California.

Stephen Laymon University of California, 

Berkeley

Conducted 5-year study of riparian ecosystem at Dog Island 

on the Sacramento River; presented papers at California 

conferences. Involved in riparian revegetation at TNC Kern River 

Preserve.

Robert E. Manning University of Vermont 1979 review paper on impacts of recreation use on riparian 

soils and vegetation. Author of best-known text on recreation 

research, “Studies in Outdoor Recreation” (1986; third edition 

2011).

Robert D. Ohmart Arizona State University Studies of riparian ecology along the lower Colorado River; 

several publications, largely with B.W. Anderson and graduate 

students on aspects of riparian ecology. 

Duncan T. Patten Arizona State University; 

Montana State University 

Founding President, Arizona Riparian Council 1985-1989. 

Studies of several southwestern and eastern Sierra riparian areas 

and mentor of several leading riparian ecologists; numerous 

publications.

David R. Patton USDA Forest Service Published numerous papers on management and research needs 

for riparian habitat.

William S. Platts USDA Forest Service Studied impacts of grazing on riparian ecosystems; published 

large number of scientific papers. Lead author of 1987 

“Methods for Evaluating Riparian Habitats with Applications to 

Management.”

John N. Rinne USDA Forest Service Studied impacts of grazing on riparian and aquatic ecosystems, 

with emphasis on fish and reptiles. 

Anne Sands Riparian Systems, Davis, CA Coordinator and Editor of 1977 California conference; member 

of Advisory Committee for 1988 California conference; author 

or co-author of papers presented at riparian conferences; chair 

of the Riverlands Council. Co-author of 1989 USFWS report 

profiling riparian habitats of southern California.

Table A.1—Continued.
216 USDA Forest Service RMRS-GTR-377. 2018

Person Affiliation Riparian activities and results

David J. Schmidly TX A&M University Studied riparian mammals, recreation, and livestock impacts in 

Big Bend National Park.

James R. Sedell Oregon State University; 

Weyerhauser Corp.; USDA 

Forest Service

Studied forest-stream interactions in Pacific Northwest. Coauthor of River Continuum Concept, led aquatic component of 

FEMAT/NW Forest.

Bo Shelby Oregon State University; 

Confluence Research and 

Consulting

River/riparian recreation studies in Grand Canyon (1976), later 

Alaska and Lower 48; studies of flows for recreation from Alaska 

to Georgia; organized national conferences and agency training 

programs. Publications on recreation impacts, flows, and hydrolicensing. 

Felix E. Smith US Fish and Wildlife Service Presented papers at 1976, 1977 and 1981 California riparian 

symposia. 

Lawrence E. Stevens Prescott College; Northern 

Arizona University; Museum 

of Northern Arizona

Assisted Carothers and Johnson in Verde River study. Leading 

expert on North American springs; founder of Springs 

Stewardship Institute; published studies of many taxa, especially 

insects, in the Colorado Plateau region and Colorado River 

corridor. 

Frederick J. Swanson Oregon State University; 

USDA Forest Service 

Studied forest-stream interactions in Pacific Northwest region; 

co-author of numerous publications; input into development of 

management policies. 

Dale A. Zimmerman Western New Mexico 

University

Studied birds of the Gila River, New Mexico. Several U.S. Forest 

Service reports; published notes on riparian birds.

Bill Zeedyk USDA Forest Service; 

Restoration Services

Published manual on riparian restoration “Let the Water do the 

Work.”

1980s 

Betty Andrews Friends of the River; Philip 

Williams and Associates, San 

Francisco, CA

Provided hydrologic and geomorphic consulting expertise for 

multiple stream and riparian restoration projects in California; 

member of steering committee for 1981 California conference.

Ward Brady Arizona State University Several riparian papers, especially on regeneration and growth 

of riparian forests.

Bryan T. Brown National Park Service; 

University of Arizona; SWCA 

Riparian studies on birds along the Colorado River in the Grand 

Canyon; studied fauna of Rio Yaqui in Sonora, Mexico. 

Thomas C. Brown USDA Forest Service Organized 1989 Forest Service funding which led to 1992 short 

book on instream flows for recreation and aesthetics with Bo 

Shelby and Jonathan Taylor. Several other articles and book 

chapters about benefits of instream flows for recreation and 

aesthetics.

David N. Cole Aldo Leopold Wilderness 

Research Institute 

Publications and agency training on measuring and mitigating 

recreation impacts at camps and attraction sites. Organized 

conferences and training sessions on stewardship, with a focus 

on recreation impacts in alpine and riparian settings.

David W. Crumpacker University of Colorado, 

Boulder

Published several papers on riparian ecosystems and their 

management.

Clifford N. Dahm University of New Mexico Published numerous papers on riverine activities and 

interactions between streams and riparian ecosystems.

Wayne Elmore Bureau of Land Management Studied effects of livestock grazing on riparian areas and 

restoration methods on public lands in the Great Basin 

of Eastern Oregon. Presented papers at 1988 California 

conference. Author or co-author of several papers on riparian 

ecology.

Table A.1—Continued.
USDA Forest Service RMRS-GTR-377. 2018 217

Person Affiliation Riparian activities and results

Christopher C. Estes Prescott College; Washington 

State University; Alaska 

Dept. Fish and Game; Chalk 

Board Enterprises, LLC

Aquatic Resources & Habitat Scientist, involved in Verde 

Valley, AZ, riparian avian studies in 1970s; as Chief of Alaska 

Department of Fish and Game, Aquatic Resources Coordination 

Unit and At-Large Director, Instream Flow Council, integrated 

riparian habitat assessment elements in instream flow/water 

level conservation actions and publications.

Phillis M. Faber CA Native Plant Society; 

Pickleweed Press, Mill 

Valley, CA

Editor of articles on riparian plants of California in CNPS 

journal Fremontia (1983-1999). Co-author of 1988 field guide 

to common riparian plants of California. Lead author of 1989 

USFWS biological report on ecology of riparian habitats of 

southern California. Editor of 2001 California riparian systems 

conference.

Deborah M. Finch USDA Forest Service Field researcher and co-author or author of more than 150 

published papers, especially on riparian birds; technical 

coordinator or editor of many USFS research reports. 

Randall L. Gray Natural Resources 

Conservation Service

Co-author of several publications, and presentations at riparian 

conferences; co-founder of California Riparian Revegetation 

Study Group. 

Stanley V. Gregory Oregon State University Studied forest-stream interactions in Pacific Northwest region. 

F. Thomas Griggs The Nature Conservancy, CA; 

River Partners, Chico, CA 

TNC Project Manager for riparian restoration throughout 

California. Served on Advisory Committee for 1988 California 

conference. Author/co-author of several papers on riparian 

habitat restoration. Author of California Riparian Habitat 

Restoration Handbook.

Barry Hecht HEA, Berkeley, CA; Balance 

Hydrologics Inc., Berkeley, 

CA

Provided hydrologic and geomorphic consulting expertise for 

multiple stream and riparian restoration projects in California. 

Member of CA Riparian Revegetation Study Group; presented 

paper at 1981 California riparian conference and other 

conferences.

Dean A. Hendrickson Arizona State University; 

University of Texas, Austin

Primarily an ichthyologist but also a leading U.S. expert on 

cienegas. 

Kathleen M. Hendrix University of California, 

Davis

Assistant Coordinator and Assistant Editor of 1981 California 

Riparian Systems Conference. Co-author of California 

Department of Fish and Game report on the riparian resources 

of the Central Valley and California desert.

William C. Hunter Arizona State University; 

U.S. Fish and Wildlife 

Service

Author or co-author of several papers on riparian birds. Coauthor with Ohmart, Anderson, and Rosenberg of book on Birds 

of the Lower Colorado River Valley (published in 1991). 

Diana Jacobs California State Lands 

Commission

Member of Advisory Committee for 1988 California Riparian 

Systems Conference. Chair of SB1086 Riparian Habitat 

Committee overseeing preparation of Sacramento River Riparian 

Habitat Management Plan. Major contributor to California 

public trust reports on California rivers and delta.

William L. Jackson Bureau of Land Management 

and National Park Service

Early flow studies for BLM. Lead author on 1989 value-based 

interdisciplinary approach for protecting instream flows applied 

to rivers in Arizona, Colorado, and Alaska. Managed water 

operations branch for NPS, including the National Water 

Quality Assessment in the national parks. 

Sherman E. Jensen White Horse Associates, 

Smithfield, UT

Studied riparian-stream interactions and soils. Several 

publications, especially on Great Basin riparian and wetland 

soils; co-author of riverine publications.

K. Bruce Jones Environmental Protection 

Agency, Las Vegas, NV

Author of numerous publications on amphibians and reptiles; 

authority on the riparian herpetofauna of the Southwest.

Table A.1—Continued.
218 USDA Forest Service RMRS-GTR-377. 2018

Person Affiliation Riparian activities and results

Edwin F. Katibah University of California, 

Berkeley

Principal investigator for California Department of Fish and 

Game Central Valley Riparian Mapping Project. Presented paper 

at 1981 California conference on distribution and condition of 

riparian resources.

Kenneth J. Kingsley University of Arizona; 

National Park Service; SWCA 

Published studies on irrigated pecan orchard as a riparian 

ecosystem and riparian taxa associated with springs, temporary 

aquatic habitats, and washes; presented paper at 1985 Arizona 

conference.

Fritz L. Knopf US Fish and Wildlife Service Author or co-author of several papers on riparian birds and 

conservation of riparian woodlands.

G. Mathias Kondolf University of California; 

Santa Cruz and Berkeley

Consulting hydrologist for multiple stream and riparian 

restoration projects. Developed a database of stream restoration 

projects in California as part of the National River Restoration 

Science Synthesis. Presented papers at 1981 and 1988 

California conferences and other conferences. 

Jon Kusler Founder and Executive 

Director, Association of State 

Wetland Managers

With a law degree has been instrumental in formulating and 

improving wetland and floodplain programs and laws for 

numerous states; served on the staffs of the University of Mass. 

and Harvard University; editor of numerous publications on 

wetlands and floodplains. 

Laurel Marcus California Coastal 

Conservancy

Managed cooperative planning for stream and riparian 

restoration; published papers on riparian restoration and 

watershed management at 1987 SWS Seattle conference, 1988 

AWM Oakland symposium and 1988 California conference.

Jeffrey L. Marion Virginia Tech University 1988 review paper on recreation impacts in eastern riparian 

forests with David Cole. Leader in recreation ecology research, 

with several review articles.

Joe R. McBride University of California, 

Berkeley

Conducted and published research on riparian trees, seed 

dispersal and hydrology, role of fire in riparian woodlands, and 

urban forestry. 

L. Richard Mewaldt San Jose State University Founder of San Jose State University Avian Biology Laboratory, 

Point Reyes Bird Observatory, San Francisco Bay Bird 

Observatory and Coyote Creek Riparian Station. Active in bird 

banding in riparian systems; author or co-author of numerous 

avian papers. 

G. Wayne Minshall, Idaho State University Numerous publications on interaction between aquatic and 

riparian ecosystems; co-author of River Continuum Concept.

Robert J. Naiman Oregon State University; 

University of Washington

Early studies on beavers and riparian ecosystems; studied 

forest-stream interactions in Pacific Northwest region; numerous 

publications. 

John Rieger California Department of 

Transportation

Research and publications on nesting habitat requirements 

for endangered Least Bell’s Vireo; designed and supervised 

installation of riparian revegetation projects in San Diego region 

for Caltrans; coordinator and editor of 1985 and 1987 native 

plant revegetation symposia; member of California Riparian 

Revegetation Study Group. 

Michael Rigney Coyote Creek Riparian 

Station, San Jose, CA

Co-founder/director of Coyote Creek Riparian Station; 

participated in design of riparian revegetation projects; 

monitored avian use of riparian revegetation sites. Member of 

California Riparian Revegetation Study Group. Presented paper 

at 1988 California conference.

Table A.1—Continued.
USDA Forest Service RMRS-GTR-377. 2018 219

Person Affiliation Riparian activities and results

Ann L. Riley California State Water 

Resources Control Board; 

Waterways Restoration 

Institute

Co-founder of the Urban Creeks Council and Coalition to 

Restore Urban Waters; Executive Director of Waterways 

Restoration Institute; coordinated redesign of flood channel 

restoration projects; member of conference advisory 

committees; author of books on restoration of urban riparian 

areas and waterways; member of California Riparian 

Revegetation Study Group. Presented papers at multiple 

conferences.

Stewart B. Rood University of Lethbridge, 

Alberta, Canada

Plant physiologist with many publications on the ecology of 

riparian trees, especially water needs and establishment of 

Populus spp. 

Kenneth V. Rosenberg Arizona State University; 

Louisiana State University; 

Cornell University

Published several papers on riparian birds and suburban 

habitats. 

David L. Rosgen USDA Forest Service; 

Wildland Hydrology

Developed a system of stream classification based on 

morphology, and related this to evaluating riparian conditions 

and planning restoration efforts. Involved in the design of 

numerous stream and riparian restoration projects.

Ronald F. Schultze USDA Soil Conservation 

Service, Davis, CA 

Member of steering committee for 1981 and 1988 California 

conferences. Presented papers on restoration at 1981 CA 

conference and 1985 Arizona conference. Co-founder of 

California Riparian Revegetation Study Group (ca. 1981).

John T. Stanley Harvey and Stanley 

Associates, Inc., Alviso, CA; 

The Habitat Restoration 

Group, CA 

Involved in planning and installation of riparian and stream/

watershed restoration projects; co-chair of CA Riparian 

Revegetation Study Group; active in organizing several 

conferences, authoring or co-authoring publications, including 

chapter on riparian wetland creation and restoration in the far 

west in 1989 EPA publication.

Juliet C. Stromberg Arizona State University Many publications and presentations on riparian plant ecology, 

particularly along the San Pedro River; active in Arizona 

Riparian Council and activist for riparian conservation. 

Robert C. Szaro USDA Forest Service Several publications on riparian vegetation and wildlife, 

including studies of effects of various perturbations.

Richard E. Warner Field Studies Center, 

University of California, 

Davis

Editor of 1979 USFWS workshop proceedings on fish and 

wildlife resources needs in riparian ecosystems. Technical 

Coordinator and Editor of 1981 California riparian conference. 

Presented paper at the 1981 California Riparian Systems 

Conference. Lead author of 1985 California Department of 

Game and Fish report on the riparian resources of the Central 

Valley and California Desert.

Doug Whittaker Bureau of Land 

Management, National Park 

Service, and Confluence 

Research and Consulting

River/riparian recreation planning and management studies 

on Alaska and Lower 48 rivers starting in the late 1980s. Coorganized first national conferences and agency training on 

flows and recreation. Papers on recreation impacts, flows, and 

hydro-licensing. 

Philip B. Williams Philip Williams and 

Associates, Inc., San 

Francisco, CA

Hydrologist/civil engineer involved in design of numerous 

stream and riparian restoration projects; member of California 

Riparian Revegetation Study Group. Author or co-author of 

several papers on riparian corridor management and restoration, 

esp. in relation to flood control. 

Table A.1—Continued.
220 USDA Forest Service RMRS-GTR-377. 2018

Appendix B: Conferences, Symposia, and Other Gatherings 

Pertaining To Riparian Ecosystems, Riparian Ecology, 

Riparian Habitat Restoration, and Riparian Area Conservation

Compiled by: John T. Stanley, R. Roy Johnson, Steven W. Carothers, Duncan T. 

Patten, and John S. Richardson

Beginning in the late 1970s, research of riparian systems increased. This ecosystem of decreasing cover across the semi-arid West captured more interest among 

academia, government agencies, and the interested public. Consequently, several conferences and symposia were developed to share information on riparian ecosystems, their 

processes, and management of these threatened ecosystems.

Table B.1 is a list of conferences and symposia that focused on riparian ecosystems, riparian ecology, riparian habitat restoration, and/or riparian area conservation. 

This list was compiled to provide an understanding of the timing and progression of the 

evolution of the field of riparian ecology in western North America. Some, but not all, 

of these conferences/symposia have the word riparian in their title. Not every conference focused exclusively on riparian ecosystems; however, those that addressed other 

ecosystems (e.g., wetlands other than riparian ecosystems) also contained a significant 

number of papers related to riparian ecosystems and are included here. Emphasis was 

placed on documenting conferences that occurred in the early years of development 

of the field (1970s and 1980s) while conferences from the 1990s and 2000s are also 

included (cut-off date is 2010). This cut-off date was selected by the team members 

because, after exhaustive research, we were able to find information on conferences and 

symposia up to that date; however, since 2010 many more agencies and organizations 

have been conducting riparian conferences and symposia unfamiliar to us. Due to space 

limitations, we have omitted almost all of the worthwhile riparian workshops that also 

occurred during these years.

We have referenced each conference/symposium by the editor(s), date, and title of 

the conference proceedings. Conferences are listed in chronological order in which they 

were held, which is not necessarily the years in which the proceedings were published. 

Some of the types of conferences that have been omitted from this table include those 

addressing aquatic ecosystems, including instream habitat improvement, salmonid 

stream restoration, trout stream improvement, as well as floodplain management, watershed management, rangeland management, and invasive species management, even 

though many of these conferences have also included some papers dealing with riparian 

ecosystems.

Many of these conferences were organized by individuals who were experts in 

riparian ecology or had a great interest in helping disseminate information on riparian 

ecosystems. The names of these individuals are shown below as editors and/or coordinators of the conference proceedings listed; see also the Appendix A, riparian pioneers, 

this volume. We have also provided the URL for those conference materials, which are 

hard to locate on the web.
USDA Forest Service RMRS-GTR-377. 2018 221

The following criteria were applied for the selection of conferences included in 

table B.1:

1. Word “riparian” is included in conference title and majority of papers deal with 

riparian ecosystems, riparian ecology, riparian habitat restoration, and/or riparian area 

conservation and management.

2. Word riparian is not included in conference title; however, a sizeable percentage of 

the papers deal with the above.

3. Selected papers presented at conference reported on some of the earliest work leading 

to the development of the “riparian” fields mentioned above.

4. Conference represents a key link in the history of the development of the “riparian” 

fields mentioned above.

5. Conference was held in the western United States, or if held elsewhere, a sizeable 

percentage of the papers dealt with western riparian areas.

6. Conference may have focused on river systems and their management; however, 

a sizeable percentage of the papers addressed the role of riparian ecosystems as a 

component of riverine systems and their restoration.

Annotations next to date in the left-hand column indicate the relative availability 

of conference proceedings:

a = Conference proceedings were published in hard copy shortly after, or within a 

few years of, conference date; proceedings are readily available from academic 

institutions, government agencies, or on the web.

b = Conference proceedings were published in hard copy within a few years of 

conference date; however, proceedings do not appear to be readily available from 

academic institutions, government agencies, or on the web.

c = Conference proceedings were published, but only electronically, within a few years 

of conference date; proceedings are readily available on the web.

d = Conference proceedings were not published at time of conference; however, 

proceedings have been scanned and are currently available on the web from an 

electronic content provider (generally at a fee).

e = Conference proceedings were not published (or cannot confirm publication) nor are 

the proceedings available on the web; however, selected papers from the conference 

have been published in scientific journals and/or are available on the web.

f = Conference proceedings were never published nor were selected papers ever 

published; only hard copy materials were distributed to individuals in attendance.

g = Conference proceedings were never published nor were selected papers ever 

published; however, abstracts and/or a summary/overview were prepared and 

distributed to participants but are not readily available.

h = Conference proceedings were never published nor were selected papers ever 

published; no abstracts or summary were prepared.
222 USDA Forest Service RMRS-GTR-377. 2018

Since the early 1980s, a number of statewide organizations have been formed by 

individuals interested in exchanging information on riparian ecosystems, riparian area 

conservation, and riparian habitat restoration. Two of the earliest were the California 

Riparian Revegetation Study Group, organized circa 1981, and the Arizona Riparian 

Council, organized in 1985–1986. Table B.2 lists some of these organizations that 

have held annual, semi-annual, or regular meetings, workshops, and/or conferences 

on riparian topics. Not included in this table are the many professional societies (e.g., 

American Fisheries Society, American Society of Civil Engineers, California Invasive 

Plant Council, California Society for Ecological Restoration, Society for Ecological 

Restoration, Society for Range Management, Society of Wetland Scientists, The 

Wildlife Society) and conservation organizations (e.g., watershed councils, State native plant societies) and their chapters that have included riparian issues at their annual 

conferences and workshops and are to be credited with playing a significant role in 

advancing the field of riparian ecology. Additionally, a myriad of place-based non-profit 

organizations were formed in the latter half of the 20th century to address the protection, 

management, and restoration of specific river systems and their watersheds. Typically 

named after the river, creek, or watershed where they were located, these organizations 

also have held conferences and trainings promoting the care and management of riparian corridors. Finally, there have been agency personnel and academic faculty who have 

taken it upon themselves to conduct workshops and field trainings pertaining to riparian 

ecology and riparian ecosystem restoration. We mention two of these individuals in 

table B.1.
USDA Forest Service RMRS-GTR-377. 2018 223

Table B.1—Conferences, symposia, and other gatherings focused on riparian ecosystems, riparian ecology, riparian habitat 

restoration, and/or riparian area conservation.

Year held Publications including date and location of event

1976g Gaines, D., ed. 1976. Abstracts from the conference on the riparian forests of the Sacramento Valley; 

1976 May 22; Chico, CA. Oakland, CA: Davis and Altacal Audubon Societies. 25 p. (Hard copy but not 

published).

1977a Sands, A., ed. 1977. Riparian forests in California: Their ecology and conservation: A symposium; 1977 

May 14; Davis, CA. Davis, CA: Institute of Ecology, University of California, Davis, and Davis Audubon 

Society. Institute of Ecology Publication No. 15. 122 p.

1977e

Menke, J., ed. Circa 1981. A symposium on livestock interactions with wildlife, fisheries, and their 

environments; 1977 May; Sparks, NV. Berkeley, CA: U.S. Department of Agriculture, Forest Service, 

Pacific Southwest Forest and Range Experiment Station. Refer to: Platts, W.S.1979. Livestock interactions 

with fish and their environments: A symposium summary. p. 92-96. http://www.tws-west.org/transactions/

Platts%20’78.pdf.

1977a Johnson, R.R.; Jones, D.A., tech. coords. 1977. Importance, preservation, and management of riparian 

habitat: A symposium; 1977 July 9; Tucson, AZ. Gen. Tech. Rep. RM-43. Fort Collins, CO: U.S. 

Department of Agriculture, Forest Service, Rocky Mountain Research Station. 217 p.

1978b Graul, W.D.; Bissell, S.J., tech. coords. 1978. Lowland river and stream habitat in Colorado: A 

symposium; 1978 October 4–5; Greeley, CO. Greeley, CO: Colorado Chapter, Wildlife Society and 

Colorado Audubon Council. 195 p.

1978a Cope, O.B., ed. 1979. Proceedings of the forum: Grazing and riparian/stream ecosystems; 1978 Nov. 

3–4; Denver, CO. Arlington, VA: Trout Unlimited. 94 p.

1978a

Johnson, R.R.; McCormick, J.F., tech. coords. 1979. Strategies for protection and management of 

floodplain wetlands and other riparian ecosystems: Proceedings of the symposium; 1978 December 

11–13; Callaway Gardens, GA. Gen. Tech. Rep. GTR-WO-12. Washington, DC: U.S. Department of 

Agriculture, Forest Service. 410 p.

1979b

Warner, R. E. 1979. Proceedings of a workshop on fish and wildlife resource needs in riparian 

ecosystems; 1979 May 30–31; Harpers Ferry, WV. Kearneysville, WV: U.S. Department of Interior, Fish 

and Wildlife Service, National Water Resources Analysis Group, Eastern Energy and Land Use Team, U.S. 

Fish and Wildlife Service. 53 p.

1979b

Great Plains Agricultural Council, Forestry Committee (Program Committee) 1979. Riparian and wetland 

habitats of the Great Plains: Proceedings of the 31st annual meeting; 1979 June 18–21; Fort Collins, 

CO. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range 

Experiment Station. Great Plains Agricultural. Council. Publ. No. 91. 88 p.

1981a Warner, R.E.; Hendrix, K.M., eds. 1984. California riparian systems: Ecology, conservation and productive 

management; 1981 September 17–19; University of California, Davis, CA. Berkeley, CA: University of 

California Press. 1,035 p.

1984a Rieger, J.P.; Steele, B.A., eds. 1985. Proceedings of the native plant revegetation symposium; 1984 

November 15; San Diego, CA. California Native Plant Society. 99 p.

1985a

Johnson, R.R.; Ziebell, C.D.; Patton, D.R.; Ffolliott, P.F.; Hamre, R.H., tech. coords. 1985. Riparian 

ecosystems and their management: Reconciling conflicting uses - First North American riparian 

conference; 1985 April 16–18; Tucson, AZ. Gen. Tech. Rep. RM-120. Fort Collins, CO: U.S. Department 

of Agriculture, Forest Service, Rocky Mountain Research Station. 523 p.

1985g Patten, D.T., ed. 1986. Report to Pacific Gas and Electric Co. and Southern California Edison Co.: 

Riparian workshop; 1985 November 13–14; Sam Ramon, CA. Pacific Gas and Electric Company, 

Research and Development. Tempe, AZ: Arizona State University, Center for Environmental Studies. 24 p.

1986e Salo, E.O.; Cundy, D.W., eds. 1987. Streamside management: Forestry and fishery interactions 

- proceedings of the symposium; 1986 February 12–14; Seattle, WA. Seattle, WA: University of 

Washington, Institute of Forest Resources, Contribution No. 57. 471 p.

1987a Mutz, K.M.; Lee, L.C., tech. coords. 1987. Proceedings of the Society of Wetland Scientists’ eighth annual 

meeting: Wetland and riparian ecosystems of the American west; 1987 May 26–29; Seattle, WA. Denver, 

CO: Planning Information Corporation. 346 p.
224 USDA Forest Service RMRS-GTR-377. 2018

Year held Publications including date and location of event

1987a Raedeke, K. J., ed. 1988. Streamside management: Riparian wildlife and forestry interactions; 1987 

February 11–13; Seattle, WA. Seattle, WA: University of Washington, College of Forest Resources, 

Institute of Forest Resources, Contribution No. 59. 277 p.

1987a Rieger, J.P.; Williams, B.K., eds. 1988. Proceedings of the second native plant revegetation symposium; 

1987 April 15–18; San Diego, CA. Madison, WI: Society for Ecological Restoration and Management. 

220 p.

1988f USFS Region 5 1988. Riparian management workshop; 1988 June 14–17; Reno, NV. U.S. Department of 

Agriculture, Forest Service, Region 5. (Unpublished binder of presentation materials.) 

1988a Kusler, J.A.; Daly, S.; Brooks, G., eds. 1988. Proceedings of the national wetland symposium: Urban 

wetlands; 1988 June 26–29; Oakland, CA. Berne, NY: Association of State Wetland Managers, Inc. 402 p.

1988b Mutz, K.M.; Cooper, D.J; Scott, M.L.; [et al.], tech. coords. 1988. Restoration, creation and management 

of wetland and riparian ecosystems in the American West: Proceedings of a symposium; 1988 November 

14–16; Denver, CO. Boulder, CO: Society of Wetland Scientists, Rocky Mountain Chapter. 239 p.

1988a

Abell, D.L., tech. coord. 1989. Proceedings of the California riparian systems conference: Protection, 

management and restoration for the 1990’s; 1988 September 22–24; Davis, CA. Gen. Tech. Rep. PSW110. Berkeley, CA: U.S. Department of Agriculture, Forest Service, Pacific Southwest Forest and Range 

Experiment Station. 544 p. 

1989a Gresswell, R.E.; Barton, B.A.; Kershner, J.L., eds. 1989. Practical approaches to riparian resource 

management: An educational workshop; 1989 May 8–11; Billings, MT. Billings, MT: U.S. Department of 

the Interior, Bureau of Land Management. 193 p.

1989f USFS Region 5 1989. Integrated approaches to riparian area management through forest plan 

implementation; 1989 November 13–17; Sequoia National Forest, Porterville, CA. U.S. Department of 

Agriculture, Forest Service, Region 5. (Unpublished binder of presentation materials.)

1990h Kusler, J. 1990. Urban stream and river corridors: A multi-objective management symposium; 1990 April 

23–28; Portland, OR. Portland Oregon Audubon Society in cooperation with Association of State Wetland 

Managers.

1991d Anonymous 1994. Riparian resources: Proceedings of a symposium on the disturbances, management, 

economics, and conflicts associated with riparian ecosystems; 1991 April 18–19; Logan, UT. Logan, UT: 

Utah State University, College of Natural Resources. Natural Resources and Environmental Issues 1. 82 p.

1991a Clary, W.P.; McArthur, E.D.; Bedunah, D.; [et al.], comps. 1992. Proceedings—Symposium on ecology 

and management of riparian shrub communities; 1991 May 29–31; Sun Valley, ID. Gen. Tech. Rep. INT289. U.S. Department of Agriculture, Forest Service, Intermountain Research Station. 232 p.

1992d

McCoy, M.C., ed. 1992. California’s river heritage: A conference on conservation issues, policy and 

implementation strategies; 1992 May 18–19; Sacramento, CA. Davis, CA: University of California 

Extension. (Not numbered.) https://books.google.com/books/about/California_s_River_Heritage.

html?id=092oPgAACAAJ

1993a

Tellman, B.; Cortner, H.; Wallace, M.; [et al.], tech. coords. 1993. Riparian management: Common 

threads and shared interests, a western regional conference on river management strategies; 1993 

February 4–6; Albuquerque, NM. Gen. Tech. Rep. RM-226. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station. 419 p.

1994f McCoy, M.C., ed. 1994. Conservation of California’s rivers: A practicum in tools and techniques; 1994 

June 15–16; Sacramento, CA. (Not numbered.)

1995f Laird, A.; Jacobs, D.; McCoy, M., chairs. 1995. California’s riparian-river ecosystems conference IV: 

Addressing current land use and resource conflicts; 1995 November 14–16; Sacramento, CA. Davis, CA: 

UC Davis Extension. (Not numbered.)

1995a Laursen, S.B., ed. 1995. At the water’s edge: The science of riparian forestry: Conference proceedings; 

1995 June 19–20; Duluth, MN. University of Minnesota, Minnesota Extension Service, Minnesota 

Environmental Initiative. 160 p.

1995a

Macdonald, K.B.; Weinmann, F., eds. 1997. Wetland and riparian restoration: Taking a broader view—

Proceedings of a conference; 1995 September 14–16; Seattle, WA. Society for Ecological Restoration 

International Conference. EPA 910-R-97-007. Seattle, WA: U.S. Environmental Protection Agency, Region 

10, Office of Ecosystems and Communities, Aquatic Resources Unit. 326 p.

Table B.1—Continued.
USDA Forest Service RMRS-GTR-377. 2018 225

Year held Publications including date and location of event

1995a

Shaw, D.W.; Finch, D.M., eds.1996. Desired future conditions for Southwestern riparian ecosystems: 

Bringing interests and concerns together; 1995 September 18–22; Albuquerque, NM. Gen. Tech. Rep. 

RM-GTR-272. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest 

and Range Experiment Station. 359 p.

1998a

Finch, D.M.; Whitney, J.C.; Kelly, J.F.; [et al.], eds. 1999. Rio Grande ecosystems: Linking land, water, and 

people: Toward a sustainable future for the Middle Rio Grande Basin; 1998 June 2–5; Albuquerque, NM. 

Proc. RMRS-P-7. Ogden, UT: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research 

Station. 245 p.

1999a

Native Plant Society of New Mexico 1999. Wetland and riparian plant communities of New Mexico: 

New Mexico Native Plant Society 1999 symposium proceedings; 1999 August 26–29; Albuquerque, NM. 

Santa Fe, NM: New Mexico Native Plant Society. (Copy at UNM Centennial Science and Engineering 

Library.)

2000a Wigington, P.J., Jr.; Beschta, R.L., eds. 2000. International conference on riparian ecology and 

management in multi-land use watersheds; 2000 August 28–31; Portland, OR. AWRA 2000 Summer 

Specialty Conference. Middleburg, VA: American Water Resources Association. 616 p.

2001a Faber, P.M., ed. 2003. California riparian systems: Processes and floodplain management, ecology, and 

restoration; 2001 March 12–15; Sacramento, CA. Sacramento, CA: Riparian Habitat Joint Venture. Mill 

Valley, CA: Pickleweed Press. 557 p.

2002e

UBC 2002. Symposium on small stream channels and their riparian zones: Their form, function, and 

ecological importance in a watershed context; Abstracts. 2002 February 18–20 [or 19–21]; Vancouver, 

B.C. The University of British Columbia (UBC), Vancouver, B.C., Canada. 38 p. (Limited number of papers 

published in the Canadian Journal of Forest Research. 33(8). Refer to Moore, R.D.; Richardson, J.S. 2003. 

Introduction: Progress towards understanding the structure, function, and ecological significance of small 

stream channels and their riparian zones. Canadian Journal of Forest Research. 33: 1349–1351.) (Some 

papers also published in Journal of the American Water Resources Association [JAWRA] in August 2005. 

Refer to Moore, R.D. 2005. Small stream channels and their riparian zones in forested catchments of the 

pacific northwest: Introduction. JAWRA August 2005: 759–761.)

2004e Dwire, K.A.; Lowrance, R.R. 2006. Riparian ecosystems and buffers—Multiscale structure, function, and 

management: Introduction. Journal of the American Water Resources Association (JAWRA).42(1): 1–4. 

February 2006. (AWRA Summer Specialty Conference. June 2004, Olympic Valley, CA.)

2005e

OHRC 2005. Science and management of headwater streams in the pacific northwest; 2005 November 

17–18; Corvallis, OR. Oregon Headwaters Research Cooperative (OHRC). (Refer to Danehy, R.J.; Ice, 

G.G. 2007. Introduction to special issue on headwater streams. Forest Science. 53(2): 101–103. Society 

of American Foresters. 

2007c RHJV 2007. Riparian habitat conservation and flood management in California; 2007 December 4–6; 

Sacramento, CA. Sacramento, CA: Riparian Habitat Joint Venture (RHJV). 116 p. http://www.prbo.org/

calpif/rhjvconference/proceedings/2007RHJVConfProceedings_complete.pdf

2007g

Richardson, J.S. 2007. Riparian management in headwater catchments: Translating science into 

management; 2007 February 19–21; University of British Columbia, Vancouver, BC, Canada. Refer to 

Maclsaac, E.A.: Moore, R.D.; Richardson, J.S. 2007. Riparian management in headwater catchments: 

Translating science into management—Meeting summary. Streamline Watershed Management Bulletin. 

11(1): 1–4.

2008a

Ryan, D.F.; Calhoun, J.M., tech. eds. 2010. Riparian adaptive management symposium: A conversation 

between scientists and management; 2008 November 3–4; Forks, WA. University of Washington, 

Olympic Natural Resources Center. Portland, OR: Gen. Tech. Rep. PNW-GTR-830. U.S. Department of 

Agriculture, Forest Service, Pacific Northwest Research Station. 135 p.

Table B.1—Continued.
226 USDA Forest Service RMRS-GTR-377. 2018

Table B.2—Organizations holding annual, semi-annual, or regular meetings pertaining to riparian ecosystems, riparian 

ecology, riparian habitat restoration, and/or riparian area conservation.

Start date Organizations including date formed and beginning and end of meetings 

1981 California Riparian Revegetation Study Group. (Organized circa 1981; Ron Schultze and Randy 

Gray.) Annual or semi-annual meetings in both northern and southern California (1981–1991). No 

proceedings. Function taken over by SERCAL Riparian Guild in 1992.

1983 Salmonid Restoration Federation (SRF). (Incorporated 1986.) Annual Salmonid Restoration 

Conference including riparian topics held in California since 1983. Proceedings for conferences 

since 2006 published on website.

1985 North American Riparian Council. (Organizational 1985; R. Roy Johnson, Founding President; 

disbanded by 1990.) No proceedings.

1986 Arizona Riparian Council. (Organized 1985–1986; Duncan Patten, Founding President.) Annual 

meetings held in Arizona (1986 to present.) No proceedings.

1989 Colorado Riparian Association (CRA). (Formed in 1989.) Sustaining Colorado Watersheds 

Conference (2006 to present). (See archives on website.)

1992 SERCAL Riparian Guild, California Society for Ecological Restoration. (Organized 1992: John 

Stanley.) Annual meetings (1992 to present). No proceedings; riparian topics often addressed at 

annual SERCAL conferences.

1995 New Mexico Riparian Council. (Incorporated 1995.)

1995 Great Basin Ecosystem Management Project (GBEMP): Restoring and Maintaining Riparian 

Ecosystem Integrity. (Organized 1995; Jeanne Chambers) U.S. Department of Agriculture, Forest 

Service, Rocky Mountain Research Station—Annual Symposium and Field Tour of riparian ecosystem 

research in Great Basin (1995 to present); renamed RiversEdge West (2018).

2001 Tamarisk Coalition. (Incorporated 2002 in Colorado) Annual riparian restoration research 

conferences and symposiums since 2001. Presentations recorded and made available online at 

Tamarisk Coalition’s YouTube page.

2005 Berkeley River Restoration Symposium. (Organized 2005; G. Matt Kondolf.) Annual symposium held 

at U.C. Berkeley (2005 to present).

2008 Texas Riparian Association. Annual meetings (2008 to present). Sponsor of Urban Riparian Symposia 

and Southwest Stream Restoration conferences (2013 to present). No proceedings.

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