Riparian Research and Management: 

Past, Present, Future Volume 2

United States Department of Agriculture

Forest 

Service

Rocky Mountain 

Research Station

General Technical Report June 2020

RMRS-GTR-411
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. 

Carothers, Steven W.; Johnson, R. Roy; Finch, Deborah M.; Kingsley, Kenneth J.; Hamre, Robert H., tech. eds. 2020. Riparian 

research and management: Past, present, future.Volume 2. Gen.Tech. Rep. RMRS-GTR-411 Fort Collins, CO: U.S. 

Department of Agriculture, Forest Service, Rocky Mountain Research Station. 268 p. https://doi.org/10.2737/RMRS-GTR-411

Abstract

In the Preface to volume 1, we discuss the development of riparian ecology as one of the newest of ecological fields that 

gained significant momentum in the 1950s and 1960s as part of the general “riparian movement” in the United States. The 

field expanded rapidly throughout the latter half of the 1900s. Volume 2 involves more than two dozen authors—most with 

decades of experience—who expand upon riparian and other topics introduced in volume 1. Two important recent 

developments are global climate change and impacts of introduced tamarisk leaf beetles (Diorhabda spp.) in the American 

West. Other chapters in volume 2 that provide current information evaluate the losses of riparian habitat, including 

“extirpation” of a large number of mesquite bosques (woodlands) in the Southwest; the restoration of riparian ecosystems 

damaged by anthropogenic activities; the importance of a watershed; and the importance of riparian ecosystems to 

recreation. The combination of volumes 1 and 2 examines the evolving understanding of scientific implications and 

anthropogenic threats to those ecosystems from Euro-American settlement of the region to present. 

__________________________

Keywords: riparian, ecosystem, ecology, riparian processes, riparian losses, restoration, aquatic, arid, semiarid, upland, 

freshwater, groundwater, hydrology, watershed, tamarisk, tamarisk leaf beetles (Diorhabda spp.)

Cover: A comparison of the effects of tamarisk leaf beetle defoliation. Top photo 2005, before leaf beetle 

defoliation. Photo by Steven W. Carothers. Bottom 2017, after defoliation. Photo by Robb Irwin Eidemiller. 

Colorado River, mile 7.5; Glen Canyon National Recreation Area, Arizona.
i

Technical Editors

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

Consultants, Phoenix, Arizona; Adjunct Professor, Department of Biological Sciences, Northern Arizona University, 

Flagstaff, Arizona.

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

Retired Research Scientist, National Park Service, Tucson, 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, National Park Service.

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

Station, Fort Collins, Colorado.

Contributing Authors to Volume 2

Gerard J. Allan, Professor of Biology and Director of the Environmental Genetics and Genomics Facility at Northern 

Arizona University, Flagstaff, Arizona. 

Randy K. Bangert, Adjunct Professor, Department of Biological Sciences, Northern Arizona University, Flagstaff, Arizona; 

School of Earth Sciences and Environmental Sustainability, Northern Arizona University, Flagstaff, Arizona. 

Davis E. Blasini, Ph.D. student in the School of Life Sciences at Arizona State University, Tempe, Arizona.

Helen M. Bothwell, Post-Doctoral Fellow working jointly with the Environmental Genetics and Genomics Center in the 

Department of Biological Sciences at Northern Arizona University, and with the USDA Forest Service, Rocky Mountain 

Research Station, Flagstaff, Arizona.

Abraham E. Cadmus, Ph.D. student and National Science Foundation graduate research fellow in the Department of 

Biological Sciences at Northern Arizona University, Flagstaff, Arizona. 

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

Consultants, Phoenix, Arizona; Adjunct Professor, Department of Biological Sciences, Northern Arizona University, 

Flagstaff, Arizona.

Tanner S. Carothers, Biologist, Springs Stewardship Institute, Museum of Northern Arizona, Flagstaff, Arizona.

Hillary F. Cooper, Ph.D. candidate in the Department of Biological Sciences at Northern Arizona University, Flagstaff, 

Arizona.

Christopher Estes, Aquatic Resources and Habitat Scientist at Chalk Board Enterprises, LLC, and Retired Fisheries 

Scientist and Chief of the Statewide Aquatic Resources Coordination Unit, Alaska Department of Fish and Game, 

Anchorage, Alaska.

Catherine A. Gehring, Lucking Family Professor in the Department of Biological Sciences and Merriam-Powell Center for 

Environmental Research at Northern Arizona University, Flagstaff, Arizona.

Kevin C. Grady, Research Assistant Professor in the School of Forestry and the Merriam-Powell Center for Environmental 

Research at Northern Arizona University, Flagstaff, Arizona.

F. Thomas Griggs, Practicing Restoration Ecologist with a long history of designing and implementing restoration projects 

in the Central Valley of California, recently retired from River Partners, Chico, California.
ii

Dorothy A. House, Technical Writer and Editor, SWCA Environmental Consultants, Flagstaff, Arizona. 

Julia B. Hull, Ph.D. candidate in the Department of Biological Sciences at Northern Arizona University, Flagstaff, Arizona.

Kevin R. Hultine, Research Ecologist in the Department of Research, Conservation and Collections at the Desert Botanical 

Garden in Phoenix, Arizona.

Dana H. Ikeda, Post-doctoral fellow with the Cottonwood Ecology group at Northern Arizona University, Flagstaff, Arizona.

Dale A. Jones, Retired Wildlife, Fish, and Endangered Species Director, USDA Forest Service, Washington, DC.

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, 

National Park Service.

Stefan Lorenzato, Chief, Riverine Ecosystem Section, Riverine Stewardship Program, California Department of Water 

Resources-IRWM, Sacramento, California. 

Lisa Markovchick, Ph.D. student and Presidential Fellow in the Department of Biological Sciences at Northern Arizona 

University, Flagstaff, Arizona.

Jackie Parker, Ph.D. student in the Department of Biological Sciences at Northern Arizona University, Flagstaff, Arizona. 

John P. Rieger, Restoration Ecologist; Retired Senior Biologist and Project Manager, California Department of Transportation; 

Founding President, Society for Ecological Restoration.

Stephen M. Shuster, Professor of Invertebrate Zoology and Curator of Marine Invertebrates and Mollusks with the Department 

of Biological Sciences at Northern Arizona University, Flagstaff, Arizona. 

Bo Shelby, Senior Social Scientist, Confluence Research and Consulting, and Emeritus Professor, Department of Forest 

Ecosystems and Society, Oregon State University, Corvallis, Oregon.

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

Ecologist, The Habitat Restoration Group; and Program Manager, Tahoe Regional Planning Agency; Paradise, California.

Lawrence E. Stevens, Curator of Ecology at the Museum of Northern Arizona, where he also serves as Director of the Springs 

Stewardship Institute.

Thomas G. Whitham, Regents’ Professor with the Merriam-Powell Center for Environmental Research and the Department of 

Biological Sciences at Northern Arizona University, Flagstaff, Arizona, and is the Principal Investigator for the National Science 

Foundation-funded Southwest Experimental Garden Array (SEGA). 

Doug Whittaker, Senior Researcher and Planner, Confluence Research and Consulting, formerly region-wide Outdoor 

Recreation Planner, National Park Service and Bureau of Land Management, Anchorage, Alaska.
iii

Foreword

I was honored when Roy Johnson and Steve Carothers asked me to write a Foreword for volume 2 of Riparian Research and 

Management: Past, Present, Future, but then realism hit me. When I read this book, I realized how much time, experience, and 

knowledge had passed me by. But, in my memories I could still envision the day in 1968 when Doug Morrison called me from 

the Flagstaff office of the Coconino National Forest to request funds for two “birders” to make a census of breeding birds along a 

section of the Verde River. The two “birders” were ornithologists Steve Carothers, Museum of Northern Arizona, and Roy Johnson, 

Prescott College. 

Salt River Project, a Phoenix-based water management company, had been cutting down mature cottonwood trees along the 

Verde River in central Arizona. Salt River Project was determined to rid Arizona of the “phreatophytes that were robbing the people 

of Arizona of their water,” but when they hit the Coconino National Forest boundary, they found out there were other resources 

to be considered. Our search for money to do the survey came up with a measly $528. I called Doug back knowing no biologist 

worth his salt would agree to such a meager sum! I don’t know if it was Doug’s pleading, or Steve’s and Roy’s knowledge that 

this inventory was far more important than money, but they agreed to make this survey. Steve and Roy found a large number of 

species of birds nesting in the cottonwood forests of the Verde Valley. Even more exciting, they also found the highest population 

densities of breeding birds ever reported for North America! These results changed the minds of a great number of people about 

the vital importance of riparian habitat, including mine. 

In 1977, I moved from my position as Wildlife, Fish, and Endangered Species Director for the Southwest Region of the United 

States Forest Service in Albuquerque, New Mexico, to the same position for the Chief of the Forest Service in Washington, DC 

Early that year, Roy Johnson, Steve Carothers, and I planned a riparian conference for the western United States that was held on 

July 9 in downtown Tucson, Arizona. We were unaware of concurrent planning for another riparian conference on “Riparian Forests 

in California,” held 2 months earlier, on May 14, 1977, at the University of California at Davis. The proceedings from those two 

conferences were the first riparian proceedings published for the United States. This current two volume publication is an example 

of the U.S. Forest Service’s continuing commitment to research in riparian ecosystems.

While I was reviewing volume 2, it was interesting to see that the 1970s were the years that riparian studies, conferences, and 

publications happened in many areas of the country. The symposia in California and Arizona may have influenced this, but a 

lot of the awareness just seemed to happen. The other shocking thing to me as I read volume 2 was that this manuscript was 

sometimes difficult to understand for biologists of my era. Such rapid advances have been made in riparian ecology since my 

retirement that I spent more time reading the dictionary than I did the manuscript of some chapters; but the great news is that 

today and tomorrow’s biologists will find volume 2 of Riparian Research and Management: Past, Present, Future extremely helpful.

Dale A. Jones, Retired

Wildlife, Fish, and Endangered Species Director 

USDA Forest Service

Washington, DC.
iv

Preface for Volume 2

This is the second of two volumes of scientific papers on semi-arid West and wet Pacific Northwest riparian ecosystems. 

Volume 1 entailed a primarily historical approach to the evolution of the science of riparian ecology and concerns and 

conservation needs for and on riparian habitats. It also reviewed past effects of anthropomorphic impacts to these highly 

productive ecosystems (chapters 1, 2, and 3). Another chapter discussed the turn-of-the-century invasion of the nonnative 

tree/shrub Tamarix spp. (tamarisk or salt cedar) and the past several decades of mostly failed attempts at its suppression and 

habitat restoration with native woody species (chapter 4). 

Following this was a detailed description of biocontrol efforts by the Department of Agriculture’s Animal and Plant Health 

Inspection Services (APHIS) on tens of thousands of acres where this exotic plant has invaded throughout the western United 

States over the past 100-plus years (chapter 5). The success of biocontrol, beginning in earnest in 2001, has surprised 

everyone, including program proponents with how rapidly introduced Asian tamarisk leaf beetles (Diorhabda spp.) have 

spread and defoliated their target. Results of biocontrol reveal a surprising management conundrum in the plight of an 

endangered songbird, the southwestern willow flycatcher (Empidonax traillii extimus), and its apparent reliance on nonnative 

Tamarix spp. for nesting habitat. Without this “invasive” tree species, there is little suitable nesting habitat remaining 

throughout much of the flycatcher’s range. 

As Tamarix declines under the influence of biocontrol, it has become obvious that original assumptions that native plant 

species would eventually replace the invader have not come to pass. Most areas, but not all, where Tamarix has been 

defoliated revegetate with a mixture of native and nonnative forbs and grasses, poor wildlife habitat replacement for the 

declining invader (chapter 4). Chapters 6 and 7 in volume 1 focused on beavers, their near-extirpation by early EuroAmericans, and how reintroduction of beavers can have many positive benefits on riparian health.

Volume 1 also provided a review of the extensive watershed alterations by dams and diversions in Arizona (chapter 8), 

reminding us that one of the more common and insidious impacts associated with dams is the inundation of riparian habitat 

behind the dam and change in the natural hydrograph, or seasonal patterns of flow, temperature, and sediment accumulation 

and distribution downstream. Chapter 9, using spatially extensive matched photo sets, provided the details of changes in 

riparian vegetation in canyon bound reaches of the Colorado River below these dams. The penultimate chapter, (chapter 10) 

discussed the breeding water birds of the Mexican portion of the Colorado River delta and the insidious effects that the last 

100 years of reduction in freshwater flows have had on delta habitat. 

The final chapter to volume 1 (chapter 11) documented the species richness and available data on population densities of 

land vertebrates in mesquite (Prosopis spp.) bosque habitats compared to other riparian habitat types. The appendices to 

volume 1 also provided a chronology of southwestern and northwestern United States riparian symposia and the riparian 

ecologists and other “pioneers” who made those symposia possible.

In volume 2, our first chapter explores conclusions of a widely distributed book on southwestern riparian systems, the Ribbon 

of Green (2007) by Robert Webb, Stanley Leake, and Raymond Turner. By analyzing changes in streams and rivers through 

time by photographic comparisons, this book reached a conclusion that more riparian habitat exists today than was present 

in the past. This chapter re-assesses riparian status, considering the proliferation of dams that have inundated thousands of 

miles of riparian habitat, diversions of surface water for agriculture, falling water tables, wide-spread invasion of nonnative 

species, and ever-increasing demands on water resources. Although we recognize the value of the book’s contributions on 

the status of southwestern streams and rivers in general, the results of our assessment reveal that the amount of riparian 

habitat present today has been appreciably reduced compared to earlier times, and we explain why our conclusion differs 

from Webb et al. (2007).

Throughout the remainder of volume 2, we apply a contemporary view of present-day riparian habitat by discussing the 

need to understand better the short- and long-term effects to wildlife habitat, riparian vegetation succession, and ecosystem 

recovery under the vagaries of climate change as well as remarkably successful biocontrol as was discussed in detail in 

chapter 5 of volume 1. The chapter discussing defoliation of Grand Canyon Colorado River corridor Tamarix and its potential 

influence on native vertebrates illuminates ongoing and ever-changing challenges to which native wildlife is exposed (chapter 

2). Chapter 3 continues the discussion from volume 1 on the disappearance of mesquite bosques and suggests habitat 

restoration where possible with mesquite as a replacement for Tamarix. One of the more creative approaches to riparian 

habitat maintenance and restoration is discussed in the chapter on the recent development of experimental gardens in arid 

lands and attempts to identify and deploy plant genotypes that are likely to survive vagaries of climate change. The combined 
v

effects of climate change undoubtedly will have an impact on our ability to maintain and restore riparian ecosystems, even to 

the level of defining specific ecotypes of plants that can be used in each specific location (chapter 4).

Additional volume 2 chapters detail and discuss complexities of the watershed continuum model and how that model may 

provide a better understanding and approach to habitat restoration (chapter 5). Chapter 6 describes a wealth of opportunities 

in the Anthropocene (the period during which human activity has been a dominant influence on the environment) for riparian 

restoration using agricultural runoff, unlined irrigation canals, waste water outfalls and orchards, that are often a part of 

human communities. Threats to riparian communities are not just limited to arid land systems, as are detailed in the two 

chapters describing unique challenges in California streams and rivers (chapters 7 and 8). 

The final two chapters in volume 2 discuss riparian ecosystem function where and how recreation and habitat restoration can 

be conflict-free. A compelling argument is made that as we engineer physical or biological habitat, we are also engineering 

recreation habitat, either by default or by design (chapters 9 and 10).

In both volumes of this publication, various threats to riparian ecosystems have been mentioned and discussed. From our 

viewpoints, threats are those processes, natural and anthropogenic, that disrupt or interfere with dynamic processes of 

native biotic and abiotic constituents functioning together in a healthy riparian ecosystem. Our concepts of threats to riparian 

habitats are value-based and biased by our professional and personal experiences as biologists and ecologists. With that 

viewpoint, it becomes obvious that there are many threats and most of these are interacting with each other simultaneously 

or sequentially. Our understanding of losses and gains in habitat through time are difficult to quantify because of different 

viewpoints and inadequate historical records. We know that things are not what they were in riparian areas before the coming 

of modern civilization, but accurate quantification of losses and threats is impossible, as is a complete return to those former 

conditions. Throughout both volumes, an emphasis on the natural hydrograph of healthy riparian systems is a common 

theme. The altered hydrograph usually results in a loss of native woody riparian vegetation that is often replaced by nonnative 

invasive species, e.g., Tamarix spp. or Elaeagnus angustifolia (Russian olive). 

It has become clear that without a natural flow regime, most native riparian vegetation cannot become reestablished as a 

viable community and, in its absence, wildlife productivity along our rivers and streams is seriously compromised. Invasive 

species are often pointed to as threats, but as is evidenced by the volume 1 chapter on unintended consequences of 

biocontrol and volume 2 chapters on Grand Canyon habitats, sometimes invasive species, especially Tamarix spp., can have 

multiple effects— “good” and “bad” in the context of riparian habitat and wildlife it supports. 

Collectively we are learning how to restore formerly disturbed riparian ecosystems. Many people, including most of those 

involved in the production of these two volumes, had various roles as “pioneers” in developing our current understanding of 

the functions and values of healthy native riparian areas. Our culture appears to have learned that riparian woodlands are not 

evil phreatophytes, needing control by removing all vegetation so that more water might be available for human usage. But 

these woodlands are valuable natural resources to be protected. Management orientation of riparian resources has shifted 

from control to protection and appreciation, throughout the Southwest, Pacific Northwest, and Mexico. Remnants of riparian 

areas are now being protected, and new riparian areas have been created, either intentionally or inadvertently, and then 

managed for conservation and aesthetic objectives and recreational uses. The collective consciousness of American culture 

appears to be moving toward a broader understanding of the role of riparian areas in the context of the watershed continuum 

and the world in which most of us want to live. 

Clearly, one of the greatest challenges riparian ecologists will face for the next several decades will be in evaluating impacts 

of Tamarix biocontrol and developing prescriptions for habitat restoration and recovery in areas where existing native wildlife 

resources have been significantly compromised due to unintended and unanticipated results of biocontrol. As documented in 

this volume, with few exceptions, passive revegetation with wildlife-enhancing native woody species does not occur postbiocontrol. Depending upon the quality of soil and water resources available, most areas where Tamarix has been defoliated 

by the tamarisk leaf beetle will likely require active, and expensive, revegetation efforts to achieve wildlife productivity 

approaching the levels that invasive Tamarix has heretofore accommodated. 

We, the authors and editors of these two volumes, hope that this work will serve as a foundation for continuing research, 

directed management, and societal appreciation of our western riparian resources.

Steven W. Carothers and R. Roy Johnson 

Co-Directors

The Western Riparian Project
vi

Acknowledgments

This is the second 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 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 ecological aspects of riparian habitat productivity, 

management, and conservation. Beginning in 2013, the Western Riparian Project is in its seventh 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 in the 1960s. These two volumes examine the progress made in western aridland riparian ecology, management, and 

conservation since those earliest studies. The idea for the beginnings of the project are outlined in the Acknowledgements of 

volume 1. 

Many of the authors of the chapters in this General Technical Report (GTR) have spent decades documenting the importance 

of conserving and managing riparian habitats due to their disproportionate value to wildlife and people compared to upland 

habitats. We greatly appreciate the efforts of our reviewers. Professionals that reviewed the entire volume 2 include Dale A. 

Jones, Patricia M. Woodruff, and Harley G. Shaw. Others that reviewed one or more chapters include Bertin W. Anderson, 

Kai Caraher, Bruce K. Orr, and Richard Valdez. We thank David L. Hawksworth from the Rocky Mountain Research Station for 

help in formatting the chapters. Special thanks to Lane Eskew, contracted editor for the Forest Service, who did a brilliant job 

combing through every word, number, figure, caption, table, etc. and made many suggestions that improved the product. 

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. Their dedicated efforts have greatly contributed to the success of this project. Finally, we thank Dale 

Jones for writing the insightful Foreword to this volume. Dale is a real riparian pioneer, one of the earliest to recognize the 

importance of riparian ecosystems and undertake steps for their conservation and management. 
vii

Table of Contents

Foreword . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iii

Preface for Volume 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iv

Chapter 1. Understanding Gains and Losses of Riparian Habitat: Interpreting Change, Its Causes 

and Consequences . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

Steven W. Carothers and Dorothy A. House

Chapter 2. A Naturalized Riparian Ecosystem: Consequences of Tamarisk Leaf Beetle 

(Diorhabda Spp.) Biocontrol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

Steven W. Carothers, R. Roy Johnson, and Kenneth J. Kingsley

Chapter 3. Vanishing Riparian Mesquite Bosques: Their Uniqueness and Recovery 

Potential . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

R. Roy Johnson, Tanner S. Carothers, and Steven W. Carothers

Chapter 4. Using the Southwest Experimental Garden Array to Enhance Riparian Restoration 

In Response to Global Environmental Change: Identifying and Deploying Genotypes and 

Populations For Current and Future Environments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63

Thomas G. Whitham, Catherine A. Gehring, Helen M. Bothwell, Hillary F. Cooper, Julia B. Hull, Gerard J. Allan, 

Kevin C. Grady, Lisa Markovchick, Stephen M. Shuster, Jackie Parker, Abraham E. Cadmus, Dana H. Ikeda, 

Randy K. Bangert, Kevin R. Hultine, and Davis E. Blasini

Chapter 5. The Watershed Continuum: A Conceptual Model of Fluvial-Riparian 

Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 80

Lawrence E. Stevens, R. Roy Johnson, and Christopher Estes

Chapter 6. It’s Not All Bad News—Riparian Areas in The Anthropocene . . . . . . . . . . . . . . . . . . 138

Kenneth J. Kingsley

Chapter 7. The Development of Riparian Ecosystem Restoration in California . . . . . . . . . . . . 154

John Stanley, F. Thomas Griggs, and John Rieger 

Chapter 8. Sacramento-San Joaquin System . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 211

F. Thomas Griggs and Stefan Lorenzato 

Chapter 9. Recreation Habitat vs. Ecological Habitat In Riparian Areas: Can We 

Manage For Both? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 233

Bo Shelby and Doug Whittaker

Chapter 10. Intended Versus Unintended Effects During Riparian Restoration Creating High 

Quality Recreation Habitat . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 250

Bo Shelby and Doug Whittaker
viii
USDA Forest Service RMRS-GTR-411. 2020. 1

Chapter 1. Understanding Gains and Losses of 

Riparian Habitat: Interpreting Change, Its Causes and 

Consequences

Steven W. Carothers and Dorothy A. House

Introduction

One of the most wide-ranging scientific treatments of riparian vegetation of the 

American Southwest (exclusive of New Mexico and southwestern Colorado) can be found in 

the 2007 book, The Ribbon of Green by Robert Webb, Stanley Leake, and Raymond Turner 

(Webb et al. hereafter). In The Ribbon of Green, the authors use sets of matched photographs 

to evaluate and interpret historical changes in riparian vegetation along southwestern streams 

that have taken place since the late 19th century. As stated in the book’s preface, the authors 

“concentrate on the question of long-term changes, addressing whether we have endured a 

net loss of woody riparian vegetation” over that time (Webb et al. 2007:x). 

More narrowly, the authors challenge the accuracy of a widely-repeated claim that 90 

percent (or a similarly large percentage) of riparian vegetation in the southwestern United 

States has been lost due to human activities. In The Ribbon of Green, Webb et al. either 

explicitly or implicitly conclude that: (1) the Southwest has not endured a net loss of 

woody riparian vegetation over time; (2) claims of such loss have been exaggerated and 

misguided; (3) the 90 percent loss figure is myth; (4) riparian vegetation has changed to 

some degree, but change is to be expected in a dynamic ecological system and is neither 

bad nor good; and (5) the role of human actions in bringing about change in riparian 

systems has been overblown relative to the influences of natural processes. In this chapter, 

we examine the validity of those conclusions and the methodology that led to them, and 

offer further information and interpretation. 

For their study, Webb et al. compared 2,724 sets of repeat photography with native 

and nonnative woody riparian vegetation visible in at least one of the matched photos. Of 

the 2,724 sets examined, 215 sets totaling 480 individual photographs are published in 

the book. Lengths of time spanned by the published sets vary widely, from a few recent 

decades (e.g., 1988 compared to 2000) to more than a century (e.g., 1863 compared to 

2000). The book is more than just a comparison of photographs, however. Webb et al. also 

summarized the biology and distribution of riparian ecosystems in the study area, drawing 

upon numerous publications and research projects completed prior to 2007. They look at 

human influences, climatic fluctuations, hydrologic changes, and the life history strategies 

of dominant woody vegetation species in the selected riparian ecosystems.

The Ribbon of Green is a remarkable reference book for many things riparian in the 

geographic areas they address. That said, we believe Webb et al. are doing a disservice 

to the very resource they celebrate by painting too sanguine a picture of the history and 

status of southwestern riparian ecosystems. The scientists who initially raised concerns 

about the historical trajectory, condition, and future of riparian habitats were leaders who 

drew attention to conservation needs for these ecosystems. In their preface, Webb et al. 

dispute the concerns of those scientists and generally understate the level of threat to riparian 

vegetation that existed in the mid-20th century. At its core, the book is revisionist, an attempt 
2 USDA Forest Service RMRS-GTR-411. 2020.

to correct a record that, in the authors’ view, had been distorted. But in their enthusiasm 

to counter often-exaggerated estimates of historical riparian losses and their interest in 

promoting the effects of climate and other natural processes on riparian systems as opposed 

to human impacts, the authors risk tipping the scales too far in the opposite direction. 

There is danger in understating human impacts on southwestern riparian habitats. The 

risk is a desensitization of policy makers, resource managers, and the public to existing 

and future threats. Another risk is a weakening of the political will to avoid, minimize, 

and mitigate such threats. Vigilance is particularly important now because southwestern 

riparian habitats are undergoing highly visible changes resulting from release of the 

tamarisk leaf beetle (Diorhabda spp.), a biocontrol effort that was in its infancy when 

Webb et al. wrote The Ribbon of Green. The long-term ecological consequences of the 

beetle release are unknown and much debated, but in the short term it is clear that the 

amount of woody riparian vegetation available to wildlife is shrinking as nonnative, 

invasive tamarisk (Tamarix spp.) defoliates and dies across the western United States (see 

Carothers et al., this volume; McLeod volume 1). 

Have We Endured a Net Loss of Woody Riparian Vegetation?

Webb et al. pose the question of riparian losses in the book’s preface, and then they 

answer it in the book’s final two chapters with the following statements:

“In general, riparian vegetation has had either an [a small] increase (49 

percent [of the photographs]) or a large increase (24 percent [of the photographs]) 

in comparisons involving all years ....” (Webb et al. 2007:388)

“The time series of change ... suggests that, at least in the specific reaches 

for which we have documentation, riparian vegetation began to increase after 

about 1940. Increases in density of plants appear to have accelerated after the 

1970s, followed by increases in the size of plants. Many deviations from this 

general pattern have been noted in the chapters on specific reaches.” (Webb et al. 

2007:388)

“Our interpretations of the available imagery ... suggest more gains than 

losses within our photographic views.” (Webb et al. 2007:403)

“Although we observed local decreases in riparian vegetation in many places, 

reach-scale decreases in the 20th century were present only in a few locations.” 

(Webb et al. 2007:405)

In these statements, the authors couch their findings in terms of the specific locations 

photographed, or, at most, the river reach in which the photographs were taken. These 

geographic constraints are appropriate considering the spatial limitations of repeat 

photography, which we discuss at length below. However, a stated purpose of the study 

is to address allegations of riparian habitat loss across a very large area, the southwestern 

United States. Given this regional premise, readers can be forgiven if they understand the 

scope of authors’ ultimate conclusions to be region-wide as well. The book’s unmistakable 

message is we have not endured a net loss of woody riparian vegetation in the American 

Southwest despite more than a century of intensive agricultural, urban, and water resource 

development. 

In the book’s preface and accompanying notes (page 413), Webb et al. place the 

responsibility for originating the “perception of diminishing or degraded riparian areas” 
USDA Forest Service RMRS-GTR-411. 2020. 3

with academic and agency scientists who, for the most part, began publishing their 

concerns in the late 1970s. Specifically, Webb et al. report the claim of a 90 percent loss as 

having been traced back to a single paper (i.e., Ohmart et al. 1977). Webb et al. imply that 

the concerned scientists—many of whom pioneered the discipline of riparian ecology—

exaggerated the extent of riparian losses right from the beginning. How does one account 

for the conflicting views of Webb et al. and those they disagree with? 

Several explanations are possible. First, Webb et al. and the early concerned scientists 

came to the issue from different professional perspectives, and those perspectives carried 

with them different priorities and biases that likely colored their conclusions. Second, 

Webb et al. and the early concerned scientists defined the issue somewhat differently. 

The scientists who first raised the alarm were concerned about the loss of native riparian 

vegetation, not all riparian vegetation, and they were concerned about the reduction of 

wildlife habitat quality, particularly that posed by the proliferation of nonnative plant 

species. Webb et al., although they differentiate between native and nonnative species, 

concentrate on changes in the total amount of riparian vegetation. They pay little attention 

to habitat quality. 

Third, Webb et al. and the concerned scientists arrived at their conclusions via 

different pathways, and their divergent approaches influenced their conclusions. 

Fourth, Webb et al. and the concerned scientists operated in different timeframes. The 

circumstances affecting riparian areas in the early 2000s, when The Ribbon of Green was 

written, differed in many ways from those in the 1970s, when the alarm was first raised. 

Moreover, Webb et al.’s temporal scope was necessarily limited to the periods when 

photographs were taken, while the concerned scientists were troubled by changes that 

have taken place since first European settlement. We discuss these points in the following 

sections and revisit the question of where that 90 percent loss figure originated. 

Different Perspectives, Value Judgements, and the Issue of Habitat Quality

Robert Webb and Stanley Leake are hydrologists/geomorphologists, and their 

interest and expertise in physical processes is evident in the attention they pay to climatic 

variation, streamflow, flood records, and stream channel morphology. Raymond Turner, a 

plant biologist, has long favored climate shifts as the primary driver of vegetation changes 

in the Southwest (see Bahre 1991; Hastings and Turner 1965). In The Ribbon of Green,

a focus on physical processes as opposed to human influences (as well as a conviction 

that relatively little riparian vegetation has been lost) is revealed in this statement: “That 

some patches of vegetation, growing as they do in a dynamic geomorphic milieu, are 

gone is to be expected” (Webb et al. 2007:ix). This is not to say that Webb et al. do not 

acknowledge human impacts—they do—but they place decidedly more emphasis on 

climatic fluctuations, hydrology, and geomorphological changes to explain the differences 

in riparian vegetation between compared photographs. 

When describing the vegetation visible in their photographs, Webb et al. make no 

overt value judgments about what they see. From their perspective, differences observed 

between compared photographs reflect changes that are neither desirable nor undesirable, 

neither good nor bad. In their preface, Webb et al. decry the fact that some scientists do

make judgments about the rightness or wrongness of changes they perceive. The authors 

criticize the “... value judgments and assumptions that creep into the issue of long-term 
4 USDA Forest Service RMRS-GTR-411. 2020.

change in the region’s rivers. Some authors use pejorative words such as destroyed

to describe what has happened to these watercourses historically and lost to describe 

changes in riparian vegetation.... These judgments derive from some scientists’ perception 

that humans are the ultimate cause of regional change and therefore that any changes must 

be ‘bad’” (Webb et al. 2007:ix)

Webb et al. are correct that value judgments have been made by others, and words 

such as “destroyed” and “lost” have been (and are) commonly used, but the authors’ 

presumption that fellow scientists believe “humans are the ultimate cause of regional 

change and therefore that any changes must be ‘bad’” is unsupportable and unmerited. 

One need not look far in the literature to find a different and more obvious motive 

among the scientists who launched the movement to conserve native riparian systems. 

Those scientists were generally ecologically-oriented ornithologists who saw riparian 

areas primarily as habitat for wildlife, particularly birds (e.g., Anderson and Ohmart 

1976; Hehnke and Stone 1979; Hubbard 1971; Johnson and Simpson 1971; Knopf 

1985; Phillips and Monson 1964; Rea 1983; Schmitt 1976; Stevens et al. 1977). Their 

orientation was (and is) to view riparian systems primarily through the lens of services 

provided, not to humans directly, but to something humans (biologists in particular) care 

about: the health and functioning of ecological communities. 

From that perspective, a reduction in either the amount of riparian habitat or a 

diminishment in the quality of that habitat—as measured in species richness, diversity, 

and density—is undesirable. That is a commonly held view whether the undesirable 

condition is caused by human activity or not. Value judgments did not creep into the 

assessments of the concerned scientists; value judgements were an overt and legitimate 

aspect of their assessments. 

Different Pathways—Repeat Photography

Webb et al. approach the issue of riparian change through a methodical study using 

compared photographs. An implicit premise of the study is that the collection of historical 

photographs available to the authors reliably documents conditions of riparian vegetation 

across the Southwest at the time of Euro-American settlement (any effects of earlier 

Native American, Spanish, and Mexican activities are necessarily moot because such 

activities predated photography). We believe this premise and other aspects of the study 

are open to question and explain why below. 

First, while a comparative photographic record provides valuable visual 

documentation of two points in time for a specific location, such a record has limited 

application and utility for extrapolation over a large area. Constrained by nonrandom 

snapshots that are rigidly confined both spatially and temporally, repeat photography is 

not well suited to provide a systemwide assessment of riparian condition and change over 

time. Achieving adequate geographic coverage is difficult if not impossible for such an 

assessment. On page 404 of The Ribbon of Green, the authors state, “In this book, we 

used thousands of repeat photographs to address the claim that 90 percent of riparian 

vegetation has been lost in the southwestern United States” (Webb et al. 2007:404). On 

the same page, however, they themselves cast doubt about the appropriateness of their 

approach by acknowledging they could not fully meet this objective for reasons that 

include lack of “full regional coverage.” 
USDA Forest Service RMRS-GTR-411. 2020. 5

The acknowledgment also alludes to an important weakness of the study and the 

conclusions based on it. The regional coverage falls far short of “full.” While 2,724 sets 

of photographs may seem to be a great many, the actual amount of riparian environment 

documented in those photographs barely signifies when compared to the tens of 

thousands of miles of rivers, streams, and washes in the American Southwest. In Arizona 

alone, there are approximately 14,605 miles of perennial and intermittent waterways 

and another 112,900 miles of ephemeral washes (Arizona Department of Environmental 

Quality 2011). Coverage of the rivers and streams included in Webb et al.’s study is 

sparse and spotty, and many waterways (most notably the Rio Grande and its tributaries) 

are not represented at all. 

Moreover, the images used in the study suffer from the spatial limitation common to 

all photography. That is, the area depicted in each image is only as large as the camera’s 

field of view. Each photograph in this study represents only one narrowly circumscribed 

point in a reach of stream that may extend scores or hundreds of miles. A photograph 

cannot speak to riparian conditions that exist a few yards, let alone miles, upstream or 

downstream of the camera station. 

The study also suffers from what practitioners of repeat photography call “spatial 

bias” (Villarreal et al. 2013). The most obvious example of this is the fact that 46 percent 

(1,251 of 2,724) of the matched photographs used in the study were taken in a single river 

reach: the Colorado River through Grand Canyon. That overrepresentation biases the 

result toward circumstances that are not characteristic of southwestern rivers and streams 

in general. Grand Canyon is a severely canyon-confined reach, while most southwestern 

waterways are more alluvial and open in nature. Spatial bias is an inherent weakness of a 

repeat photography project that uses pre-existing baseline imagery (Bahre 1991; Pickard 

2002; Pupo-Correia et al. 2014; Villarreal et al. 2013). The camera stations in such studies 

are not purposefully and randomly selected to achieve a spatially unbiased sample. 

Rather, they are accidents of history: functions of photographs that happen to have been 

taken years earlier, usually (as in this case) for reasons unrelated to the question at hand. 

In the Webb et al. study, just two river reaches (Grand Canyon and Cataract Canyon) of 

the over 37 reaches covered account for over half (55 percent) of all matched images. 

The historical photographs in those matches happen to exist because photographers 

accompanied two expeditions in the 19th century. 

Other locales that appear to be overrepresented (based on the published images) include 

stream gaging stations and readily accessible sites near bridges and communities. Types 

of environments that appear to be underrepresented in those images include river reaches 

and associated riparian vegetation lost or altered by reservoirs, urban development, and 

bottomland agriculture. For example, four Colorado River reservoirs (Lakes Powell, Mead, 

Mohave, and Havasu) that collectively inundated approximately 400 mainstem river miles 

are represented in the book by only seven matched pairs of photographs. 

In the book’s preface, Webb et al. use a term implicative of a second common 

problem of repeat photography: limited temporal coverage. The authors write: “... 

[repeat photography] is the only method that provides a glimpse [our emphasis] into 

predevelopment conditions in our region, a glimpse that is required to address the 

question of gains or losses of riparian vegetation” (Webb et al. 2007:iv). “Glimpse” 

implies a fleeting moment, and a moment in time is all each photograph in the study 

captures. The validity of the study, however, rests on the premise that each image 
6 USDA Forest Service RMRS-GTR-411. 2020.

represents the usual conditions at that location in the “past” compared to the usual 

conditions in the “present.” This assumption is difficult to defend for environments as 

dynamic as southwestern riparian systems, systems that are subject to unpredictable 

scouring floods and characterized by fast-growing plant species. The vegetation 

incidentally captured in a photograph may be very different from the vegetation that 

would have been shown if the photograph had been taken at the same location a year, a 

month, or even a week earlier or later. 

The problem of temporal coverage is aggravated by temporal inconsistency. Webb et al. 

maintain that their “repeat photography establishes the visual condition of woody riparian 

vegetation at the end of the 19th century and the beginning of the 20th century” (Webb et 

al. 2007:403). However, this narrowly defined baseline does not exist. There is no common 

historical baseline of any kind. Far from it. Webb et al. report that the earliest dates of the 

initial photographs used to assess change since “predevelopment” range from 1863 (for the 

lower Colorado and Mohave rivers) to 1940 (for the Agua Fria River). The oldest images for 

all other stream reaches fall somewhere between those 2 years (Webb et al. 2007:44). 

Using dates ranging as far apart as 1863 and 1940 to represent riparian conditions 

“predevelopment” is questionable, but the temporal inconsistency is even greater than that. 

In individual matched sets reproduced in the book, dates of the older or oldest photographs 

range well into the 1980s. Overall, the “baseline” dates number in the dozens and are 

scattered over more than a century. Dates of the “new” photographs in matched sets, those 

documenting recent conditions, generally range from the late 1990s to the early 2000s. 

Thus, the time span of individual matched sets is highly variable, ranging from over 130 

years (e.g., 1863 compared to 2000) to under 20 years (e.g., 1988 compared to 2000).

The relatively recent dates of many baseline photographs point to yet another problem 

with the study. Webb et al. refer to it when they admit that “for many watercourses ... the 

initial photography postdates a strong human influence on the watershed” (Webb et al. 

2007:403). Certainly, assessing the role of human activity in what is essentially a “before 

and after” exercise is hindered when the baseline (i.e., “before”) photographs were taken 

after—sometimes long after—human activities (e.g., cattle grazing, beaver trapping, 

damming) commenced in the Southwest. 

Different Pathways – Different Times

The initial concern over historical losses of riparian habitat in the Southwest and the 

realization that what was left was seriously threatened took place amidst a nationwide 

groundswell of environmental awareness beginning in the 1960s and 1970s. Ecology was a 

young, rapidly expanding science, and the role of humans as an ecological factor was gaining 

recognition. In the Southwest, ecologically oriented ornithologists were among the first to 

recognize a link between riparian habitat value, human activities, reduction in available 

riparian habitat, and species decline. They began to see a decrease in riparian avian species 

richness and abundance about the same time they were discovering the disproportionate 

importance of riparian habitat for birds (see, e.g., Johnson et al. 1977; Rea 1977). 

Biologists assessed avian population status and trajectories in riparian areas by 

comparing the results of recent bird surveys with historical accounts and distribution lists 

(e.g., Stephens 1885; Swarth 1914). Following one of these assessments, Hunter et al. 

(1987) reported that, since the year 1900, 17 riparian bird species had either declined or 
USDA Forest Service RMRS-GTR-411. 2020. 7

been extirpated from one or more of 16 southwestern river reaches. The lower Colorado, 

lower Gila, lower Salt, and lower Santa Cruz Rivers were hardest hit, with 12 to 13 

species lost or declining in each system. These were the reaches most altered by dams, 

diversions, channelization, bottomland agriculture, urban and industrial development, 

and/or lowered water tables due to heavy groundwater withdrawals. In contrast, the 

river reaches upstream of dams and in less developed areas (e.g., the Upper Gila, Upper 

Salt, Upper Santa Cruz, and San Pedro Rivers) experienced the least change in riparian 

avifauna, with 0–2 species lost or declining in each system Hunter et al. (1987).

The fact that riparian bird species declined concurrently with human modification of 

river systems was not lost on observers. A reduction in the availability of suitable riparian 

habitat due to human activities was the logical inference, an inference supported by 

observed deviations from historical descriptions of vegetation found in written accounts 

from the 18th through the early 20th centuries (see, for example, Bartlett 1854; Bell 1869; 

Bolton 1919; Emory 1857; Froebel 1859; Johnston 1848; Mearns 1907; Rusling 1875; 

Swarth 1905; Willard 1912). 

By the middle of the 20th century, the stands of riparian trees—mostly cottonwood 

(Populus spp.), willow (Salix spp.), and mesquite (Prosopis spp.)—described in those 

accounts had often dwindled or disappeared (Bahre 1991; Dobyns 1981; Rea 1983; 

Webb et al. 2014). The reduction in the amount of streamside vegetation was strikingly 

apparent and troubling because of the rarity and relative importance of that vegetation 

in the Southwest compared to other regions. In the Southwest, riparian habitats make 

up a very small fraction of the generally arid and semiarid landscapes and are naturally 

more disjunct than those of wetter climes. At best, they are connected only along narrow 

ribbon-like bands and usually separated by vast expanses devoid of similar vegetation 

(Skagen et al. 1998). 

 The most easily attributable (although not the earliest) human-caused changes in 

southwestern riparian vegetation resulted from the dams and diversion structures that 

were built on nearly every river in the Southwest following passage of the Federal 

Reclamation Act of 1902. That Act created the Bureau of Reclamation and authorized 

construction of Roosevelt Dam, the first of seven dams built on Arizona’s Salt River 

and its largest tributary, the Verde River (see table 1). Combined, the seven dams and 

associated diversions choked off all but flood flow in the once perennial lower Salt River. 

By the end of the 1920s, loss of discharge from the Salt River (the Gila River’s largest 

tributary), combined with damming of the Gila mainstem, converted the lower Gila River 

into a mostly dry channel (Huckleberry 1996; Werner volume 1). The Agua Fria River, 

another Gila River tributary in Arizona, suffered a similar fate. Altogether, the dewatered 

portions of just these three rivers totaled more than 340 miles (see table 1). 

The dewatering of the Salt, Agua Fria, and Gila River channels had a particularly 

egregious effect on downstream riparian systems, but only a few dams in the Southwest 

impeded flow to that degree. Every dam, however, altered downstream flow in some 

fashion, typically by disrupting the natural hydrograph, modulating and decreasing 

downstream flow, and reducing the magnitude of peak flows. Downstream sediment loads 

were diminished, stream channel and floodplain morphology altered, and water tables 

lowered. These modifications affected the persistence and recruitment of native riparian 

vegetation downstream of dams. Upstream of dams, impounded water drowned whatever 

riparian vegetation once grew along the flooded mainstem channel (over 250 miles on the 
8 USDA Forest Service RMRS-GTR-411. 2020.

lower Colorado River alone) as well as along flooded tributary channels (e.g., the Virgin 

River arm of Lake Mead and the San Juan arm of Lake Powell) (see table 1). 

The impacts of dams and diversions on riparian vegetation were perhaps the most 

obvious when concern about riparian habitat loss and impairment was first raised, but 

those impacts were not the only ones. Large swaths of bottomlands capable of supporting 

native vegetation had been converted to agriculture and other human uses. No regionwide estimate of the magnitude of this conversion is available, but local studies have been 

done. For example, Ohmart et al. (1988) quantified the shift to agriculture for a 21,504-ha, 

71-mile-long stretch of the lower Colorado River. Between 1938 and 1986, the proportion 

of that area covered by native riparian vegetation decreased from 88.6 to 19.8 percent, while 

crops increased from 0.3 to 61.1 percent. Even greater losses of native riparian vegetation to 

agriculture (ca. 95 percent) were calculated along the lower Gila River (Lacey et al. 1975). 

Riparian displacements of similar magnitude were thought to be typical of fertile, easily 

irrigated bottomlands throughout the Southwest (Ohmart et al. 1988). 

Additional reductions in native riparian vegetation were reported due to urbanization, 

groundwater pumping, bank stabilization, channelization, levee construction, and other 

activities (National Research Council 2002; Poff et al. 2011). Throughout the 19th and 

early 20th centuries, cattle grazing and agriculture contributed to erosion and loss of 

riparian integrity in many waterways of the Southwest, including small streams and arroyos 

(Dobyns 1981). Even woodcutting took a toll. As early as 1910, for example, it was noted 

that woodcutting for fuel had resulted in the “practical disappearance” of mesquite along 

much of the lower course of the Colorado River (Grinnell 1914), an area where 50 years 

Source: R. Roy Johnson

a

 Reaches of rivers below dams into which controlled releases are made for irrigation, domestic and industrial water, and power production.

b Total miles of stream channel that flow only during flood stage and that have been dewatered and desertified, losing nearly all their riparian 

habitat.

c

 Total includes previous three categories in table.

d

 From Glen Canyon Dam to U.S.-Mexico boundary below Yuma. 

e Includes only reservoirs behind Hoover, Davis, Parker, and Headgate Rock Dams (i.e., Lakes Mead, Mohave, Havasu, and Moovalya, 

respectively), four of the nine major storage and diversion dams on the Lower Colorado River. 

f

 Includes Coolidge Dam, San Carlos Reservoir; Ashurst-Hayden Diversion Dam.

g Includes Roosevelt Dam, Roosevelt Lake; Horse Mesa Dam, Apache Lake; Mormon Flat Dam, Canyon Lake; Stewart Mountain Dam, Saguaro 

Lake; Granite Reef Diversion Dam. 

h

 Includes Horseshoe Dam, Horseshoe Reservoir; Bartlett Dam, Bartlett Reservoir; Granite Reef Diversion Dam at confluence with the Salt River.

Table 1—Approximate linear miles of selected rivers impacted by dam construction in Arizona.

River

Total miles in 

Arizona

Inundated channel 

upstream

Regulated flow 

downstreama

Dewatered channel 

downstreamb

Total miles (%) not 

free-flowingc

Coloradod 704 250+e 454 0 704 (100%)

Gilaf 440 20 55 265 340 (77.3%)

Saltg 200 40 11 38 89 (44.5%)

Verdeh 170 20 20 0 40 (23.5%)

Agua Fria 120 12 0 38 50 (41.7%)

Total 1,634 342+ 540 341 1,223 (74.7%)
USDA Forest Service RMRS-GTR-411. 2020. 9

earlier observers reported large, at times impenetrable, mesquite woodlands (Bartlett 1854; 

Bell 1869; Emory 1857). 

A huge mesquite bosque on the Salt River estimated to cover 25 square miles had been 

replaced by metropolitan Phoenix and surrounding fields and orchards (Jacobs and Rice 

2002). Additional large, well-documented bosques that had been decimated or obliterated 

by the mid-1900s included the San Xavier Thicket (or Great Mesquite Forest; Webb et al. 

2014) on the middle Santa Cruz River; the Komatke Thicket (or New York Thicket; Rea 

1983) at the confluence of the Santa Cruz and Gila Rivers; and the Casa Grande National 

Monument bosque just south of the Gila River (Johnson et al., this volume; Judd et al. 

1971). These bosques succumbed to a combination of activities, including wood cutting, 

clearing for agriculture, upstream dam construction, and finally groundwater overdraft and 

cessation of flowing surface water (Phillips et al. 1964; Rea 1983; Webb et al. 2014).

In the mid- and late 1900s, scientists and others certainly lamented the riparian habitat 

that had been lost and degraded up to that time, but it was ongoing and accelerating threats 

that incited a call for action. Demands for water in the arid Southwest were growing along 

with the human population and economic development, and those demands were being met 

by ever more ambitious water development and storage projects. The projects authorized 

by the U.S. Bureau of Reclamation in the Colorado River basin during that period included 

eight1

 new large dams, numerous smaller dams, and a score of large water diversion 

projects, including the massive Central Arizona Project (Bureau of Reclamation 2018a,b). 

Unprecedented threats to the integrity of riparian habitat included the rampant spread 

of nonnative tamarisk along southwestern drainages and an intentional assault on riparian 

vegetation called “phreatophyte control.” The spread of tamarisk was troubling for biologists 

largely because, in the early days of riparian study, tamarisk was almost universally believed 

to out-compete native vegetation and replace it with inferior wildlife habitat (Bateman et 

al. 2013; Haase 1972; Harris 1966; Rosenberg et al. 1991; Sogge et al. 2013). Phreatophyte 

control was a coordinated strategy of land and water managing agencies, water suppliers, 

and industrial water users to increase water yield for beneficial (i.e., human) use by removing 

deep-rooted, high-water-consumptive plants (phreatophytes) from along streams and rivers. 

While tamarisk was the principal focus of eradication efforts, native species were targeted as 

well. Woody riparian plants would be replaced, if possible, with low-water-consumptive and/

or economically useful vegetation (e.g., grass forage for livestock).

In Arizona, phreatophyte control was first field-tested in the 1940s as a means to 

provide water needed to expand wartime copper production at the Phelps Dodge Morenci 

Mine (Gatewood et al. 1950). By 1969, thousands of acres of riparian vegetation had been 

treated and many more projects were planned. In that year, the Pacific Southwest InterAgency Phreatophyte Subcommittee conducted a survey of phreatophyte control projects 

in Arizona. The results of the survey were reported by Affleck (1975) and are summarized 

here in table 2. A total of 11,909 acres of riparian vegetation had been treated as of the date 

of the survey; 20,415 acres of treatment were underway; and 80,191 acres of treatment were 

planned, for a total of 112,515 acres (table 2). 

1 Three of the big dams were never built. Marble Canyon Dam and Bridge Canyon Dam (aka Hualapai Dam), both planned for the 

Grand Canyon, were removed from consideration in 1968. Hooker Dam, slated for the Upper Gila River, remained a possibility 

into the 1980s. Serious interest in building an alternative to Hooker Dam for water storage on the Upper Gila was revived with 

passage of the Arizona Water Settlements Act of 2004 (P.L. 108–451), but nothing has materialized as of this writing (Pearson 

1994; Utton Transboundary Resources Center 2014).
10 USDA Forest Service RMRS-GTR-411. 2020.

Of that total, 11 percent of the vegetation was native (cottonwood, willow, mesquite); 

14 percent was tamarisk; and 75 percent was mixed tamarisk and native. At the time, 

estimates of the total coverage of riparian vegetation in Arizona outside of National Parks 

and other protected areas ranged from 279,600 acres (Babcock 1968) to 320,000 acres 

(Lower Colorado Region Comprehensive Framework Study, Appendix VI 1971). Based 

on these estimates, it would have been reasonable to conclude that between 35.2 and 40.2 

percent of the remaining unprotected riparian habitat in the State was threatened with 

imminent destruction.

In short, in the mid-20th century, the perception of loss and decline of native 

southwestern riparian habitat was well justified, and the threat of substantial additional 

harm was acute. By the time Webb et al. launched their study in the early 21st century, 

the context was quite different. The ecological importance of riparian systems had 

become universally recognized, and some of the threats that loomed largest in the mid20th century had diminished. Public and Federal enthusiasm for new dam construction 

had dwindled; intentional destruction of native riparian vegetation for water salvage 

had become socially and politically unacceptable; conservation of riparian habitats had 

become the norm in natural resource management; and several restoration projects were 

planned and underway. This is not to say old threats had not persisted (e.g., escalating 

water demand associated with human population growth), or that new threats had not 

surfaced (e.g., predicted increases in drought occurrence and severity), but by the turn of 

the 21st century, riparian habitats were widely valued and had powerful advocates. 

So ... Have We Endured a 90 Percent Loss of Woody Riparian Vegetation in 

the Southwest or Not?

Answering this question has been hampered by the lack of a credible baseline from 

which to measure change. To our knowledge, no one has attempted to estimate the total 

amount of woody riparian vegetation that existed throughout the Southwest before 

European settlement, and no one has attempted to quantify the amount of change that has 

taken place since that time. What researchers have done is estimate the extent of presettlement riparian vegetation and historical change for a few individual river reaches. 

One of these estimates, an effort by Ohmart et al. (1977) to quantify the loss of 

cottonwood-willow in the Lower Colorado River Valley, is of special interest. It is this 

publication that Webb et al. identifies as the ultimate “scientific justification” for the 

statement, “We’ve lost 90 percent of the riparian vegetation in the southwestern United 

Source: Calculated from data provided by Affleck (1975) in table 4.

Table 2—Acres of phreatophyte treatments completed, underway, and planned in Arizona according to a survey conducted by the 

Pacific Southwest Inter-Agency Phreatophyte Subcommittee in 1969.

Phreatophyte type Treatment completed Treatment underway Treatment planned Total acres (%)

Native 4,657 7,710 0 12,367 (11)

Tamarisk 5,894 7,455 2,711 16,060 (14)

Mixed native and tamarisk 1,358 5,250 77,480 84,088 (75)

TOTAL 11,909 20,415 80,191 112,515 (100)
USDA Forest Service RMRS-GTR-411. 2020. 11

States” (Webb et al. 2007:ix, endnote 3). To be fair to Ohmart et al. (1977), they made no 

such claim, nor did they present information that justified such a claim. 

To begin with, although Webb et al. placed quotation marks around the statement, 

“We’ve lost 90 percent of the riparian vegetation in the southwestern United States,” 

they gave no attribution. Rather than an actual quotation, this appears to be a generic 

representation of a type of claim seen in the literature. The connection Webb et al. make 

between this generic “statement” and Ohmart et al. (1977) is through a chain of citations 

associated with an actual statement made in support of an executive order issued by 

Arizona Governor Rose Mofford for the protection of riparian areas (Governor’s Riparian 

Habitat Task Force 1990). The actual statement reads as follows: “According to most 

estimates, over 90 percent of the native riparian areas along Arizona’s major desert 

watercourses have been lost, altered or degraded as a result of man’s activities (SCORP 

1989; TANC 1987; Warner 1979)”2

. The actual statement varies significantly from Webb 

et al.’s generic statement. Note that the actual statement refers: (1) only to “native riparian 

areas,” not to all riparian areas; (2) only to “Arizona’s major desert watercourses,” not 

to the entire “southwestern United States”; and (3) to “lost, altered or degraded” riparian 

areas, not just to “lost” riparian areas. 

The first two references cited by the Task Force (SCORP 1989 and TANC 1987) 

both lead to the third reference (Warner 1979). Warner (1979) discusses local losses 

(not Statewide losses) of riparian vegetation in Arizona on the order of 95 percent and 

cites Ohmart et al. (1977) and Lacey et al. (1975). The analysis conducted by Ohmart 

et al. (1977) was limited to 200 miles of the lower Colorado River and to only one 

component of riparian habitat (the cottonwood-willow community). Ohmart et al. (1977) 

hypothesized a reduction of that community over time but did not calculate a percentage 

loss. They stated that:

“Cottonwood communities have declined from high abundance (5,000 acres plus) 

along the lower Colorado River in the 1600s to scattered groves containing a few mature 

individuals today. Anderson and Ohmart (1976) have estimated that only 2,800 acres of 

cottonwood-willow community remain along the lower river. If one was to consider pure 

cottonwood communities, it would be less than 500 acres” (Ohmart et al. 1977:45)3

.

Lacey et al. (1975) performed a similar analysis for predominantly mesquite riparian 

habitat along a 58-mile reach of the lower Gila River and concluded that only about 5 

percent of the theoretical 1860 riparian base remained.

In subsequent years, the findings of both Ohmart et al. (1977) and Lacey et al. 

(1975), as well as the understanding of those findings reported in Warner 1979, 

were misunderstood and distorted in numerous publications. For example, the 

“cottonwood-willow” and “pure cottonwood” communities of Ohmart et al. (1977) were 

misunderstood to represent all riparian vegetation in the lower Colorado River (Council 

on Environmental Quality 1978:317–318). The 5 percent of original riparian habitat 

2 SCORP is the Arizona Statewide Comprehensive Outdoor Recreation Plan, and TANC is The Arizona Nature Conservancy.

3

 The amount of vegetation cited by Ohmart et al. (1977) as having been mapped by Anderson and Ohmart in 1976—i.e., 2,800 

acres of cottonwood-willow community and 500 acres of pure cottonwood for a total of 3,300 acres—is suspect. The acreage 

numbers conflict with those in Ohmart et al. 1988 (table 22), which lists a total 3,354 ha (8,288 acres) of cottonwood-willow 

(apparently inclusive of “pure cottonwood”) mapped in the same river reach (Lower Colorado River) by the same researchers 

(Anderson and Ohmart) in the same year (1976). The similarity of numbers (3,300 acres and 3,354 ha) suggests Ohmart et al. 

(1977) confused hectares and acres. If so, their entire analysis is thrown into question. 
12 USDA Forest Service RMRS-GTR-411. 2020.

remaining on the lower Gila River (Lacey et al. 1975) became 5 percent of original 

riparian habitat remaining in Arizona (Arizona State Parks 1989:60). 

Eventually, it became common in both scholarly and popular literature to refer to a 90 

percent (or over 90 percent, or 90–95 percent) loss (or loss and degradation) of original 

(or native) riparian habitat in Arizona (or in Arizona and New Mexico or in the western 

United States) (e.g., Fleishner 1994; Johnson 1989; Mac et al. 1998; National Research 

Council 2002; Noss 1995; U.S. Fish and Wildlife Service 2000; Yong and Finch 1996; 

Zaimes 2007). Few authors cite Ohmart et al. (1977) and Lacey et al. (1975) directly, 

but citation trails commonly lead back to them or to a more comprehensive analysis of 

riparian change on the lower Colorado River completed by Ohmart et al. (1988). 

Citing Mearns (1907), Ohmart et al. (1988) assumed 160,000 ha (400,000 acres) 

of indigenous riparian vegetation, not just cottonwood-willow, covered lower Colorado 

River bottomlands at the turn of the 19th century.4

 Based on vegetation mapping 

completed by Anderson and Ohmart (1984) and Younker and Andersen (1986), Ohmart et 

al. (1988) concluded that only 40,000 ha of the original 160,000 ha of riparian vegetation 

remained. Of that 40,000 ha, only about 17 percent (ca. 6,800 ha) consisted of native 

trees and shrubs. The balance of the remaining vegetation (ca. 33,200 acres) consisted 

of tamarisk and mixed native/tamarisk communities. These figures suggest that, in less 

than a century, the total volume of riparian vegetation along the lower Colorado River 

was reduced by 75 percent, and native riparian vegetation, unmixed with tamarisk, was 

reduced by more than 95 percent.

These few regional analyses are clearly inadequate to support the claim that 90 

percent (or a similarly large percentage) of riparian vegetation has been lost, or lost and 

degraded, across the entire Southwest. In this regard, we agree with Webb et al.; the “90 

percent loss” claim is unsupported by quantitative data or analysis. It simply became 

conventional wisdom legitimatized through repetition.

Postscript

The urgent concern that led to Governor Mofford’s 1990 executive order for the 

protection of riparian areas was about the loss of indigenous cottonwood, willow, and 

mesquite-dominated riparian woodland communities due to human activities. These 

woodlands are known to have increased in some drainages, notably along the Upper San 

Pedro River (Stromberg and Tellman 2009), but the preponderance of evidence suggests a 

net decrease across the Southwest. The impacts of dams, water diversions, channelization, 

bank stabilization, groundwater withdrawals, agriculture, urban development, grazing, 

invasive species, and a multitude of other factors have been extensive and well 

documented, if not well-quantified. 

4

 Ohmart et al. (1988:24) state that, “In 1894, Mearns (1907) estimated that there were about 160,000 to 180,000 ha (400,000 

to 450,000 acres) of alluvial bottomland between Fort Mohave and Fort Yuma covered by riparian vegetation.” To set the record 

straight, the estimate was not made in 1894, nor was it made by Mearns, nor was the maximum 450,000 acres. What Mearns 

(1907:125) actually wrote was, “As a result of an investigation along the Colorado River, made in January, 1902, by the hydrographic branch of the U. S. Geological Survey, the extent of the alluvial bottom land ... was found to be from 400,000 to 

500,000 acres.” Mearns, who did not venture north of Yuma during his visit to the Colorado River in 1894, likely obtained the 

cited information from the Reclamation Service’s First Annual Report, in which Lippincott and Davis (1903) refer to 400,000 to 

500,000 acres of “irrigable lands” surveyed along the Lower Colorado River in the preceding year. 
USDA Forest Service RMRS-GTR-411. 2020. 13

Webb et al. endeavored to address the issue of riparian change through time in a more 

systematic way than has been attempted in the past. In our judgement, however, repeat 

photography is not the appropriate tool. It is not sufficiently comprehensive, rigorous, or 

objective to measure change over such a large area and long period of time. Weaknesses 

of that approach for addressing this question include sparse and inadequate spatial and 

temporal coverage, spatial and temporal bias, temporal inconsistency, and a multitude of 

baselines from which to measure vegetation change in a highly changeable ecosystem. 

What is called for is a comprehensive, stream-by-stream quantitative analysis that 

encompasses, at a minimum, the entire States of Arizona and New Mexico. While we 

currently lack a well-reasoned baseline from which to quantify historical change, the 

tools needed to accomplish that task are fast improving. Geospatial technology designed 

to reconstruct pre-settlement riparian conditions has become increasingly refined 

(e.g., LANDFIRE 2008) and shows promise for assessing historical change across a 

broad, multi-state scale. Macfarlane et al. (2016), for example, used such technology to 

determine the ratio of existing native riparian vegetation cover to pre-European settlement 

riparian vegetation cover for 25,685 km of perennial streams and rivers in Utah. They 

found that approximately 62 percent of riparian vegetation along those streams showed 

significant (> 33 percent) to large (> 66 percent) departures from historical conditions. 

The departures were predominantly attributed to human land use impacts and conversion 

of native riparian vegetation to nonnative or upland vegetation types. A comparable 

analysis for the Southwest would be welcome, not only as a window on historical riparian 

conditions and past changes, but as a tool to help assess the effects of tamarisk biocontrol 

as it plays out in the coming years.

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18 USDA Forest Service RMRS-GTR-411. 2020.

Chapter 2. A Naturalized Riparian Ecosystem: 

Consequences of Tamarisk Leaf Beetle (Diorhabda spp.) 

Biocontrol 

Steven W. Carothers, R. Roy Johnson, and Kenneth J. Kingsley

Introduction 

Presence of the introduced genus Tamarix has been a perplexing problem for decades 

along rivers of the southwestern States. It is clearly an invasive species occurring along 

most perennial, ephemeral, and intermittent drainages of the Southwest including rivers, 

small streams, and normally dry washes. It seems to reach highest densities and form 

monocultures along waterways with altered flow regimes, but it can also invade unaltered 

streams and small springs, where it may become the dominant or exclusive woody species. 

Approximately 150 years after Tamarix was first introduced into the Southwest, it was 

being reviled as a notorious phreatophyte that was thought to measurably deplete ground 

and surface water at the expense of native riparian habitat (Chew 2013; Robinson 1952). 

Over the past 60-plus years, much research and effort have gone into understanding 

Tamarix biology and the largely failed struggles to eradicate this nonnative shrub/tree’s 

species (Chew 2009, 2013; El Waer et al. volume 1; Horton 1964, 1977; Sher and Quigley 

2013; Zavaleta 2000, 2013). On the issue of water consumption by Tamarix, recent 

findings discourage generalizations as to excessive water use. Findings indicate that the 

species’ complex is a stress-adapted group with a low to moderate water consumption 

that primarily replaces native vegetation when conditions within watersheds become 

unsuitable for native species colonization (Nagler and Glenn 2013; Nagler et al. 2012). 

Nagler and Glenn (2013) also demonstrated that from 1967 to 1982, water salvage 

projects by Tamarix removal did not achieve a sustainable recovery of water. Analysis of 

more recent water salvage/Tamarix removal literature concluded that increases in water 

yield after removal are only likely to occur when a Tamarix stand with high leaf area is 

replaced with low leaf area recolonizers (Shafroth et al. 2005). 

It is also well documented that its two dominant species, T. ramosissima and T. 

chinensis and their hybrids, can become a naturalized part of the landscape in many 

areas and provide a unique Tamarix-dominated habitat type (Brown et al. 1987; Chew 

2013; Johnson and Carothers 1987; Scott et al. volume 1). Although not all riparian birds 

find Tamarix to be suitable nesting habitat (i.e., woodpeckers, most cavity nesters, large 

raptors), studies over the past 30 years show that remarkably productive wildlife habitat 

is provided where Tamarix dominates and mixes with native riparian woody species 

(Bateman et al. 2013a; Brown et al. 1987; Brown and Trosset 1989; Darrah and van Riper 

2017; Hunter et al. 1988; Johnson et al. 2012; Sogge et al. 2005, 2008, 2013; van Riper 

et al. 2008). In this chapter, we call this a “naturalized community,” which is a unique 

vegetation assemblage or group of woody riparian plant species, largely dominated by 
USDA Forest Service RMRS-GTR-411. 2020. 19

the nonnative Tamarix, but intermixed with native species like mesquite (Prosopis spp.), 

cottonwood (Populus spp.), willow (Salix spp.), arrowweed (Pluchea sericea), seepwillow 

(Baccharis spp.), and others that are repeatedly found together within a riparian corridor.

The relatively brief reign of Tamarix appears to be coming to an end. In 1999, the 

Department of Agriculture through its division of Animal and Plant Health Inspection 

Service (APHIS) issued research permits for the release of Diorhabda elongata15

(Coleoptera: Chrysomelidae), a species of nonnative tamarisk leaf feeding beetle, in 

secure field cages at 10 sites in six States. Based on the success of the field trial, in 

2001 permits for the release of the beetle from field cages were issued by APHIS to the 

Agricultural Research Service, Bureau of Land Management, and Bureau of Reclamation. 

By August of 2005, the biocontrol program was officially underway. This was the 

culmination of a Tamarix biocontrol effort that had been in development for at least two 

decades (Bean et al. 2013; DeLoach et al. 2003; McLeod volume 1). These beetles of 

Asian origin rapidly spread throughout the West, beginning in limited areas of Colorado, 

Nevada, and Utah. 

By 2015, the expanding beetle invasion and their repeated episodic defoliation of 

Tamarix eventually led to varying rates of dieback. This included significant levels of 

plant mortality in some riparian ecosystems throughout the West (Bloodworth et al. 

2016; Hultine et al. 2015). The leaf-beetles have not only become more widespread than 

once predicted, they are also rapidly adapting their life cycles and behavior to facilitate 

penetration into areas where they were never expected (Hultine et al. 2015; Meinhardt and 

Ghering 2012). Thus, it finally appears as if the Tamarix invasion of wet places in aridland 

regions of North America has met its first serious challenge as a result of successful 

biocontrol. 

Many biocontrol efforts have had unforeseen consequences (Howarth 1983, 1991; 

Louda et al. 2003; Simberloff and Stilling 1996; many others). We believe that one 

unforeseen consequence of the Tamarix biocontrol triumph has been the rapid alteration 

of the functional benefit of thousands of acres of riparian habitat as breeding, stop-over, 

and wintering habitat for many species of wildlife, especially birds, in areas where the 

beetle impacts have led to repeated defoliation and plant mortality (McLeod volume 1). 

Moreover, we are not as optimistic as biocontrol proponents that the long-term benefit 

of the Tamarix destruction will be unassisted native species habitat restoration. With 

limited exceptions (see discussion on Grand Canyon below), most available evidence to 

date indicates that without massive efforts at active habitat rehabilitation, once Tamarix

is removed, it is most often replaced by a mixture of native and nonnative grasses and 

herbaceous cover, not the woody vegetation necessary to support riparian wildlife 

(Gonzales et al. 2017). 

In the arid Southwest, the beetle biocontrol has had a significant adverse impact 

on riparian habitat and wildlife conservation. Contrary to the assurances of the original 

proponents of the biocontrol efforts (DeLoach and Tracy 1997), there is little evidence 

that reestablishment of native woody species will naturally occur in most areas where 

1

 Diorhabda elongata was later reclassified as D. carinulata and after that species release, three additional species (D. carinata, D. 

elongate, and D. sublineata) were tested in cages between 2002 and 2009. All four species were eventually released at about 70 

sites in Texas (see McLeod volume 1). Today, D. carinulata and D. sublineata are known to be established and are responsible for 

most Tamarix defoliation with the former occupying the Colorado River Basin and expanding in range south and east while the 

latter is moving north and west from its original release sites in Texas.
20 USDA Forest Service RMRS-GTR-411. 2020.

biocontrol has resulted in significant defoliation and mortality of Tamarix. No serious 

efforts at habitat restoration are currently planned in any reasonable timeframe.

In this chapter, we have three objectives: first, we document the value of nonnative 

Tamarix as summer habitat for birds compared to native riparian habitats of mesquite 

bosques and cottonwood/willow, and mixed deciduous gallery woodlands; second, we 

specifically focus on the unintended consequences to native avifauna of dam construction, 

Tamarix invasion, native vertebrate colonization of the Tamarix-dominated riparian 

habitat, and subsequent biocontrol along approximately 300 miles of the Colorado River 

in Grand and Glen Canyons; and, third, we briefly review current allelopathic studies on 

the potential long-term fate of native woody riparian vegetation when growing alongside 

Tamarix. 

Importance of Tamarix to Riparian Birds 

Tamarix has not always been appreciated for its importance as wildlife habitat. 

Beginning in the late 1950s, investigations into the avian use of Tamarix-dominated 

riparian areas consistently demonstrated that the loss of native riparian vegetation and 

invasion of Tamarix had a negative effect on the population sizes of riparian birds 

(Anderson and Ohmart 1977). However, as verification of the fact that ecological 

generalizations concerning Tamarix are elusive, there were exceptions. Rosenberg et al. 

(1991) reported 14 pairs of nesting white-winged doves per acre in Tamarix-screwbean 

mesquite (Prosopis pubescens) communities in the Lower Colorado River Valley. In the 

late 1950s, Shaw (1961) and Shaw and Jett (1959) reported nesting white-winged dove 

(Zenaida asiatica) populations as high as 60 nests per acre in the “saltcedar” thickets 

adjacent to declining mesquite bosques near the mouth of the Salt and Gila Rivers near 

Gillespie Dam in central Arizona. 

Hunter et al. (1988) compared avian use on three river systems, the Lower Colorado, 

Rio Grande, and Pecos Rivers. They found that on the Pecos River in New Mexico, where 

riparian trees were mostly rare before the invasion of Tamarix, several species of obligate 

riparian birds expanded their range into the Pecos River Valley coincident with the arrival 

of Tamarix. The Hunter et al. (1988) findings were surprising as avian use of Tamarix on the 

Lower Colorado and Rio Grande was comparatively low. Brown et al. (1987) and Carothers 

and Brown (1991) also documented avian range and population increases along the 

Colorado River in Grand Canyon and linked those increases to the arrival of Tamarix. For 

both the Pecos River and Colorado River in Grand Canyon, quantitative estimates of native 

vegetation cover and bird species density prior to the invasion of Tamarix were mostly nonexistent. In the Pecos River Basin, Tamarix is reported to have invaded in 1912 (Hildebrant 

and Ohmart 1982). In Grand Canyon, Tamarix was first reported as relatively rare with only 

isolated occurrences along the river corridor in 1938 (Clover and Jotter 1944); however, as 

early as 1936 a pair of southwestern willow flycatchers was found nesting in Tamarix below 

Glen Canyon at Lees Ferry (Woodbury and Russell 1945). 

Quantitative estimates of bird life along the Pecos and Colorado Rivers are only 

available long after the respective Tamarix invasions. However, prior to the invasion of 

Tamarix along the Pecos River, Hildebrandt and Ohmart (1982) report, without citing 

the source for their conclusion, that there were few existing tall and mature stands of 

vegetation that could be used by riparian birds. In the Grand Canyon, Tamarix only 
USDA Forest Service RMRS-GTR-411. 2020. 21

proliferated significantly in the riparian zone after 1963, coincident with the construction 

and operation of Glen Canyon Dam and the significant reduction of annual scouring 

floods (see more below on Glen Canyon Dam river corridor impacts).

It was not until the surprising rate of spread and rapid and repeated defoliation and 

tree mortality caused by the Diorhabda became obvious that the value of Tamarix as 

wildlife habitat was quantified. Sogge et al. (2008) reviewed some available literature 

(see Corman and Wise-Gervais 2005) for nesting birds in Arizona and elsewhere and 

determined that 49 species used Tamarix as nesting habitat. Over 75 percent of lowand mid-elevation riparian birds in Arizona are known to build their nests in Tamarix

(Bateman et al. 2013a; Corman and Wise-Gervais 2005; Sogge et al. 2008). 

In this chapter, we have expanded the Tamarix-bird-use literature search to include: 

(1) over 50 publications spanning several decades; (2) more southwestern States than 

just Arizona; and (3) birds recorded as using Tamarix for foraging and roosting as well as 

for nesting. For reference, we compare bird use of Tamarix to that of riparian vegetation 

associations dominated by native species, including mesquite, cottonwood/willow, 

and mixed deciduous habitats (table 3). Of a total of 143 species of lowland birds in 

the southwestern United States normally found during the spring and summer months 

associated with breeding activities, 105 (73 percent) have been recorded in Tamarix, 98 

(69 percent) in mesquite, 81 (57 percent) in cottonwood-willow, and 67 (47 percent) 

in mixed deciduous habitat types. While not all the species listed in table 3 necessarily 

build their nests in the specific tree types or associations, the lists do serve to emphasize 

the relative importance of the Tamarix shrub/tree species as a wildlife resource.

The value of Tamarix to riparian wildlife is especially evident where native riparian 

species have significantly declined over the past decades. It is also apparent in disturbed 

drainages where native vegetation species can no longer recruit and survive due to land 

conversion by urbanization, agriculture, and/or alteration of the natural hydrograph 

(Sogge et al. 2013). Where native riparian species are specifically precluded, Tamarix

is a far superior habitat to no woody habitat at all. It provides structural diversity for 

riparian wildlife species that does not exist when there are no woody tree species and/or 

low-growing native and nonnative grasses and forbs. We believe increased use by birds 

of Tamarix-dominated habitats versus exclusively native species-dominated habitats is 

the result of increased general habitat structure provided by the increased foliage volume 

and foliage height diversity. A Tamarix-dominated understory is normally unavailable in 

the exclusively native stands. The Tamarix growth form provides vegetative cover and 

foraging areas from the ground up, while riparian habitat consisting of mostly native 

species normally has low-growing herbaceous plants under the gallery forest or mesquite 

bosque canopies (see Bateman et al. 2013a). In addition, Tamarix’s ability to rapidly 

establish after disturbance and its high stress tolerance compared to native vegetation 

has resulted in the rapid proliferation of vegetative cover. Thus wildlife habitat is quickly 

available where previously either low-growing, non-woody cover predominated, or only 

bare ground was present (Hultine and Dudley 2013). 

Moreover, it has been demonstrated that invertebrate density and diversity can 

increase within the riparian ecosystem in some areas where Tamarix has invaded 

(Stevens 1976, 1985; Strudly and Dalin 2013). Within a little over two decades following 

the closing of the floodgates of Glen Canyon Dam, Tamarix supported a relatively 

low species richness of invertebrates compared with native species, but it supported 
22 USDA Forest Service RMRS-GTR-411. 2020.

Table Key

* Birds recorded from the Colorado River and its tributaries in Grand Canyon (Brown et al. 1987; Corman and 

Wise-Gervais 2005).

Riparian dependency at lower elevations: 

F = facultative, occurring approximately equally in or out of riparian habitats; 

O = obligate, occurring in riparian or similar wetland habitats >90 percent of the time; 

P = preferential, occurring in riparian habitats < 90 percent of the time but more often than in nonriparian or other wetland 

habitats; 

W = may occur in riparian habitats but often occurring in other wetland types, e.g., marshes, open water, or openings near water. 

Although most species nest in the habitat in which they have been listed they may forage in or over the vegetation types 

listed below.

--------------- Vegetation type ---------------

Common name Scientific nameb

Riparian 

dependency 

code Tamarixc Mesquited

CottonwoodWilloweMixedDeciduousf

Abert’s towhee Melozone aberti O 2,11,15 X X

Acorn woodpecker Melanerpes formicivorus X

American coot Fulica americana * W 6,13,16 X

American dipper Cinclus mexicanus * O 12,17,21

American kestrel Falco sparverius * P 4,18 X X X

American robin Turdus migratorius P X X

Anna’s hummingbird Calypte anna P X X X

Arizona woodpecker Picoides stricklandi X

Ash-throated flycatcher Myiarchus cinerascens * F 2,3,15,16 X X X

Bald eagle Haliaeetus leucocephalus O 6,13 X X

Barn owl Tyto alba 2 X X X

Bell’s vireok Vireo bellii * P 1,11,15 X X X

Belted kingfisher Megaceryle alcyon O 12,17

Bendire’s thrasher Toxostoma bendirei F X

Bewick’s wren Thryomanes bewickii * 2,10,11 X X X

Black phoebe Sayornis nigricans * O 6,12,15 X X X

Black rail Laterallus jamaicensis O W 21 X

Black-caped gnatcatcher Polioptila nicriceps P X X X

Black-chinned hummingbird Archilochus alexandri * P 1,2,17,18 X X X

Black-crowned night-heron Nycticorax nycticorax * O 1,12,13 X X

Black-headed grosbeak Pheucticus 

melanocephalus

X

Black-tailed gnatcatcher Polioptila melanura * P 4,12,22 X X

Black-throated sparrow Amphispiza bilineata X

Blue grosbeak Passerina caerulea * O 1,2,3,15 X X

Blue-gray gnatcatcher Polioptila caerulea * 1,5 X X

Botteri’s Sparrow Peucaea botterii F X X X

Table 3—A comparison of breeding birds recorded in lowland riparian habitats) of the southwestern United States.
USDA Forest Service RMRS-GTR-411. 2020. 23

Table 3—Continued.

--------------- Vegetation type ---------------

Common name Scientific nameb

Riparian 

dependency 

code Tamarixc Mesquited

CottonwoodWilloweMixedDeciduousf

Bridled titmouse Baeolophus wollweberi O X X

Broad-billed hummingbird Cynanthus latirostris P X X X

Bronzed cowbird Molothrus aeneus O X X

Brown thrasher Toxostoma rufum 12,17,20 X

Brown-crested flycatcher Myiarchus tyrannulus P 12,18 X X X

Brown-headed cowbird Molothrus ater * F 1,2,3,16 X X X

Bullock’s oriole Icterus bullockii * O 2,3,15,18 X X X

Cactus wren Campylorhynchus 

brunneicapillus 

* 3,12,15 X

Canyon towhee Melozone fusca * 18 X

Canyon wren Catherpes mexicanus * F X

Cassin’s kingbird Tyrannus vociferans F 18 X X

Cattle egret Bubulcus ibis P 17,20,21 X

Chihuahuan raven Corvus cryptoleucus 12,17,18 X X X

Clapper rail Rallus crepitans O W 21 X

Clark’s grebe Aechmophorus clarkii * W 5

Cliff swallowl Petrochelidon pyrrhonota * O 4,12,17 X X X

Common black hawk Buteogallus anthracinus O X X X

Common ground-dove Columbina passerina P 18,21 X

Common moorhen Gallinula chloropus W 6,12,13 X

Common poorwill Phalaenoptilus nuttallii X

Common raven Corvus corax * 4,10,18 X X X

Common yellowthroat Geothlypis trichas * O W 11,12,15 X X X

Cooper’s hawk Accipiter cooperii * O 18 X X X

Costa’s hummingbird Calypte costae * 14 X

Couch’s kingbird Tyrannus couchii X

Crissal thrasher Toxostoma crissale * P 3,9,12,15 X

Curve-billed thrasher Toxostoma curvirostre F 7 X

Double-crested Cormorant Phalacrocorax auritus * O 21 X X

Dusky-capped flycatcher Myiarchus tuberculifer X

Eastern bluebird Sialia sialis X X

Elf owl Micrathene whitneyi P 12,20,17 X X X

European starling Sturnus vulgaris * P 4 X X X

Ferruginous pygmy-owl Glaucidium brasilianum P X X

Gambel’s quailg Callipepla gambelii * P 6,11,13 X X X

Gila woodpecker Melanerpes uropygialis P 4,15,18 X X X
24 USDA Forest Service RMRS-GTR-411. 2020.

--------------- Vegetation type ---------------

Common name Scientific nameb

Riparian 

dependency 

code Tamarixc Mesquited

CottonwoodWilloweMixedDeciduousf

Gilded flicker Colaptes chrysoides F 4 X X

Golden-fronted woodpecker Melanerpes aurifrons P

Gray hawk Buteo plagiatus O 18 X X X

Great blue heron Ardea herodias O 4,6 X X

Great egret Ardea alba O 21

Great horned owl Bubo virginianus * 4 X X X

Great-tailed grackle Quiscalus mexicanus * O 11,15

Greater roadrunner Geococcyx californianus * F 3,12,22 X X X

Green heron Butorides virescens O 6,13 X

Harris hawk Parabuteo unicinctus P X X

Hooded oriole Icterus cucullatus * P X X X

House finch F Haemorhous mexicanus * F 11,14,15 X X X

House sparrow Passer domesticus * X X

Hutton’s vireo Vireo huttoni F X

Indigo bunting Passerina cyanea * O 12,17 X X X

Killdeer Charadrius vociferus W P 6,14,16 X

Ladder-backed woodpecker Picoides scalaris * P 2,3,15,16 X X X

Lazuli bunting Passerina amoena * O 12,17 X X

Least bittern Ixobrychus exilis W 5,10 X X

Lesser goldfinch Spinus psaltria * P 11,15,18 X X X

Lesser nighthawk Chordeiles acutipennis * 4,6,13 X X X

Loggerhead shrike Lanius ludovicianus 3,21 X

Long-billed thrasher Toxostoma longirostre X

Lucy’s warbler Oreothlypis luciae * O 1,2,15,18 X X X

Marsh wren Cistothorus palustris 4,5

Meadowlark Sturnella sp. 7

Mexican jay Aphelocoma wollweberi X

Mississippi kite Ictinia mississippiensis O 19,21 X X

Montezuma quail Cyrtonyx montezumae X

Mourning dove Zenaida macroura * P 2, 6,11,13 X X X

Northern beardlesstyrannulet Camptostoma imberbe O 18 X X X

Northern cardinal Cardinalis cardinalis P 9,18 X X

Northern flicker Colaptes auratus X X X

Northern mockingbird Mimus polyglottos P 3,11,7,18 X X X

Table 3—Continued.
USDA Forest Service RMRS-GTR-411. 2020. 25

--------------- Vegetation type ---------------

Common name Scientific nameb

Riparian 

dependency 

code Tamarixc Mesquited

CottonwoodWilloweMixedDeciduousf

Northern rough-winged 

swallow Stelgidopteryx 

serripennis

P 6,12,13 X X X

Orchard oriole Icterus spurius P X X

Painted bunting Passerina ciris 22

Peregrine falconj Falco peregrinus * P 5

Phainopepla Phainopepla nitens * P 4,14,18 X X X

Pied-billed grebe Podilymbus podiceps W 5,6,10

Prairie falcon Falco mexicanus *

Purple martin Progne subis X

Pyrrhuloxia Cardinalis sinuatus P 3,12,22 X

Red-tailed hawk Buteo jamaicensis * 14,21 X X X

Red-winged blackbird Agelaius phoeniceus * W P 2,8,14,16 X

Ring-necked pheasant Phasianus colchicus 7

Rose-throated becard Pachyramphus aglaiae O X X

Ruddy duck Oxyura jamaicensis W 5,13

Rufous-winged sparrow Peucaea carpalis P X

Say’s phoebe Sayornis saya 14,15,16

Snowy egret Egretta thula O 21

Song sparrow Melospiza melodia * O W 2,6,10,15 X X X

Sora Porzana carolina O W 5 X

Spotted sandpiper Actitis macularius * W 1,21 X

Spotted towhee Pipilo maculatus 7

Sulphur-bellied flycatcher Myiodynastes luteiventris O X

Summer tanager Piranga rubra * O 2,9,12,17 X X X

Swainson’s hawk Buteo swainsoni 21 X

Thick-billed kingbird Tyrannus crassirostris P 21

Tricolored blackbird Agelaius tricolor W P 12,17,20

Tropical kingbird Tyrannus melancholicus P 9 X X

Turkey vulture Cathartes aura * 4 X X X

Varied bunting Passerina versicolor F 18 X

Verdin Auriparus flaviceps P 2,14,15 X

Vermilion flycatcher Pyrocephalus rubinus 2,18 X X X

Violet-green swallow Tachycineta thalassina * X

Virginia rail Rallus limicola O W 5 X

Western grebeh Aechmophorus 

occidentalis

* W 5

Table 3—Continued.
26 USDA Forest Service RMRS-GTR-411. 2020.

--------------- Vegetation type ---------------

Common name Scientific nameb

Riparian 

dependency 

code Tamarixc Mesquited

CottonwoodWilloweMixedDeciduousf

Western kingbird Tyrannus verticalis * F 3,15,18 X X X

Western meadowlark Sturnella neglecta 4

Western screech-owl Megascops kennicottii * P 2 X X X

Western wood-pewee Contopus sordidulus O X X X

White-breasted nuthatch Sitta carolinensis X

White-faced ibis Plegadis chihi W 12,20,17

White-throated swift Aeronautes saxatalis * 5

White-winged dove Zenaida asiatica P 2,6,8,13 X X X

Willow flycatcher Empidonax traillii * O 9,17 X

Yellow warbler Setophaga petechia * O 1,2,15,18 X X X

Yellow-billed cuckooi Coccyzus americanus * O 2,3,7,17 X X X

Yellow-breasted chat Icteria virens * O 1,13,15 X X

Yellow-headed blackbird Xanthocephalus 

xanthocephalus

O W 4,5,12,17

Zone-tailed hawk Buteo albonotatus P X X X

*58 Grand Canyon Total species: 143 105 98 81 67

Total Total Total Total

a After Johnson et al. (1977, 1987) with modifications from information gathered since those publications. Several species of cavity nesters are obligate 

or preferential riparian nesters except for using saguaros (Carnegiea gigantea), e.g., western screech-owl, elf owl, Gila woodpecker, gilded flicker, 

and American kestrel. 

b Common and scientific names after the 2017 checklist of North and Middle American birds by the North American 

Classification Committee (NACC) of the American Ornithologists’ Union. 

c Tamarisk or saltcedar; after 3 or 4 references are listed for a given species additional references are not cited.

d Velvet (Prosopis velutina), honey (P. glandulosa), or screwbean mesquites (P. pubscens). 

e Fremont (Populus fremontii) and plains cottonwoods (P. deltoides), and Goodding willow (Salix gooddingii). 

f Largely Arizona (Platanus wrightii) and California sycamores (P. racemosa); ash (Fraxinus spp.), and walnut (Juglans spp.).

g Although numerous historic records of Gambel’s quail the species was extirpated by the mid-1900s.

h Western and Clark’s grebes have been recently separated so there is confusion about which occurs (one or both?) on upper Lake Mead at the lower 

end of Grand Canyon. 

i Formerly the Yellow-billed cuckoo was a rare summer resident in the Canyon but last recorded in 1971 (Brown et al. 1987).

j Peregrines commonly nest on cliffs close to water from where they hunt for birds on and over water.

k Bell’s vireo has progressed steadily upstream in Tamarix thickets since the construction of upstream Glen Canyon Dam; one of the most noticeable 

species because of its persistent song (Brown et al. 1983).

l The cliff swallow was formerly a common breeding species along the Colorado River in Grand Canyon but by the mid-1970s it was extirpated, 

apparently due to sediment being entrapped by upstream Glen Canyon Dam, thus a lack of mud for nest-building (Brown et al. 1987). 

m The overall period covered was from 1978 to 2012 with the states of CA, AZ, NM, TX, and NV all represented and 

although the same locations were sometimes sampled more than once each sampling listed different species. 

n van Riper et al. (2008) does not list how species use riparian habitat or differentiate between use of Tamarix and other riparian vegetation for each 

species. 

o https://birdsna.org/Species-Account/bna/species/ 

Table 3—Continued.
USDA Forest Service RMRS-GTR-411. 2020. 27

a 10-fold higher biomass of invertebrates compared with the co-occurring native 

willow (Salix exigua) (Stevens 1985). Most of the invertebrate biomass increase found 

in Tamarix reported by Stevens (1985) was due to high density of the Tamarix hostspecific, nonnative leaf hopper (Opsius stactogalus)—a species that Grand Canyon birds 

commonly feed upon during the breeding season (Yard et al. 2004). In addition, Tamarix

is a rapid invader and fast to grow in conditions otherwise marginal for most native 

riparian species. These conditions include inflows and outflows to reservoirs like Lakes 

Mead and Powell, where annual tailwater flows are severely reduced from pre-dam flows, 

and where receding reservoirs expose bare ground. 

The unintended and unexpected consequence of a major reduction in wildlife habitat 

caused by the rapid and widespread loss of Tamarix-dominated habitat in many areas has 

now been well documented (Bateman et al. 2015, 2013b; Darrah and van Riper 2017; 

McLeod volume 1; Paxton et al. 2011; Sogge et al. 2008, 2013; van Riper et al. 2008; 

Yard et al. 2004). Sogge et al. (2008) also determined that there were no negative effects 

I. Tamarix citationsm Rivers and states 

1. Brown et al. (1987) Colorado R. in Grand Canyon, AZ

2. Rosenberg et al. (1991) Lower Colorado R. Valley, AZ & CA

3. Hunter et al. (1988) Pecos, Rio Grande, lower Colorado rivers, TX, AZ & CA 

4. Anderson (2017) Lower Colorado R. Valley, AZ & CA

5. R. Roy Johnson (Pers. observ.), unpublished Various, SW U.S.

6. Johnson and Simpson (MS b) Salt R., AZ

7. Livingston and Schemnitz (1996). Pecos R., NM

8. Rea (2007). Gila R., AZ

9. Hunter et al. (1987) Colorado, Gila, Pecos, Rio Grande, Salt, San Pedro, Santa Cruz, Verde rivers, TX, AZ & CA

10. Johnson et al. (2000) Lower Salt River, AZ

11. Engel-Wilson and Ohmart (1979) Rio Grande, TX

12. Sogge et al. (2008) Various, SW U.S.

13. Johnson and Simpson (MS a) Gila, Salt, Verde rivers, AZ

14. Cardiff et al. (1978a) Mojave R., CA

15. van Riper et al. (2008)n Lower Colorado R., AZ & CA

16. Cardiff et al. (1978b) Mojave R., CA

17. Paxton et al. (2011) Southwestern U.S.

18. Brand et al. (2008) San Pedro R., AZ

19. Glinski and Ohmart (1983) San Pedro R., AZ

20. Birds of North America accountso Various, SW U.S.

21. Corman and Wise-Gevais (2005) Arizona rivers, AZ

22. Hunter et al. (1985) Colorado, Pecos, Rio Grande rivers, SW U.S.

II. Mesquite citations: (Arnold 1940; Bendire 1872, 1892; Brandt 1951; Brown 1987; Brown et al. 1984; Carothers and Brown 1991; Carothers 

and Sharber 1976; Carothers et al. 1976; Corman and Wise-Gervais 2005; Dawson 1921; Gavin and Sowles 1975; Glinski and Ohmart 

1983; Huels et al. 2013; Hunter et al. 1987; Johnson and Simpson Ms A, Ms B; Johnson et al. 2000; Monson and Phillips 1981; Ohmart et al. 

1988; Phillips et al. 1964; Rea 1983, 2007; Rosenberg et al. 1991; Stamp 1978; Webb et al. 2014; Willson and Carothers 1979).

III. Cottonwood-willow citations: (Brandt 1951; Carothers and Johnson 1976; Carothers et al. 1974; Corman and Wise-Gervais 2005; EngelWilson and Ohmart 1979; Glinski and Ohmart 1983; Hunter et al. 1987; Johnson and Simpson Ms A, Ms B; Johnson et al. 2000; Monson 

and Phillips 1981; Ohmart et al. 1988; Phillips et al. 1964; Rea 1983, 2007; Rosenberg et al. 1991; Stamp 1978; Webb et al. 2014).

IV. Mixed deciduous citations: (Bock and Bock 1984; Carothers et al. 1974; Monson and Phillips 1981; Phillips et al. 1964).

Table 3—Continued.
28 USDA Forest Service RMRS-GTR-411. 2020.

from breeding in Tamarix habitats expressed in several studies on the southwestern willow 

flycatcher (Empidonax traillii extimis). This finding contradicted the conventional wisdom 

that birds that nested in Tamarix had lower reproductive fitness as measured in nest and 

fledgling survival (see Brand et al. 2008; DeLoach et al. 2000). In Grand Canyon, the 

earliest southwestern willow flycatcher breeding record (1936) reported by Woodbury 

and Russell (1945) in the Lees Ferry area was in Tamarix, and every nest found along the 

Colorado River in Grand Canyon by Brown (1988) was also in Tamarix (table 4)2

.

6

Sogge 

et al. (2008) warned that the overall ecological costs and benefits of Tamarix control were 

difficult to predict and that restoration projects that resulted in removal of Tamarix without 

replacement with high quality habitat had the potential to reduce net riparian habitat value 

for local and/or regional bird populations. Their predictions were prophetic.

Darrah and van Riper (2017) proposed that the impacts to riparian bird species 

resulting from Diorhabda biocontrol will continue until regrowth of native vegetation 

is established. Whereas they recognized that active restoration may be necessary in 

some areas, they did not discuss the challenges that will be faced by private, State, and 

Federal organizations in providing human and financial resources necessary for active 

habitat restoration37

on thousands of acres of Tamarix habitat destroyed by the beetle. 

One estimate for comprehensive eradication and restoration of Tamarix-dominated 

habitat prior to biocontrol concluded that, although costs per acre are difficult to average 

depending on local site conditions and other factors, they normally reach thousands of 

dollars per acre (Zavaleta 2000). 

An early review of the potential for habitat restoration following anticipated successful 

biocontrol questioned whether the habitat that occurs following Tamarix control and 

revegetation was any better for wildlife than the original habitat (Shafroth et al. 2005). 

Indeed, in their review, Shafroth et al. (2005) observed that following Tamarix control 

efforts, failure to plan and implement restoration efforts could result in recolonization of 

a site by other exotic species. This prediction has been largely verified by Gonzales et al. 

(2017) in their study of 244 sites where Tamarix was removed or disadvantaged by one 

means or another. In another study of a restoration project on Las Vegas Wash in Nevada 

where Tamarix was removed from several sites, replaced by native trees and shrubs, and 

then monitored for avian use, benefits to birds were not as evident as predicted (Shanahan 

et al. 2011). However, the apparent lack of wildlife benefits in Tamarix removal projects 

reviewed to date may be partially a consequence of the assessments being too early after 

control impacts. There may have been insufficient time for revegetation following the 

defoliation and mortality of the exotic species. 

2

 Nesting records for the endangered southwestern willow flycatcher are important not only because of the rarity of the species 

but because of its nesting in Tamarix in Grand Canyon rather than in willows or other native species. The flycatcher had formerly 

occurred throughout much of the state (Phillips et al. 1964) but had largely disappeared by the 1970s. For example, 34 nests 

were collected along the Colorado River in the Yuma area in 1902 by Herbert Brown (Unitt 1987) but the species was later extirpated as a breeding bird for that region (Rosenberg et al. 1991). Monson and Phillips (1981) wrote, “No nests found [in Arizona] 

since 1970,” obviously unaware of our nesting records (table 2) and those published by Carothers and Johnson (1975). Thus, 

the southwestern willow flycatchers nesting in Grand Canyon during the 1970s and into the 1980s represented the only known 

breeding population for the state at that time.

3

 We are distinguishing between passive restoration and active restoration as defined by Shafroth et al. 2013; i.e.: “Passive restoration refers to facilitating the return of desirable system dynamics and species composition by removing one or more underlying stressor(s). Active restoration approaches include manipulating a site to prepare it for restoration; revegetating the site by 

introducing seeds, transplant stock, or cuttings; or irrigating or otherwise manipulating the site to enhance recovery” (Shafroth et 

al. 2013, p. 411). 
USDA Forest Service RMRS-GTR-411. 2020. 29

Whereas biocontrol appears to be satisfying the long-desired eradication of this 

invasive species much more quickly than expected, there is little evidence at this time that 

native riparian vegetation will, without active restoration, eventually colonize most areas 

as originally assumed (DeLoach and Tracy 1997). Beetle-release proponents admitted 

early in the development of the biocontrol that evidence for native vegetation species to 

replace Tamarix was circumstantial and not well supported (see McLeod volume 1) and 

clearly underestimated the value of the Tamarix to native riparian wildlife species. Recent 

revegetation efforts attempted along the Colorado River in Glen Canyon indicate that once 

Tamarix is gone, reestablishment of native woody vegetation can be rapid and effective, but 

only coincident with labor intensive planting, fencing, and watering (Stevens et al. 2015). 

Table 4—Summer records of the endangered southwestern willow flycatcher (Empidonax traillii extimus) from the Colorado River 

in Grand Canyon prior to 1989; see especially Brown (1988).

Date Location Observation Observer and reference

June 7, 1933 RM 21 Female collected Woodbury and Russell (1945)

August 18, 1936 Near Lees Ferry Nest collected from Tamarix Woodbury and Russell (1945)

June 16, 1953 Lees Ferry Specimen collected R.W. Dickerman; Monson (1953)

June 17, 1953 Mouth of Little Colorado River Specimen collected R.W. Dickerman; Monson (1953)

July 12, 1971 Cardenas Marsh Nest with eggs in Goodding 

willow

Johnson; Carothers and Johnson (1975)

July 27, 1971 Cardenas Marsh 1 pair Carothers; Carothers and Johnson (1975)

May 20, 1974 Cardenas Marsh 1 pair Johnson; Carothers and Johnson (1975)

1974-1976 225 mi of Colorado R. 1 known pair Carothers et al. 1976; Carothers and 

Sharber (1976)

1974-1976 225 mi of Colorado R. 1 known pair Carothers et al. 1976; Carothers and 

Sharber (1976)

1982 Saddle Can. to Nankoweap Crk. 1 singing male Brown (1988)

1982 Cardenas Marsh 1 singing male Brown (1988)

1983 Saddle Can. to Nankoweap Crk. 4 singing males Brown (1988)

1984 Saddle Can. to Nankoweap Crk. 3 singing males, 

2 nests* In Tamarix

Brown (1988)

1984 Cardenas Marsh 1 singing male Brown (1988)

1985 Saddle Can. to Nankoweap Crk. 7 singing males,

4 nests* In Tamarix

Brown (1988)

1985 Cardenas Marsh 1 singing male Brown (1988)

1986 Saddle Can. to Nankoweap Crk. 8 singing males 

2 nests* In Tamarix

Brown (1988)

1986 Nankoweap Crk. to Kwagunt Crk. 1 singing male Brown (1988)

1986 Cardenas Marsh 2 singing males Brown (1988)

1987 Saddle Can. to Nankoweap Crk. 4 singing males

2 nests* In Tamarix

Brown (1988)

1987 Cardenas Marsh 3 singing males Brown (1988)

1

 River Miles downstream from Lees Ferry.

*Exact location of nests not given by Brown (1988).
30 USDA Forest Service RMRS-GTR-411. 2020.

However, in other areas where active restoration has not been implemented after 

Tamarix removal, passive revegetation is almost exclusively limited to recolonization by a 

mixture of native and nonnative grasses and herbaceous cover with the conspicuous absence 

of woody species (González et al. 2017; Sher et al. 2018). In a study in Grand Canyon 

tributaries, Belote et al. (2010) found that 1 to 3 years after mechanical removal of Tamarix, 

there was no recruitment or increase in the number of native species; however, there was 

approximately a 50 percent decline in precipitation between pre- and post-restoration periods 

that could have prevented a greater success in restoration of native species.

Tamarix-Dominated Riparian Habitat In Grand Canyon: A New Wildlife 

Resource 

Prior to the closing of the flood gates of Glen Canyon Dam in 1963 and creation of 

Lake Powell, annual flood scour in the river channel precluded riparian growth in all but a 

narrow margin at the edge of the high water line. Composed of honey mesquite (Prosopis

glandulosa), netleaf hackberry (Celtis reticulata), four species of Baccharis, and western 

redbud (Cercis occidentalis), this narrow band of vegetation constituted the extent of 

riparian habitat in the river corridor. Below this narrow band of high water line vegetation, 

the annual scour zone was largely devoid of woody vegetation and composed of annual 

forbs and grasses that could establish between flood flows. After the dam began to hold 

back annual highwater and annual scouring floods were largely controlled, a Tamarixdominated vegetation zone became established in the river corridor. This included Tamarix

(T. chinensis/ramosissima) intermixed with native species, arrowweed (Pluchea sericea), 

Baccharis spp., coyote willow (Salix exigua), and Goodding’s willow (Salix gooddingii) 

where previously only forbs and grasses occurred. Figures 1 and 2 provide a graphic 

representation of the riparian vegetation condition of the river corridor in Grand Canyon 

before and after the Dam. Scott et al. (volume 1) demonstrate these vegetation changes with 

photographic comparisons. 

By the time Glen Canyon Dam had been in place for 20 years (1963-1983), natural 

resources studies below the Dam were focused on attempting to quantify changes that damcontrolled flow had on aquatic and terrestrial ecosystems within Grand Canyon National 

Park. At that time, approximately 300 miles of river corridor from the dam to Lake Mead 

were under control and management of the National Park Service, but dam releases were 

under control of the Bureau of Reclamation and fluctuated according to power needs 

by cities. Also, during this time, researchers first recognized that the terrestrial Tamarixdominated habitat was becoming increasingly more utilized by river corridor wildlife in 

Grand Canyon. The New High Water Zone (NHWZ; see fig. 2) habitat was recognized as a 

boon to native wildlife species. 

Within only two decades of the closing of the flood gates at Glen Canyon Dam in 

1963, Tamarix-dominated habitat below the dam became remarkably productive for 

native vertebrates and invertebrates and some nonnative invertebrates as well (Brown et 

al. 1987; Stevens 1985; Yard et al. 2004). The NHWZ vegetation introduced an extensive 

“new” river margin zone of vegetation along the river corridor from Glen Canyon Dam to 

the inflow of Lake Mead. Moreover, studies soon demonstrated that the NHWZ not only 

increased the overall density of canyon birds, but also contributed to the range expansions 

of several avian species that invaded from the south where native habitats and some avian 
USDA Forest Service RMRS-GTR-411. 2020. 31

species were in rapid decline (Brown et al. 1987; Rosenberg et al. 1991). Table 5 presents 

a list of the birds and mammals that have colonized or experienced range extensions into 

Tamarix-dominated habitat along the Colorado River in Grand Canyon since the closing of 

Glen Canyon Dam floodgates. 

Remarkably, it soon became obvious that NHWZ was expanding rapidly and that 

it was significantly more productive and robust in vegetative growth and in carrying 

capacity for riparian birds and other vertebrates and invertebrates than the native 

mesquite-dominated Old High Water Zone (OHWZ; fig. 1) (Brown and Trosset 1989; 

Carothers and Brown 1991; Holmes et al. 2005; Ruffner et al. 1978; Stevens 1995; Yard 

et al. 2004). Studies targeted specifically on OHWZ tree/shrub species indicated that 

plants in this zone were slowly senescing, presumably due to the absence of flows no 

longer reaching the OHWZ habitats \(Anderson and Ruffner 1987, 1988).

Figure 1—Pre-dam floods 

normally scoured the 

lower-lying river terraces 

of all woody vegetation, 

allowing only a few 

annual grasses and herbs 

to develop in the predam flood zone. From 

Grand Canyon Birds by 

Bryan T. Brown, Steven 

W. Carothers, and R. Roy 

Johnson, © 1987 The 

Arizona Board of Regents. 

Reprinted by permission 

of the University of 

Arizona Press.

Figure 2—Beginning in 1963, 

Glen Canyon Dam prevented 

the annual scouring floods. 

A new zone of riparian 

vegetation, dominated by the 

exotic Tamarix, developed 

in the pre-dam flood zone 

and was quickly colonized 

by riparian breeding birds 

and other native vertebrates. 

From Grand Canyon Birds 

by Bryan T. Brown, Steven 

W. Carothers, and R. Roy 

Johnson, © 1987 The Arizona 

Board of Regents. Reprinted 

by permission of the 

University of Arizona Press.
32 USDA Forest Service RMRS-GTR-411. 2020.

Table 5—Birds and mammals that have colonized (newly arrived species) or experienced range extensions (moved upriver from 

downstream breeding locations) into Tamarix-dominated habitat along the Colorado River in Grand Canyon since the closing 

of Glen Canyon Dam floodgates in 1963 (modified from Johnson and Carothers 1987).

Common name Scientific name Status and notes

Birds

Western grebea Aechmophorus occidentalis Colonized, nesting in dead Tamarix at upper Lake Mead

Clark’s grebea A. clarkii Colonized, nesting in dead Tamarix at upper Lake Mead

Black-chinned hummingbird Archilochus alexandri Colonized

Green heron Butorides virescens Nonbreeding

Black-crowned night-heron Nycticorax nycticorax Colonized, nesting in dead Tamarix at upper Lake Mead

Willow flycatcher Empidonax traillii Colonized, endangered species

Vermilion flycatcher Pyrocephalus rubinus Range expansion

Bell’s vireo Vireo bellii Range expansion

Lesser goldfinch Spinus psaltria Colonized

Yellow-breasted chat Icteria virens Colonized

Hooded oriole Icterus cucullatus Range expansion

Bullock's oriole I. bullockii Colonized

Red-winged blackbird Agelaius phoeniceus Colonized

Brown-headed cowbird Molothrus ater Colonized

Great-tailed grackle Quiscalus mexicanus Range expansion

Lucy's warbler Oreothlypis luciae Colonized

Common yellowthroat Geothlypis trichas Colonized

Yellow warbler Setophaga petchia Colonized

Summer tanager Piranga rubra Range expansion

Blue grosbeak Passerina caerulea Colonized

Lazuli bunting P. amoena Colonized

Indigo bunting P. cyanea Range expansion

Mammals

Beaver Castor canadensis Colonized

Brush mouse Peromyscus boylei Colonized

Deer mouse P. maniculatus Colonized

a

The western grebe was recently split into two species and it is not known if only one or both are nesting here.

Table 6—Comparison of species richness and population densities for spring and summer breeding birds showing preferences for 

Tamarix-dominated habitat relative to mesquite-dominated habitat along the Colorado River in Grand Canyon; after Brown 

(1987) and Brown and Trosett (1989). 

Type of study plot and 

year Total species Species/site

Average 

species/site

Population (densities/site 

pairs/40 ha)

Average (density/

site pairs/40 ha)

Tamarix 1984 21 1-15 7.9 200-1,200 611

Tamarix 1985 23 1-16 8.0 100-1,200 565

Prosopis 1984 20 2-14 8.6 182-986 449

Prosopis 1985 20 3-17 8.0 73-943 379

Prosopis 1985 20 3-17 8.0 73-943 379

1

 River Miles downstream from Lees Ferry.

*Exact location of nests not given by Brown (1988).
USDA Forest Service RMRS-GTR-411. 2020. 33

Table 6 provides a summary of the number of species and relative density of birds 

found in both the Tamarix- and mesquite-dominated habitats (NHWZ vs. OHWZ, 

respectively). While species richness was similar in both habitats with 20-23 nesting 

species, surprisingly, and against all conventional wisdom at the time, avian density in 

the Tamarix-dominated habitat was significantly greater than in the mesquite-dominated 

habitats with similar cover and height attributes (Brown 1987; Willson and Carothers 

1979). Additionally, most species demonstrated a preference for Tamarix as nesting 

habitat. For example, black-chinned hummingbird (Archilochus alexandri), blue-gray 

gnatcatcher (Polioptila caerulea), Lucy’s warbler (Oreothlypis luciae), yellow warbler 

(Setophaga petechia), and common yellowthroat (Geothlypis trichas) occurred most 

consistently in the Tamarix-dominated NHWZ, with these five species comprising 51.5 

percent of the total density and 31 percent of the breeding species in 1985 at the inner 

canyon study sites. 

Lucy’s warbler and black-chinned hummingbird were the most abundant and 

widespread species in Tamarix-dominated habitats during 1984 and 1985 and remarkably, 

of 24 Lucy’s warbler nests found during that time, 15 (62.5 percent) were in Tamarix. 

What is most remarkable about the Lucy’s warbler nests in Tamarix is the fact that the 

warbler is primarily a cavity-nesting species and Tamarix is not known for an abundance 

of cavities. The warblers were all found forming a cavity nest (pseudo-cavity) within the 

clumps of leaf litter debris caught in the forks of upper canopy branches (Johnson et al. 

1997). In addition, all 12 nests of the southwestern willow flycatcher in the Grand Canyon 

were in Tamarix (Brown 1988). 

The overriding preference for birds nesting in Tamarix in the Grand Canyon led 

Brown and Trosset (1989) to conclude that Tamarix-dominated habitats can be the 

ecological equivalent of native communities that are required habitat of some obligate 

riparian nesting birds. This is especially true in a situation like Grand Canyon where the 

NHWZ vegetation community did not displace a community composed of native plants 

but became established where previously there was no nesting habitat. 

While the impact of Glen Canyon Dam on terrestrial habitats of the river corridor 

clearly had created habitat for native species where previously habitat did not exist, 

dam-related changes to the aquatic ecosystem were not as favorable to the native aquatic 

species, especially the native fish. The river had been dramatically changed from an 

aquatic ecosystem driven by an annual hydrograph with seasonal flow periodicity, 

sediment laden flows, and dramatically fluctuating water temperatures to a relatively 

steady flow, sediment starved, perpetually cold system that was largely incapable of 

nurturing the native aquatic species (Carothers and Brown 1991; Gloss et al. 2005; 

Johnson and Carothers 1987). Within the National Park, the post-dam environment bore 

no semblance to the pre-dam river. One of the major findings from research on changes 

to the aquatic and terrestrial ecosystems of the riparian corridor below Glen Canyon Dam 

concluded that the exotic Tamarix species had become a wildlife-valuable “naturalized” 

element of the land-based ecosystem. This Tamarix-based ecosystem has been termed a 

naturalized ecosystem, supporting high densities of native vertebrates, while the aquatic 

ecosystem was termed an exotic ecosystem due to the loss of indigenous species and 

addition of numerous exotics (Johnson and Carothers 1987). 
34 USDA Forest Service RMRS-GTR-411. 2020.

Management of Tamarix in Grand Canyon: A Dilemma

Describing the dilemma of resource management in a National Park, where intense 

efforts are normally expended to eliminate all nonnative plants and animals, Johnson 

and Carothers (1987) opined on the futility of attempting to remove Tamarix from 

Grand Canyon. They argued for accepting and managing the relatively new and wildlifesupporting Tamarix-dominated riparian habitat as a “naturalized” community. Johnson 

and Carothers (1987) described the dam-influenced habitat as follows: 

 “... a naturalized ecosystem contains biotic communities with both indigenous and 

exotic plants and/or animals. In these communities, dominance or predominance is not 

a function of species origin (that is, native or nonnative), and the indigenous biota is not 

threatened either in species richness or population sizes by exotic species. In naturalized 

ecosystems biotic and abiotic processes have either reached or are evolving toward an 

equilibrium in which exotics do not restrict or interfere with native organisms or ecological 

processes, rather than evolving toward the destruction of components and processes of the 

original, natural ecosystem. If native species are extirpated or their populations are greatly 

reduced, the ecosystem cannot be considered naturalized. In Grand Canyon new post-dam 

riparian vegetation has led to larger populations of native species and generally has been 

beneficial to wildlife as well as recreationists.” 

In Grand Canyon, the “new” habitat became so productive for wildlife, especially birds, 

small mammals, reptiles, and amphibians that the National Park Service (NPS) eventually 

recognized the new vegetation zone as a naturalized ecosystem that would never be returned 

to its natural state as long as Glen Canyon Dam controlled the river (Sharrow 1990). As such, 

the NPS recognized the new riparian habitat as a wildlife resource and made no efforts to 

remove or attempt to control Tamarix in the dam-influenced river corridor. 

The Tamarix-dominated riparian zone in the Grand Canyon is an unusual situation 

compared to most areas that are invaded by the exotic plant. In general, conditions allowing 

colonization by and proliferation of Tamarix are varied. Contrary to popular opinion, 

Tamarix is a weak competitor compared with native riparian species in systems where the 

stream hydrograph is largely unaltered (Sher et al. 2000; Stromberg et al. 2007). Rarely 

does Tamarix have an opportunity to colonize when perennial stream systems are in their 

native state (hydroriparian ecosystems), except where portions of a stream are naturally 

intermittent (Johnson et al. 1984). 

Sometimes even minor changes in groundwater levels, stream flow and extended 

periods of stream intermittency can tip the scales away from native species dominance to 

conditions that favor the nonnative plant (Stromberg et al. 2007). However, in highly altered 

rivers—like the Lower Colorado River, where dams, channelization, groundwater pumping, 

and other factors leave no semblance of a natural hydrograph—Tamarix reaches its highest 

density and is often found in monotypic stands (Anderson 2017; Ohmart et al. 1988). 

Normally, in moderately altered systems that still maintain natural seasonal periodicity—

but where flood control structures preclude large floods—Tamarix is typically intermixed 

with native woody vegetation (Nagler et al. 2011). 

The Invasion of Tamarisk Leaf Beetles in Grand Canyon 

The first arrival of Tamarix beetles in Grand Canyon after their release in 2001 is 

unknown, but by 2009 they were commonly seen, and the effects of year-after-year 

defoliation, although unquantified until 2015, were becoming more evident. By 2015, the 
USDA Forest Service RMRS-GTR-411. 2020. 35

Tamarix along the river corridor was dying in many areas, with the mortality increasing 

with each passing year (L.E. Stevens, personal communication to S.W. Carothers). To 

quantify the rate and levels of defoliation of Tamarix, Flesh and Stevens (unpublished 

data, see fig. 3) visually estimated the percent defoliation by randomly selecting 266 

points in the 240 miles between Glen Canyon Dam and Diamond Creek. Where it 

occurred, defoliation on individual trees was conspicuously evident as the once luxuriant 

foliage was reduced to dead branches and twigs showing no living leaves over most of 

the plant. The results of the 2015 defoliation estimate showed varying rates of plant mortality 

depending on location within the 240-mile reach (fig. 3). Mortality reached very high levels 

in the Glen Canyon reach down to about river-mile 50 where 70 to 80 percent of Tamarix

was dead or dying. Below river-mile 50 to river-mile 225 defoliation was less, but mostly 

ranging from 10 to 50 percent. 

Below river-mile 225 to river-mile 280 high levels of defoliation were like those 

above Lees Ferry in Glen Canyon (S.W. Carothers 2005-2018, personal observations, and 

fig. 4). The relatively low levels of defoliation found sporadically on some beaches in the 

canyon represent areas where Tamarix density is either relatively low, or limited areas 

where the leaf beetle has not yet invaded. It is expected that within the next few years, all 

the Tamarix in Glen and Grand Canyon will be defoliated and either dead or dying and no 

longer a significant wildlife habitat resource. It is important to clarify that some once dense 

and widespread stands of Tamarix on upper terraces of Lake Mead sediments have been 

completely dead for the last decade or more as the lake has reached all-time low levels. The 

water table has dropped so rapidly that even fast-growing and drought-hardy Tamarix has not 

been able to survive. 

Figure 3—Percent defoliation of Tamarix by Diorhabda beetles along the Colorado River from below Glen Canyon Dam to the 

upper reaches of Lake Mead as determined by visual estimates in 2015. Defoliation was most extreme in the first 50-70 miles 

below the dam on south facing beaches in the upper river areas and high in the upper reaches of Lake Mead. Defoliation 

was less in the interior reaches of Grand Canyon. Data after Flesh and Stevens (unpublished manuscript 2015). Figure by 

Lawrence E. Stevens.
36 USDA Forest Service RMRS-GTR-411. 2020.

Figure 4—A comparison of the effects of tamarisk leaf beetle defoliation. Top photo 2005, before leaf beetle defoliation, photo by 

Steven W. Carothers. Bottom 2017, after defoliation, photo by Robb Irwin Eidemiller. Colorado River mile 7.5; Glen Canyon 

National Recreation Area, Arizona.
USDA Forest Service RMRS-GTR-411. 2020. 37

While it is too early to estimate the ground cover of the future in these canyon areas, 

OHWZ species are slowly moving downslope into the NHWZ and these species are 

expected to continue to increase in both size and area occupied (Holmes et al. 2005). In 

addition, the area occupied by and the structural robustness of the native shrub arrowweed, 

typically a NHWZ species, is steadily increasing in the shadow of declining Tamarix. 

Arrowweed is not generally recognized as nesting habitat for birds, but that may change as 

more of the moisture and soil nutrients previously consumed by Tamarix become available 

and arrowweed takes on a more robust growth form. 

Tamarix Disruption of Mycorrhizal Fungal Communities of Cottonwood Trees 

The iconic riparian gallery forest species of southwestern riparian habitats is the 

cottonwood tree. Though riparian ecologists often refer to the cottonwood-willow gallery 

forest community, the overriding foliage of the community is almost always provided 

by the cottonwood tree. It is the cottonwood tree, rather than the willow, that provides 

most canopy cover and root systems; and it is this species that accounts for unusually 

high wildlife productivity of the gallery forest (Johnson and Jones 1977). Willow trees 

require more water than cottonwoods and are usually found in a narrow band, one or two 

trees deep adjacent to the sides of a river or stream. Cottonwood trees can often be found 

throughout the entire floodplain where the water table is sufficient. In a recent estimate of 

the relative frequency of cottonwood versus willow trees within the gallery forest of the 

Upper San Pedro River, Carothers (2016) estimated that cottonwood trees contributed up 

to 90 percent of the riparian woodland, with willow only accounting for 5 to 10 percent. 

Thus, if cottonwood trees are disadvantaged or cannot recruit, the entire forest ecosystem 

is compromised. Cottonwood recruitment largely depends on natural river flow regimes 

(Stromberg 1993). 

Recent studies have firmly established that healthy plant communities include a wide 

variety of interacting and interdependent species at many taxonomic levels, some of them 

cryptic and poorly understood (Corenblit et al. 2014; Franklin et al. 2016). Mycorrhizae 

and fungal endophytes, as well as a host of microbes, appear to be important participants 

in healthy plant communities. The contribution of mycorrhizal and other soil fungi to 

establishment, growth, and survival of cottonwood and other trees is becoming somewhat 

better understood as a result of recent and current research. 

Conventional mycorrhizal fungi include the arbuscular endomycorrhizal (AM) 

fungi, which have been found in upward of 80 percent of terrestrial plant species, and 

ectomycorrhizal fungi (EM), which are found in about 2 percent of species (Smith and 

Read 2010). Cottonwood trees apparently have both, with probably many other associated 

microbes and fungi. The relationship is mutualistic. In this symbiotic relationship, fungi 

gain energy and carbon from tree photosynthesis while providing the tree with enhanced 

nutrient and water gathering capabilities from soil (Buckling et al. 2012). The trees 

depend on fungi for growth (Ghering et al. 2006; Meinhardt and Ghering 2012, 2013). 

Other biotic components of the rhizosphere and rhizoplane are only beginning to be 

understood, and knowledge of their roles in establishment and growth of native riparian 

plants is very limited. A few endophytic components are beginning to be understood 

(Lau et al. 2013), as are the complex interactions between species and individuals in 

establishment and maintenance of riparian communities (Corenblit et al. 2014). 
38 USDA Forest Service RMRS-GTR-411. 2020.

Studies on the influence of Tamarix on beneficial mycorrhizal fungal communities 

of native cottonwood trees reveal one of the more insidious capabilities that this invasive 

tree/shrub employs to gain a competitive advantage over native riparian woody species. 

For example, when cottonwood trees are found in close association with Tamarix, the 

mycorrhizal relationship is disrupted by the invasive species’ ability to alter soil chemistry 

by concentrating chemical compounds like salts, nitrogen, and phosphorus in leaf litter 

(Hultine et al. 2015; Meinhardt and Ghering 2012, 2013). Researchers have demonstrated 

that cottonwood trees growing in association with even low-density of Tamarix have 

two-fold reduced mycorrhizal colonization compared to native trees that do not have the 

invasive tree/shrub as a neighbor (Meinhardt and Ghering 2012). The impact of Tamarix

on the other gallery forest species is not as well studied, but both willow (Beauchamp et al. 

2006) and mesquite (Titus et al. 2003) are known to benefit from mycorrhizae and both are 

likely negatively impacted when in association with Tamarix. 

Whereas studies on complex mycorrhizal interactions within the soil and roots of 

cottonwood trees in association with Tamarix are still in their infancy, it appears that the 

long-term survival of the mixed native/nonnative Tamarix-dominated riparian habitat was 

uncertain even before the Tamarix leaf-eating beetle was released. As Hultine et al. (2015) 

emphasized, the combination of climate change and the ability of Tamarix to disrupt the 

integrity of mycorrhizal relationships in the native gallery forest foundation species is 

a harbinger of long-term deterioration of native riparian habitats when growing in the 

presence of Tamarix. Thus, as the Tamarix declines because of the biocontrol efforts, the 

possibility of long-term native species recolonization of suitable habitat may eventually be 

enhanced. 

However, the length of time and suite of conditions needed to remove or sufficiently 

attenuate the disruptive effects of Tamarix on mutualistic rhizosphere and rhizoplane 

species are not well studied. Therefore, the potential for reasonably complete restoration of 

native riparian communities is still largely unknown. Wildlife ecologists in the past several 

decades have promoted the benefits of the Tamarix-dominated riparian habitats, measured 

as increases in range and density of many vertebrate and invertebrate species. While we 

promoted the wildlife benefits, the Tamarix has apparently been altering the soil around and 

below the remaining native plant species. Hultine et al. (2015) suggest that the combination 

of mycorrhizal disruption by Tamarix and the vagaries of climate change constitute threats 

to riparian habitats that we had not previously anticipated. 

These recent findings have clear implications on the future restoration and survival of 

the cottonwood forest type that is already considered one of the most threatened vegetation 

communities in the United States (Stromberg 1993; Webb et al. 2007). Dixon et al. (2009) 

reviewed the most up-to-date model predictions on climate change and found that the 

drier, hotter southwestern climate of the future would result in decreases in cottonwoodwillow forests within the Upper San Pedro River by two-thirds. The additional mycorrhizal 

impacts to the gallery forest ecosystem when Tamarix is present are yet another threat to the 

declining riparian habitat throughout the southwestern United States. 

Conclusions 

The human life span is short, and the attention span of researchers is even shorter 

compared to the time needed for ecological change. The reign of Tamarix spp. as a dominant 
USDA Forest Service RMRS-GTR-411. 2020. 39

or co-dominating invasive species in western riparian areas has also been short. Longerterm consequences of this invasion are unknown. Along some rivers, especially those with 

drastically altered flow regimes, Tamarix has created a unique ecosystem that has been 

exploited by many wildlife species, including some that have lost habitat in the period since 

human activities started to have a significant impact on the earth’s geology and ecosystems 

(the Anthropocene, see Kingsley, this volume). Currently (2019) we are witnessing major 

changes in the Tamarix-dominated ecosystem because of the very successful biocontrol 

effort. We have documented potential riparian wildlife impacts associated with the removal 

of Tamarix and cited studies documenting that the Tamarix-dominated native/nonnative 

riparian community can support more avian species than native species-dominated habitats 

of cottonwood-willow, mesquite, and mixed deciduous woodlands. It is now apparent that 

the biocontrol program was pursued with the removal or disadvantaging of Tamarix as a 

major goal without a clear understanding of revegetation dynamics post-biocontrol. 

It appears so far that in the absence of active restoration efforts, recolonization of the 

post-biocontrol Tamarix-dominated habitat is almost exclusively limited to a mixture of 

native and nonnative grasses and often weedy, herbaceous cover with the conspicuous 

absence of woody species. Thus, at least in the short term, the biocontrol effort has resulted 

in the loss of important wildlife-producing habitat without replacement. In most areas 

where Tamarix has proliferated, both human-caused alteration of natural flow regimes and 

Tamarix-caused alteration of soils are complicating or prohibiting the establishment of 

native riparian woody plants. 

In the specific case of the Grand Canyon, an unprecedented naturalized riparian 

community became established after construction of upstream Glen Canyon Dam. The 

ultimate disposition of that community is largely unknown at present; however, short- and 

long-term changes in riparian vegetation along the river corridor are the subject of regular 

inventory and monitoring (Palmquist et al. 2018) and we are assured that the answer to what 

comes after Tamarix will eventually be known. 

What we do know now however, is that the Tamarix-dominated riparian community 

developed since Glen Canyon Dam became operational is now dramatically changing. It is 

possible, in time, that the OHWZ species (principally the mesquite and acacia) will continue 

to move downslope, as they have done for decades, and eventually replace Tamarix as it is 

reduced or eliminated. With the native OHWZ mesquites and acacias available to continue 

moving into the NHWZ, the Colorado River corridor in Glen and Grand Canyons may 

eventually be composed of mostly native species. They may provide an example, albeit 

rare so far, of the biocontrol effort realizing the potential that was originally anticipated 

by APHIS. At present, however, the decline of the Tamarix in Glen and Grand Canyons 

represents a loss of this unique naturalized habitat. Hence, we create a requiem to that 

declining community and the suite of wildlife species that were a part of it. Future research 

will document the development and replacement of an evolving community of vegetation 

and wildlife and hopefully the riparian habitat of the river corridor of the future is likely to 

be more dominated by native riparian vegetation species than it has been since Glen Canyon 

Dam altered the hydrograph. 

Thus, as we contemplate an answer to the question, “What have we lost and what 

have we gained as a result of the biocontrol?” – or, simply put – “Is Diorhabda-induced 

biocontrol an ecological disaster, or will it eventually lead to a more robust native riparian 

community?” – at this time we do not find a clear answer.
40 USDA Forest Service RMRS-GTR-411. 2020.

References 

Anderson, B.W. 2017. Death of an ecosystem: murder and autopsy of the Colorado River riparian 

ecosystem. AVVAR Books, Blyth, CA. 358 p. 

Anderson, B.W.; Ohmart, R.D. 1977. Vegetation structure and bird use in the Lower Colorado River 

Valley. In: Johnson, R.R.; Jones, D.A., tech. coords. Importance, preservation, and management 

of riparian habitat: a symposium. Gen. Tech. Rep. RM-43. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station: 23-34. 

Anderson, L.S.; Ruffner, G.A. 1987. Effects of the post-Glen Canyon Dam flow regime on the old 

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USDA Forest Service RMRS-GTR-411. 2020. 47

Chapter 3. Vanishing Riparian Mesquite Bosques: Their 

Uniqueness and Recovery Potential 

R. Roy Johnson, Tanner S. Carothers, and Steven W. Carothers

Introduction

The “mesquite bosque” (Spanish for “forest” or “woodland”), one of the most unique 

and productive southwestern riparian habitat types, was once far more abundant than it 

is today. Twenty-five years ago, Stromberg (1993), with a focus on Arizona, provided 

an excellent review on the ecology, decline, existing threats, and potential for recovery 

of these mesquite forests. By 1993 the iconic mesquite bosque riparian habitat was in 

serious decline, due primarily to anthropogenic activities. Stromberg (1993) observed that 

previous attempts at habitat restoration were of limited success and indicated that much 

of the significant bosque habitat loss was largely the result of human-induced changes in 

the biotic and abiotic conditions and processes in river floodplains specifically required by 

species of mesquite (Prosopis spp.).

In this chapter, we update elements of Stromberg’s 1993 review and provide a 

classification between two types of bosques based on distinct vegetation associations 

along a relatively dry to wet riparian continuum. We also discuss the uniqueness of 

mesquite bosques as wildlife habitat and chronicle the loss of some of the more distinctive 

of these forests in Arizona as well as conditions that led to their disappearance. Lastly, we 

suggest opportunities for a timely approach to mesquite habitat restoration that will likely 

arise as a result of recent litigation resolution between the Department of Agriculture and 

the Center for Biological Diversity and the Maricopa Audubon Society.

The Mesquite Bosque

Mesquite (Prosopis spp.) forests, or bosques, historically represented one of the most 

widespread of riparian communities in the Southwest. At one time these bosques occurred 

on floodplains, often elevated 2 to 15 m above a relatively stable water table (Stromberg 

1993) within riparian ecosystems. These ecosystems historically often spanned widths of 

several kilometers and stretched for tens of kilometers along the lower reaches of many 

southwestern rivers (Anderson and Ohmart 1977; Brown 1982; Rea 1983; Webb et al. 

2007, 2014). Today, surviving mesquite bosques are reduced to remnants of their former 

grandeur. 

The three native mesquite species found in association with bosques in the American 

southwest are bipinnately compound “legumes” that can occur as shrubs to very large 

trees. These species represent some of the more common facultative phreatophytes in 

the southwestern United States and northwestern Mexico. Honey mesquite (Prosopis

glandulosa, formerly P. juliflora) and velvet mesquite (P. velutina, also formerly classified 

as a variety of P. juliflora) occur as components of both riparian and upland ecosystems. 

A third species, screwbean mesquite or tornillo (P. pubescens), is more closely associated 

with riparian ecosystems and largely restricted to floodplains (Benson and Darrow 1981; 

Correll and Johnston 1970; Foldi 2014; Kearney and Peebles 1960; Stromberg 1993).
48 USDA Forest Service RMRS-GTR-411. 2020.

Historically, mesquite bosques provided a unique and valuable ecosystem for a variety 

of animal and plant species; however, in recent times many of these woodlands have 

been either eliminated or heavily reduced and fragmented by anthropogenic influences. 

Associated with loss and reduction of these bosques is a decrease in populations of many 

plant and animal species dependent upon this riparian type (Stromberg 1993; Johnson et 

al. volume 1).

The importance of mesquites has been demonstrated by the fact that, although there 

are only three native species in the United States, they are prevalent and important enough 

that several plants and animals have “mesquite” in their names. Two such eponymous 

grasses include curly mesquite grass (Hilaria belangeri) and vine mesquite grass (Hopia 

obtusa). Insects include the giant mesquite bug (Thasus neocalifornicus; Coreidae) and 

the mesquite borer, larva of the longhorned beetle (Placosternus difficilis; Cerambycidae). 

Vertebrates include the mesquite lizard (Sceloporus grammicus), mesquite warbler, 

also known as Lucy’s warbler (Oreothlypis luciae), and mesquite mouse (Peromyscus 

merriamii). 

A Classification of Mesquite Bosques

Riparian mesquites may occur in hydroriparian (perennial streams), mesoriparian 

(intermittent streams), or xeroriparian (ephemeral streams) situations (Johnson et al. 

1984a). Stromberg (1993) recognized that along typically ephemeral or intermittent 

streams, mesquite bosques are normally in association with Sonoran riparian scrub 

plant species (see Brown 1982). However, bosques found in association with perennial 

streams and rivers typically have more riparian vegetation assemblages, often including 

cottonwood and willow (Populus-Salix) trees. We suggest herein that bosques may be 

divided into two basic categories, Type A and Type B mesquite bosques. We base the 

classification on two factors: first, we readily agree with Stromberg’s (1993) initial 

observations on the different vegetative associations found in bosques within ephemeral/

intermittent versus perennial steamflow segments. We recognize the substantial 

differences in wildlife productivity between the bosque types. Clearly, the two types of 

bosques range along a riparian continuum from relatively dry to relatively wet. Type 

A bosques are normally found where non-perennial flow conditions exist. At the other 

end of the continuum, with perennial flow and generally wetter conditions, Type B 

bosques typically have an increased upper canopy cover of hydroriparian species such as 

cottonwood and willow. 

Type A bosques consist almost entirely of mesquites as the tallest woody plants, while 

Type B bosques have a relatively large number of other, often taller tree species. These 

categories are important because higher groundwater conditions, or more permanent flow 

in the stream in all or portions of Type B bosques, facilitate the addition of taller trees. 

This results in an increase in foliage volume and habitat structure. The increased foliage 

volume usually results in an increase in biodiversity, especially bird species (Anderson 

2017; Carothers et al. 1974; MacArthur and MacArthur 1961; MacArthur et al. 1962; 

Willson and Comet 1996). 

The greater ecological value of Type B bosques to breeding birds can be illustrated 

by the number of species recorded in each of the two classes of bosques. The only Type 

A bosque in its near-historical condition for which we find breeding bird records is along 
USDA Forest Service RMRS-GTR-411. 2020. 49

the San Pedro River near Mammoth, Arizona (Gavin and Sowls 1975). We also have 

breeding bird records for three Type B bosques—San Xavier, Ft. Lowell, and Blue Point 

Cottonwoods. In the three Type B bosques, the number of breeding species are: San 

Xavier 73, Ft. Lowell 70, and Blue Point Cottonwoods 65, respectively, compared to only 

45 breeding species for the San Pedro Type A bosque (Johnson et al. volume 1).

Type A bosques: In addition to mesquite, there may be a small number of scattered 

trees, e.g., cottonwoods (Populus spp.), willows (Salix spp.), Arizona walnut (Juglans

major), or other species, but not in sufficient density to break the homogony of a 

mesquite-dominated upper canopy. Other woody species, most often shrubs or smaller 

trees, may include acacias (Senegalia spp., Vachellia spp.), wolfberry (Lycium spp.), 

desert hackberry (Celtis ehrenbergiana), saltbush (Atriplex spp.), and numerous other 

species (Stromberg 1993). Examples of Type A bosques include the surviving large 

mesquite woodland along the San Pedro River near Mammoth, Arizona, and the 

“extirpated” Santa Cruz River arm of Komatke Thicket, also known as the New York 

Thicket (Rea 1983; Wigal 1973). 

Type B bosques: In addition to mesquite trees and the same basic shrub understory 

species also found in Type A bosques, several broad-leaf trees often occur that are 

commonly taller than the mesquites. These include cottonwoods, willows, net-leafed 

hackberry (Celtis reticulata), blue palo verde (Parkinsonia florida), velvet ash (Fraxinus

velutina), Arizona walnut, desert willow (Chilopsis linearis), and other species 

(Stromberg 1993). These trees may be scattered throughout the bosque, such as at Blue 

Point Cottonwoods on the Salt River. Or they may occur largely along the streamside as 

was the situation with: the now “extirpated” San Xavier Bosque, also known as the Great 

Mesquite Forest near Tucson; the Phoenix Bosque (Jacobs and Rice 2002); and the Yuma 

Bosque (Mearns 1907). 

Unfortunately, few studies were conducted in most of the larger bosques prior to 

their loss, and detailed accounts of their biodiversity focused primarily on studies of the 

avifauna. We know that species rare in the United States often occurred in these bosques. 

For example, three primarily Mexican (or neotropical riparian breeding bird species 

that may nest in Type B bosques) include the ferruginous pygmy-owl (Glaucidium 

brasilianum), tropical kingbird (Tyrannus melancholicus), and rose-throated becard 

(Pachyramphus aglaiae). Coincident with the loss of the Phoenix and Tucson bosques, 

the owl no longer occurs in those areas, but it formerly occurred north to the Blue 

Point bosque (Johnson and Simpson 1971; Johnson et al. 2000; Johnson et al. 2003), 

and in both the Phoenix (Breninger 1898) and Ft. Lowell bosques (Bendire 1892). The 

northernmost breeding record for the becard was in the San Xavier bosque, which is now 

completely gone (Webb et al. 2014). The kingbird is the only remaining of these three 

species that still nests along the San Pedro River among other spots throughout Arizona 

(Corman and Wise-Gervais 2005). 

Mesquite Resilience

Despite the widespread decline of mesquite bosques throughout their former range 

in the Southwest and Arizona specifically, mesquite species are well adapted to arid 

environments and are excellent candidates for habitat restoration where conditions allow. 

In riparian areas, mesquite seems to thrive when alluvial groundwater is between about 
50 USDA Forest Service RMRS-GTR-411. 2020.

2.5 and 12.2 meters. The higher the groundwater, the more robust is their growth form 

(Stromberg et al. 1992; Stromberg et al. 1993). Mesquite trees have a dimorphic root 

system, including a deep tap root that can reach alluvial aquifers as well as a system of 

lateral roots that can take advantage of precipitation and flood waters (Leenhouts et al. 

2006). Thus, mesquite species are capable of extracting water from both surface soil 

and deeper strata long after obligate phreatophytes like cottonwood and willow can no 

longer be sustained during severe droughts and/or falling water tables. These factors are 

especially important since a great deal of uncertainty currently exists on the future health 

of riparian habitats in the arid Southwest with climate change and tamarisk or saltcedar 

(Tamarix spp.) biocontrol (Carothers et al. this volume; Dixon et al. 2009). 

The relative resilience and deeper root system of mesquite may allow the species to 

replace declining cottonwood/willow gallery forests if climatic conditions are such that 

available groundwater can no longer sustain gallery forests of obligate hydroriparian and 

mesoriparian phreatophytes (Stromberg et al. 2012). We also suggest that one or more 

of the three mesquite species discussed above will be excellent candidates for riparian 

restoration in those suitable areas where Tamarix has been or will be eventually decimated 

by biocontrol (see McLeod volume 1).

Mesquites also have a shoot system adapted for high temperatures and low moisture 

conditions. It is widely accepted that microphylls (small leaves), such as those of 

mesquite species, with a reduced surface area, result in reduced evapotranspiration 

rates (Thoday 1931). This allows plants to survive in more arid environments than 

related species with larger leaves (megaphylls). In addition to microphylls, mesquites 

also regulate their stomata, pores used for gas exchange, by opening them wider when 

conditions are more mesic and closing or partially closing them to reduce water loss when 

more arid conditions prevail (Leenhouts et al. 2006). 

A less known adaptive strategy for surviving in a variety of hydrologic conditions 

occurs in the root system of mesquites. Research along the San Pedro River in 

southeastern Arizona has documented that mesquites have the ability to redistribute soil 

moisture (hydraulic redistribution) throughout the year to achieve maximum growth, 

with the direction of sap flow differing depending on environmental conditions (Hultine 

et al. 2004). For example, before the monsoon season, sap flow in the deeper reaching 

tap root of a mesquite is upward (hydraulic lift) during both day and night, and from 

there outward toward the periphery of the plant through lateral roots nearer the surface of 

the soil. After the start of the rainy season, when surface soil is moist, the plant’s water 

distribution pattern is reversed, and water moves toward the stem through lateral roots and 

then downward in the taproot toward the water table (Hultine et al. 2004; Leenhouts et al. 

2006). Thus, mesquites are well adapted to grow in both moist riparian zones where they 

can reach very large sizes, as well as relatively dry upland conditions where robust growth 

and tree size is relatively limited.

Riparian Mesquites

Riparian mesquites commonly occur on floodplains and terraces, often separated 

from the stream course by cottonwood-willow riparian gallery forests or mixed deciduous 

woodlands, such as sycamore-ash (Platanus-Fraxinus). On their more arid side, away 

from the stream course, bosques commonly grade into upland desertscrub or semidesert 
USDA Forest Service RMRS-GTR-411. 2020. 51

grassland (Brown 1982). Velvet mesquite trees once rivaled mature cottonwoods in size, 

reaching more than 4 feet in diameter and over 70 feet in height (Brandt 1951; Stromberg 

1993; Swarth 1905; see fig. 5). Such large size is at least partially because they are among 

the least disturbed woody species by floods of the major southwestern riparian trees, 

owing largely to their extensive root systems, thus attaining ages exceeding that of many 

other riparian trees (see Smith and Finch volume 1). At the time of early settlement by 

Euro-Americans, bosques with such large trees were historically widespread throughout 

the Southwest, especially along larger perennial rivers (Dobyns 1981); however, we know 

of no remaining trees approaching such large sizes. 

Although mesquites growing in riparian situations may have the same genotype 

as those occurring in adjacent upland situations, with more access to groundwater and 

nutrients, riparian mesquites attain much greater size than upland plants. Willard (1912) 

wrote of trees in the San Xavier bosque that “the mesquite trees are wonders of their 

kind. There were some whose trunks at the base scaled over four feet in diameter.” Later, 

Brandt (1951) wrote that “here there are, indeed, trees of heroic dimensions; the bole of 

one stately specimen that we measured reached a girth of 13 feet, 6 inches, and a diameter 

of more than 4 feet, 3 inches; while the height of another capitol-domed giant was 

calculated to be 72 feet.” 

In Arizona, the larger iconic bosques, especially along the perennial sections of the 

Salt, Gila, Santa Cruz, and Lower Colorado Rivers, are either gone or greatly reduced 

in size as second/third-growth trees. The best remaining examples of mature forests 

of velvet mesquite known to us are still found along several reaches of the San Pedro 

River where drought and declining water tables may eventually pose a threat to their 

long-term survival along usually dry desert watercourses. Mesquites often occur along 

Figure 5—Photograph taken in the mid1970s of Dr. Charles H. Lowe, one 

of the earliest riparian ecologists, 

standing next to a 72-foot tall, 2.5-

foot diameter at breast height velvet 

mesquite tree (Prosopis velutina) at the 

confluence of Sycamore Creek and 

the Verde River on the Fort McDowell 

Indian Reservation, Maricopa County, 

Arizona. (Photo by David E. Brown.)
52 USDA Forest Service RMRS-GTR-411. 2020.

margins adjacent to normally dry arroyos as a single line of trees along with species 

such as acacias, paloverdes, and ironwood (Olneya tesota) as well as a variety of other 

trees and shrubs (Simpson 1977; Stromberg 1993). Even mesquites of these xeroriparian 

ecosystems have been impacted by urban and rural development, especially by larger 

cities such as Phoenix and Tucson. Some of the larger remaining xeroriparian bosques are 

on protected lands such as along Growler Wash in Organ Pipe Cactus National Monument 

(Johnson et al. 1984b).

Compared to xeroriparian situations where one or two trees line a generally dry desert 

wash, more species of herbaceous riparian plants occur in Type A and B mesoriparian 

bosques that do not generally occur with xeroriparian mesquites. These include some 

relatively uncommon plants like hoary bowlesia (Bowlesia incana), Watson’s dutchman’s 

pipe (Aristolochia watsonii), and several grasses, e.g., Arizona brome (Bromus arizonicus) 

and vine mesquite grass.

Upland Mesquites

Within the past several decades, mesquite species have generally migrated out of 

the mesoriparian and xeroriparian areas and successfully invaded many upland habitats. 

The increase in mesquites in upland areas has been largely due to the natural adaptability 

of the species for surviving in a wide range of hydrologic conditions discussed above. 

Clearly, upland mesquites are the same species as riparian mesquites but have been able 

to invade the upland areas as grasslands have declined under grazing pressures or drought. 

Their spread into uplands has been aided by the foraging of cattle on mesquite seed pods 

in riparian areas. The seeds can then be carried by these cattle into adjacent grassland 

(Cable and Martin 1973). Examination of cattle droppings showed as many as 1,671 

mesquite seeds in a single cow-chip (Glendening and Paulsen 1955).

Loss of a Mesquite Bosque: Loss of an Ecosystem

The middle Santa Cruz River near Tucson, the middle Gila, and the Lower Colorado 

River from Davis Dam to the Mexican border once supported large mesquite bosques. 

The widespread loss of many of these once common bosques with trees of a far larger size 

and distribution than can be found today is well documented, especially for central and 

southern Arizona. Additionally, the extirpation of vertebrate species from portions of these 

river systems has been particularly well documented (Anderson 2017; Carothers et al. 

2013; Ohmart et al. 1988; Rea 1983; Rosenberg et al. 1991; Stromberg et al. 2004; Webb 

et al. 2014). 

Many, perhaps most, of the larger bosques once common along stream courses in 

Arizona have been “extirpated” through Euro-American settlement patterns (Dobyns 

1981; Webb et al. 2014). Stromberg (1993) lists numerous threats to mesquite bosques. 

The value of mesquite firewood is exceeded in the Southwest deserts only by desert 

ironwood and both of these species have been significantly reduced by wood-gathering 

for domestic as well as commercial use. Thousands of acres of mesquites were cut to 

provide fuel for steamships that navigated the Lower Colorado River during the last half 

of the 1800s and early 1900s (Lingenfelter 1978). Much of those mesquite woodlands 

in the Lower Colorado River Valley and along other major southwestern rivers were 
USDA Forest Service RMRS-GTR-411. 2020. 53

replaced by invasive Tamarix (Poff et al. 2011; Rosenberg et al. 1991; Stromberg et al. 

2007), which is now in decline as a result of biocontrol (Bloodworth 2016). Some bosque 

remnants are still extant along waterways for larger ephemeral and intermittent streams, 

though the woodlands are usually smaller and not as widespread as those that historically 

occurred along larger perennial and intermittent streams. Some of the best developed 

bosques remaining in Arizona occur along perennial streams with neither upstream dams 

nor urban developments, such as along the Gila River upstream from Coolidge Dam 

(Minckley and Clark 1984) and along the San Pedro River (Gavin and Sowls 1975; 

Stromberg and Tellman 2009). 

Perhaps the most severe losses of riparian mesquite ecosystems have been to 

screwbean mesquite. A recent study has shown that tornillo, or screwbean mesquite, has 

disappeared from 53 percent of the localities in which it was common more than a century 

ago (Foldi 2014). The decline of screwbean mesquite along portions of the Rio Grande 

in New Mexico and Lower Colorado River in Arizona has been as recent as the past 

few decades and this once dominant species can now only be found in isolated patches 

(Anderson 2017). In a recent study attempting to determine the cause of the decline of 

screwbean mesquite in the Lower Colorado River watershed, Madera (2016) concluded 

that some unknown pathogen is likely the cause of the species decline in that area. 

Vanishing Bosques

San Xavier Bosque: The Great Mesquite Forest, Santa Cruz River

Located on the Santa Cruz River, approximately 10 miles south of Tucson and 

adjacent to San Xavier del Bac Mission, this Type B bosque was called the Great 

Mesquite Forest because of its outstanding avian biodiversity and sheer abundance of 

birds (Brandt 1951; Webb et al. 2014). This once vast forest represents one of the best 

documented losses of an entire riparian/aquatic ecosystem in the Southwest (Webb et al. 

2014). Between the earliest journal records in the mid-1800s and its complete demise 

by the 1960s, it was the best-known mesquite woodland in the Southwest. Although 

numerous accounts of ornithological work in the bosque were published, little is known 

of its size at the time of Euro-American settlement. Willard (1912) spent “the whole 

day... skirting along the edge of the mesquite forest a distance of some six miles.” Brandt 

(1951) reported that it “bordered both banks of the Santa Cruz for a number of miles and 

appeared to be four or five miles in width at its broadest part, tapering back to the river on 

either side.” An aerial photograph taken in 1936 showed the bosque to be approximately 

7 square miles, but agricultural fields were already encroaching on it at that time (Webb 

et al. 2014). Biological reports by Euro-Americans that stopped at San Xavier del Bac 

Mission to camp along the Santa Cruz River became available by the mid-1800s. In 1850, 

Judge Benjamin Hayes of California wrote in his journal of this bosque that wolves “... 

were howling all around us, and one of very large size, was seen” (Davis 1982). In 1872, 

Major Bendire wrote of camping in this bosque on a trip from his post at Ft. Lowell in 

Tucson to the army outpost at Tubac (Bendire 1892). 

By the beginning of the 1900s, the Great Mesquite Forest became one of the bestknown sites for ornithological work in the United States. In addition to a few birds 

mentioned by Bendire (1892), the first attempt to record all the breeding birds of the 

bosque was in 1902 and 1903 (Swarth 1905; Webb et al. 2014). From Swarth’s first visit 
54 USDA Forest Service RMRS-GTR-411. 2020.

in 1902 into the early 1960s, several of the best-known ornithologists in the United States; 

including W.L. Dawson (1921) and A.C. Bent (1919-1958), visited and published lists 

of birds breeding in this bosque. However, by 1917, 2,500 cords of mesquite were being 

cut and removed annually from the San Xavier bosque by woodcutters (Dawson 1921), 

and by 1960, the bosque consisted mostly of second growth mesquites, approximately 15 

to 20 feet in height with none of the previously described giants remaining (R. Johnson, 

personal observation). The final death knell of the forest occurred during the 1970s, when 

construction of Interstate 19 bisected the remains of this once great forest (Webb et al. 

2014). It was also during the 1970s when decades of groundwater pumping for domestic 

water supplies by the city of Tucson finally drained the alluvial aquifer below the levels 

that once supported these grand mesquite trees. 

One of the unusual animals of the Great Mesquite Forest was the mesquite mouse, 

which occurs only in the States of Arizona and Sonora, Mexico (Hoffmeister 1986; 

Kingsley 2006) and is one of the rarest of native mice in the United States. The largest 

known population of this mouse lived in the Great Mesquite Forest before its destruction. 

In addition, by the start of the 20th century, several species of birds that had been 

recorded earlier along the Rillito River (Bendire 1872a, 1872b, 1892), a major tributary 

of the Santa Cruz River, had apparently been lost from the Great Mesquite Forest. These 

were largely riparian or wetland species, including killdeer (Charadrius vociferus), black 

phoebe (Sayornis nigricans), common yellowthroat (Geothlypis trichas), song sparrow 

(Melospiza melodia), gray hawk (Buteo plagiatus), and willow flycatcher (Empidonax 

trallii). Although the now endangered southwestern willow flycatcher (E. t. extimus) was 

known to breed along the Lower Colorado River during the late 1800s and early 1900s 

(Huels et al. 2013; Unitt 1987), it is not clear if the species was breeding or a migrant 

along the Santa Cruz River (Dawson 1921; Swarth 1905; Webb et al. 2014).

Komatke Thicket: New York Thicket, Confluence of Santa Cruz with Gila River

Located at the confluence of the Santa Cruz with the Gila River on the Gila River 

Indian Reservation, Komatke, or New York Thicket, was so named by local Pima Indians 

because it was so crowded with trees and wildlife (Rea 1983; Webb et al. 2014). It was 

said to be approximately 8 miles long from east to west and 3 miles wide from north 

to south (Webb 1959; Webb et al. 2014). Neff (1940) wrote that “in some places it is 

said to be fully six miles in diameter.” Records are inadequate to determine if this was a 

Type A or Type B bosque. Unfortunately, by the late 1970s, apparently about 90 percent 

of the mesquites were dead, killed by the lack of both surface water and groundwater 

(Rea 1983). In a 1970s photograph showing much of the bosque (Wigal 1973), there 

were no live mesquites and little more than dead stumps since most trees had been cut 

for firewood. Wigal (1973) called the once grand woodland “... 6,800 acres of a dying 

bosque.” At that time the Gila River had ceased to be free-flowing after the completion 

of upstream Coolidge Dam in 1928 and water flowed in the Santa Cruz only on rare 

intervals during upstream flooding (Webb et al. 2014). In addition, starting in 1902, deep 

wells were drilled on the Reservation and nearby, rapidly lowering the water table (Rea 

1983). By the late 1970s, live mesquite trees were only found at the edge of the bosque, 

where water drained from the bajada of the Sierra Estrellas, and in an ancient Hohokam 

canal over 2 meters wide (Rea 1983). Although there were studies of a large population 

of nesting white-winged doves (Zenaida asiatica) in the bosque (Neff 1940), in contrast 
USDA Forest Service RMRS-GTR-411. 2020. 55

to the Great Mesquite Forest, we find no general ornithological studies or other biological 

studies of any type from this large bosque before its demise.

Ft. Lowell Bosque, Rillito-Tanque Verde Creek

Ft. Lowell, a Type B bosque, was a commonly visited site by ornithologists looking 

for rare birds during the late 1800s and early 1900s. Although not as well studied 

ornithologically as the Great Mesquite Forest, we find 70 species of nesting birds reported 

for the Ft. Lowell area, only three species fewer than 73 species for the Great Mesquite 

Forest (Webb et al. 2014). Published ornithological studies at Ft. Lowell started earlier 

than those along the Santa Cruz River; we have records of birds from this area for three 

decades before records from the Great Mesquite Forest. The area was perhaps best known 

for the discovery of the first ferruginous pygmy-owls for the United States in “the heavy 

mesquite thickets bordering Rillito Creek, near the present site of Camp Lowell” (Bendire 

1872b, 1892). In addition to the owl, Bendire added another breeding bird of mesquite 

bosques (Bibles et al. 2002) to the regional avifauna, the gray hawk, and he also recorded 

the painted redstart (Myioborus pictus), a montane breeding species that occurs in 

mesquite bosques during migration. Bendire also added two species new to science from 

this area that occur in mesquite bosques: Bendire’s thrasher (Toxostoma bendirei) and 

rufous-winged sparrow (Peucaea carpalis) (Fischer 2001). 

After Bendire’s work in 1872-1873, the next report of birds for the Ft. Lowell area 

included several species collected by Frank Stephens in 1881 (Brewster 1882, 1883; 

Webb et al. 2014). Stephens was the second person to collect a ferruginous pygmy-owl 

in the area and he also collected the type specimen of a new subspecies of brown-crested 

flycatcher (Myiarchus tyrannulus magister) near Ft. Lowell (American Ornithologists’ 

Union 1957). Including the flycatcher, there were seven new subspecies of birds named 

from the Ft. Lowell bosque (Brandt 1951). By the start of the 1900s, groundwater 

withdrawal for the city of Tucson and surrounding Tucson Basin had resulted in a 

lowering groundwater table throughout the Basin (Webb et al. 2014). This resulted in an 

ever-increasing loss of mesquites from both Rillito/Tanque Verde Creek and the Great 

Mesquite Forest. Along with the loss of mesquites, several nesting riparian birds such as 

the gray hawk and willow flycatcher also disappeared. By the 1930s, Willis (1939) wrote 

that “the tree stand here [on the Rillito floodplain] is nearly pure mesquite and is probably 

a relic of the old forest.” Between the mid-1950s and 1980s, there was a continuing 

decline in mesquites along the Rillito floodplain, apparently due to urban development 

and decreases in both groundwater and surface flow. In 1988, Tucson increased its rate 

of withdrawal of groundwater, further stressing the remaining mesquites (Stromberg et 

al. 1992). Although the Ft. Lowell Bosque has survived longer than the Great Mesquite 

Forest on the San Xavier Indian Reservation, it is a shadow of its once great size. Urban 

encroachment and increasing groundwater withdrawal is a continuing threat to the 

remaining woodland.

Casa Grande Bosque, Gila River

Formerly located on the south bank of the Gila River, on an old floodplain (Judd et 

al. 1971) and adjacent to Coolidge, Arizona, the Casa Grande Type A bosque has been 

“extirpated” by groundwater overdraft and rural and urban development. Tree-ring dating 

aged three trees in this bosque between 110 and 137 years old (Judd et al. 1971). As with 
56 USDA Forest Service RMRS-GTR-411. 2020.

bosques at San Xavier and Komatke, a declining water table also had much to do with the 

loss of this habitat. The water table declined from a depth of 44 feet in 1923 to approximately 

100 feet in 1952 and 150 feet by 1960 (Judd et al. 1971). Mistletoe (Phoradendron spp.) 

infestation may have further hastened the death of mesquite trees (Judd et al. 1971; 

Stromberg 1993). With the completion of upstream Coolidge Dam in 1928, flow in the Gila 

River ceased, thus cutting off surface as well as groundwater to this bosque. Although there 

were later records of the ferruginous pygmy-owl from various localities along the Gila River, 

one of the earliest specimens of this rare tropical owl for the United States was taken here by 

Mearns on 10 May 1885 (Fisher 1893; Johnson et al. 2000).

Phoenix Bosque, Salt River Valley

The Salt River Valley of central Arizona is an example of a region that retains 

almost none of the area’s former wet riparian habitat or the mesquite bosques that were 

once widespread. Three of the State’s major rivers—Gila, Salt, and Verde—have all 

been dammed upstream from Phoenix to withdraw water for irrigation, municipal, and 

industrial usage in Phoenix and the surrounding area. This has resulted in desertification 

and dewatering of much of the stretches below the dams and inundation of potential 

riparian areas above the dams. In addition, several thousand square miles of xeroriparian 

ecosystems have been replaced in the Salt River Valley by urban and rural development. 

Phoenix Bosque is the name we have chosen for this Type B bosque that formerly lined 

much of the Salt River during the mid-1800s before it was entirely displaced by Phoenix. 

This large bosque covered approximately 25 square miles, extending northward for 

approximately 3 miles at its widest point, and from current Central Avenue westward for 

approximately 13 miles. A narrow band of cottonwood-willow grew immediately adjacent 

to the river with the mesquite on the large floodplain and river terraces to the north 

(Jacobs and Rice 2002). Phoenix was established amid this large bosque in 1868, and by 

1871, “98 blocks, each 300 feet long” were for sale (Barnes 1935, 1988). These lots each 

supported an average of six giant mesquite trees and early maps showed dense mesquite 

stands extending from the Salt River northward beyond Five Points (junction of Van 

Buren Street with Grand Avenue; Douglas 1938).

Before construction of dams on the Salt and Verde rivers, but long after the settlement 

of Phoenix began compromising the bosque, early irrigation canals and earthen ditches 

were commonly lined with cottonwoods and other species of trees, thus resembling 

natural watercourses (R. Johnson, personal observation). These artificial watercourses 

mimicked natural stream courses. Breninger (1898) wrote of the Phoenix area that “since 

trees planted by man have become large enough to afford nesting sites for woodpeckers, 

this Owl [cactus ferruginous pygmy-owl] has gradually worked its way from the natural 

growth of timber bordering the rivers to that bordering the banks of irrigating canals, until 

now it can be found in places ten miles from the river.” 

Yuma Bosque, Lower Colorado River Valley

At the turn of the 20th century, Edgar A. Mearns, a member of the U.S./Mexican 

Boundary Survey party, published his findings for this region and reported that the extent 

of the alluvial bottom land between Camp Mohave and Yuma was approximately 400,000 

to 500,000 acres in size. This was a Type B bosque with the river channel marked by a 

line of tall cottonwoods and a lesser fringe of willow. He also observed that the adjacent 
USDA Forest Service RMRS-GTR-411. 2020. 57

bottom lands upslope of the river were covered more or less with mesquite including 

screwbean mesquite (Mearns 1907). Mearns marveled that the size of some of the larger 

tree canopies covered an area of 50 meters or more in diameter. 

Since Mearns’ work, hundreds of thousands of acres in the Lower Colorado River 

Valley have been invaded by nonnative Tamarix. By 1986, a group of biologists working 

in the Lower Colorado River Valley estimated that the total riparian vegetation comprised 

only about 40,000 ha (88,000 acres). Thus, at that time approximately 22.5 percent 

remained of Mearns’ 1907 estimate. Of that 88,000 acres, 40 percent was covered by pure 

Tamarix with only 43 percent consisting of native plants mixed with Tamarix, and 16.3 

percent was covered by honey mesquite and/or native shrubs. Only 0.7 percent (307 ha) 

could be considered mature cottonwood or willow habitat. Thus, less than 15 percent of 

the original estimates of native riparian vegetation at the time of Mearns remained by the 

late 1900s, and much of that was mixed with nonnative Tamarix (Rosenberg et al. 1991). 

Unnamed Bosques

In addition to the above discussed bosques, others are occasionally mentioned in 

the literature without references to size or other factors. For example, Stromberg (1993) 

mentions several and Minckley and Clark (1984) mentioned a large mesquite bosque at 

Texas Hill on the lower Gila River, in the Wellton-Mohawk area. Minckley and Brown 

(1982) wrote of losses of mesquite and cottonwood forests along the lower Gila and 

Colorado Rivers without naming them or giving better locations and sizes. Historical 

documents are woefully incomplete in detailing the extent of the change affected on 

southwestern mesquite bosques. It is clear that the retreat of bosques is inextricably 

tied to the westward expansion of American settlers and subsequent ventures requiring 

large amounts of both water and land for agriculture and urbanization. The future of 

southwestern mesquite bosques is uncertain and full of challenges, but with effective 

conservation and restoration practices, opportunity exists to restore some areas with one 

or more of the three mesquite species. As stated above, mesquite species make promising 

candidates in the restoration of riparian corridors and springs, able to send roots deeper 

than cottonwoods and willows.

Perhaps the two best known surviving large bosques are a Type A bosque on the 

San Pedro River near Mammoth (Gavin and Sowls 1975) and a Type B bosque at Blue 

Point Cottonwoods, on the Salt River (Johnson and Simpson 1971; Johnson et al. 2000). 

Whereas mesquites found in these existing bosques are often smaller than those from the 

historical accounts, they still support a high degree of biodiversity. By examining extant 

bosques, one may extrapolate conditions that existed before the pumping of groundwater, 

the introduction of exotic competitors, and the large-scale clearing of mesquite for 

firewood and building material. From these case studies of historical and extant bosques, 

their distribution, extent, and size (per tree) has changed dramatically through the 

Anthropocene. Whereas any single threat has the potential to dramatically change the 

structure of bosques, these ecosystems are beset by polymorphic threats. 

The Future of Mesquite Bosques in the Shadow of Tamarisk Biocontrol

The remarkably successful biocontrol efforts to slow down invasion and/or eliminate 

the nonnative tamarisk (Tamarix spp.) shrub/tree throughout the range of the mesquite 
58 USDA Forest Service RMRS-GTR-411. 2020.

species (see McCloud volume 1; Carothers et al. this volume) has provided heretofore 

unavailable opportunities for habitat restoration of native riparian species, especially 

cottonwood and willow, but may eventually include mesquite. Potentially suitable 

conditions for cottonwood-willow and mesquite restoration are now available as the 

Tamarix declines in tens of thousands of acres. (However, see Carothers et al. this volume, 

on the potentially lasting inhibitory effects of Tamarix on reestablishment of native 

species.) The specific adaptations exhibited by mesquite (described above) for thriving in 

a variety of habitat conditions on a wet to dry riparian continuum allow for mesquite to be 

an obvious native riparian species for habitat restoration. 

Significantly, Federal funding for riparian habitat restoration has come available in 

2018 at unprecedented levels. As of June 2018, the Department of Agriculture is under a 

court order to fund the rehabilitation of riparian habitat to benefit the Federally threatened 

southwestern willow flycatcher in Arizona and New Mexico. The recently released United 

States District Court Remedial Order (Case No. 2:13-cv-1785-RFB-GWF), settling a 

lawsuit brought by the Center for Biological Diversity and Maricopa Audubon Society 

against the Department of Agriculture, is required to result in progress toward specifically 

restoring/creating cottonwood-willow habitat used for nesting by the flycatcher. 

Restoration of mesquite species as a buffer to the preferred habitat of cottonwood and 

willow could also benefit. The Court has ordered the Department of Agriculture to fund 

“intensive third-party riparian restoration efforts” in most of the living rivers and streams 

in Arizona, California, and New Mexico where the flycatcher was known to utilize or nest 

in the now-disadvantaged and dwindling Tamarix “bosques.” While the focus of the courtordered flycatcher habitat restoration is directed toward cottonwood-willow trees, buffers 

of mesquite trees around the former, as is found in Type B bosques, will be an obvious 

solution to long-term protection of plantings of the hydroriparian species. Whereas in 

many of the sites required by the court order for planting of cottonwood-willow habitat 

will not have the seasonal flooding necessary for recruitment of these hydroriparian 

species, the mesquite buffer restoration will persist and recruit once they are established.

Stromberg (1993) earlier recommended prevention of further degradation of 

mesquite bosques by removing the ongoing threats of recruitment inhibiting groundwater 

overdraft. She also recommended eliminating woodcutting, livestock grazing, and 

impacts originating from recreation. As far as restoration of mesquite habitat, Stromberg 

(1993) was far less sanguine; for example, she correctly cited one of our early papers 

(Carothers et al. 1990) as providing examples of high post-restoration mortality in 

largely failed efforts to reestablish mesquite. In the late 1980s and throughout the 1990s, 

habitat restoration science in riparian habitat was in its infancy. Those engaged in early 

attempts at restoration were often unaware of the dynamics between restoration success 

and a myriad of project-defeating biotic and abiotic factors. Not all potential restoration 

sites are equal and few are harbingers of rapid success. The previous occupation of a 

site by a particular plant species does not in and of itself offer predictions of restoration 

success. Even after evaluation of hydrologic and edaphic factors indicate restoration 

potential, beaver, elk, rabbits and pocket gophers can quickly decimate the efforts of an 

expensive, well-planned revegation project. It may take years for a restored site to become 

established and self-recruiting.
USDA Forest Service RMRS-GTR-411. 2020. 59

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USDA Forest Service RMRS-GTR-411. 2020. 63

Chapter 4. Using the Southwest Experimental Garden 

Array to Enhance Riparian Restoration in Response to 

Global Environmental Change: Identifying and Deploying 

Genotypes and Populations for Current and Future 

Environments 

Thomas G. Whitham, Catherine A. Gehring, Helen M. Bothwell, Hillary F. Cooper, Julia B. 

Hull, Gerard J. Allan, Kevin C. Grady, Lisa Markovchick, Stephen M. Shuster, Jackie Parker, 

Abraham E. Cadmus, Dana H. Ikeda, Randy K. Bangert, Kevin R. Hultine, and Davis E. Blasini

Introduction

The role of genetics in ecosystem restoration has largely revolved around the policy 

of using local genotypes (i.e., those sourced from areas near the restoration site), based on 

the logic that local plants are best adapted to local conditions (Johnson et al. 2004; Meffe 

and Carroll 1977). In a relatively stable environment, this is a sound practice; however, 

global environmental change impacts on the landscape are increasingly rendering this 

policy as inadequate at best and damaging at worst (Whitham et al. 2010). Here we define 

“global change” as ongoing changes in temperature, moisture, interactions with invasive 

plants and pathogens, increases in the frequency and severity of environmental extremes 

(e.g., fires, droughts), and other challenges (Cayan et al. 2010; Gutschick and BassiriRad 

2003; Jones and Monaco 2009). Because of rapid environmental change, plants that are 

locally adapted to current environmental conditions are likely to become increasingly 

maladapted to their changing native environments. For example, in Arizona, Grady et al. 

(2011, 2015) found that Fremont cottonwoods (Populus fremontii) were currently locally 

adapted along an elevational gradient, and that tree genotypes from lower elevations 

transplanted to higher elevations were predicted to outperform the local genotypes at 

higher elevations under climate change conditions. In regions of especially rapid change 

such as the American Southwest (Garfin et al. 2013; Seager et al. 2007), local populations 

are likely to lack sufficient genetic variation to adapt to these new environments (Aitken 

et al. 2008; O’Neill et al. 2008a). Similarly, with the rapid velocity of climate change 

coupled with fragmented landscapes (Loarie et al. 2009), many species cannot migrate 

fast enough to reach favorable environments (Aitken et al. 2008; Davis and Shaw 2001). 

Ignoring the reality of global change sets up restoration efforts for failure. Given 

that most restoration projects focus on foundation plant species that, by definition 

(Dayton 1972), shape the communities and ecosystems around them, discounting global 

change impacts to these foundation species will also lead to high biodiversity loss and 

altered ecosystem function. Failed restoration projects could also result in the loss of 

future restoration funding. Riparian and aquatic ecosystems support some of the highest 

biodiversity worldwide, and incorporating future climate change impacts in the restoration 

planning process is essential for successful preservation of these valuable habitats and 

the wealth of species they support. These problems are recognized as a global issue that 

affects riparian and aquatic organisms worldwide (Davies 2010).
64 USDA Forest Service RMRS-GTR-411. 2020.

Robust experiments are a key component of identifying which species, populations, 

and genotypes are most likely to survive both current and future environmental 

conditions. Just as agricultural field trials have been essential for crop improvement, the 

same basic field trial approach is being used for forest trees and other species in wildlands 

(e.g., Marris 2009; O’Neill et al. 2008a). In the American Southwest, we have established 

field trials of riparian cottonwoods and willows that are embedded within larger 

landscapes to be restored. We recommend such field trials as a first step in restoration to 

minimize future risks and enhance project success on adjacent lands. 

The Southwest Experimental Garden Array (http://www.sega.nau.edu) was specifically 

funded by the National Science Foundation to develop this approach. Established along 

a 1,400-meter elevational gradient, the SEGA gardens allow researchers and managers to 

understand predicted climate change impacts across five major vegetation zones in Arizona. 

Based on findings quantified in gardens (Evans et al. 2016; Grady et al. 2011, 2015) and in 

the wild (Ikeda et al. 2017), restoration biologists can deploy genotypes from populations 

that show adaptive potential and are most likely to survive in a changing environment. In 

so doing, both the foundation species and the diverse communities they support can be 

preserved for future generations (Ikeda et al. 2014). 

The following sections develop the logic of this approach. We: give evidence that 

global change results in maladaptation; emphasize the importance of focusing first 

on foundation plant species that are the community and ecosystem drivers; describe 

experimental approaches that can be used to identify adapted genotypes for use in 

restoration; and discuss the importance of maintaining high genetic diversity for the 

evolutionary potential and highest biodiversity of foundation species’ communities.

Local Adaptation in a Stable Environment and Maladaptation with Global 

Change

Many studies show that plants are generally locally adapted to their environment 

(e.g., Clausen et al. 1940; Hereford 2009; Joshi et al. 2001; Savolainen et al. 2007). Thus, 

the demonstration of local adaptation is a primary justification for using only local stock 

in restoration, which is the current practice for restoration within most national parks. 

However, climate change and other global change stressors such as invasive species, 

drought, and other extremes are causing dramatic shifts, such that locally adapted plants 

are becoming increasingly maladapted (Jones and Monaco 2009). This decoupling of 

plants from the environments they were adapted to is resulting in landscape-level dieoffs (Allan et al. 2010; Gitlin et al. 2006; van Mantgem et al. 2009) and a shift in the 

elevational and/or latitudinal distributions of species to environments that match the site 

conditions of their former home environments (e.g., Allen and Breshears 1998; Brusca et 

al. 2013; Chen et al. 2011; Walther et al. 2002). 

A mismatch between the environments that species are currently adapted to and 

what the environment has or will become in the future is illustrated in figure 6, showing 

a photograph of the Bill Williams River that is suffering from upwards of 85 percent 

mortality of Fremont cottonwood. This mortality is associated with an ongoing drought 

that began about 5 years prior to a high mortality event that first became evident in the 

summer of 2015. Concurrent with this mortality event was a lack of surface stream 
USDA Forest Service RMRS-GTR-411. 2020. 65

flows, a water table that had receded below 5 m, and an increase in extreme summer 

temperatures events ≥ 50 °C. These conditions have continued through the present 

(September 2017) and are associated with continuing mortality. High mortality of these 

foundation trees is of great concern since this forest is home to many sensitive and listed 

species (e.g., Lower Colorado River Multi-Species Conservation Program: LCR-MSCP 

2004). 

As such drought conditions persist or worsen into the future as species distribution 

models predict for the Sonoran Desert ecotype of P. fremontii (Ikeda et al. 2017), restoring 

with local stock would likely result in restoration failure for both the cottonwood and its 

many dependent species. Similar forecasts are predicted for many forest trees throughout 

western North America (Rehfeldt et al. 2006). Thus, experimentally identifying source 

populations and genotypes from non-local stocks that can survive both current and 

future conditions is crucial for restoration projects to be successful. Furthermore, when 

local plants are no longer locally adapted as figure 6 emphasizes, concerns about genetic 

pollution are largely negated. The focus should shift toward finding stocks that can 

survive in altered environments and simultaneously maintain the listed and sensitive 

species that depend upon this habitat for their survival (e.g., Bangert et al. 2013).

The Importance of Focusing on Foundation Species

Riparian foundation species such as cottonwoods and willows support hundreds 

of other organisms including sensitive and listed species (e.g., Bangert et al. 2013; 

Patla 2014; Skagen et al. 2005). Therefore, a first priority should be to identify the 

foundation species that play a central role in defining their respective communities and 

ecosystems and focus on their preservation as a first priority (Dayton 1972; Ellison et 

al. 2005; Whitham et al. 2003). Furthermore, many studies have shown that different 

plant genotypes support different communities and ecosystem processes (Crutsinger 

2016; Lamit et al. 2015; Whitham et al. 2012). The communities that these different 

genotypes support represent a heritable plant trait (Gehring et al. 2017; Keith et al. 2010; 

Shuster et al. 2006). Thus, it is essential that we consider preserving not only species, 

but also the intraspecific genetic variation within these species. High genetic diversity 

Figure 6—Photo of cottonwood riparian forest taken March 28, 2017, along the lower reaches of the Bill Williams National 

Wildlife Refuge in Arizona. Photo by co-author Hillary Cooper.
66 USDA Forest Service RMRS-GTR-411. 2020.

in foundation species drives higher biodiversity in their associated communities and 

enhanced ecosystem processes (e.g., Ferrier et al. 2012; Fischer et al. 2014; Wimp et al. 

2004). Thus, the problem of conserving riparian communities (and other vegetation types) 

is three-fold:

• Preserving foundation species.

• Preserving the genetic diversity of foundation species essential for maintaining 

their adaptive potential and role as a foundation species supporting diverse 

associated communities and ecosystem processes.

• Understanding the relative importance of genetics versus the environment and 

how they interact (G x E interactions) to affect foundation species, associated 

communities, and ecosystem processes. 

To achieve these three goals in a changing environment, we must identify the 

individual plant genotypes and populations that can survive both current and future 

environmental conditions for a given restoration site (Grady et al. 2011; Ikeda et al. 2017; 

O’Neill et al. 2008a) and identify the corridors that maintain gene flow among populations 

as landscapes become more fragmented (Bothwell et al. 2017) and may suffer from the 

loss of genetic diversity (Vranckx et al. 2012). By focusing on these foundation species 

that are community and ecosystem drivers, we can also save many other species that are 

dependent upon the critical habitat they provide. In the next section, we describe how we 

can rigorously and experimentally identify the populations and genotypes of foundation 

species most useful in restoration.

Quantifying Which Stocks Will Survive Future Environments

Experiments are essential for identifying which species, populations, and genotypes 

are most likely to survive both current and future environmental conditions. The basic 

experimental approach pioneered by Clausen et al. (1940) is to establish field trials on 

portions of lands that are slated for restoration. Because the findings from field trials are 

most valid for the biotic and abiotic environment in which they are measured, having 

these sites embedded in larger restoration sites ensures that the findings from the field 

trials have the same soils, communities, and other variables that are most relevant to the 

adjacent sites to be restored. Based on quantification of plant performance (e.g., survival, 

growth, biodiversity, community stability, ecosystem functions) from such field trials, 

managers can identify stocks that exhibit adaptive potential to changing environments 

so they can be used to restore adjacent lands. This approach maximizes both short- and 

long-term plant performance goals and reduces risk of failure. Field trials have long been 

a standard practice in agriculture for maximizing yield, but they are much less common in 

wildlands restoration that necessitate more complex metrics of success (e.g., biodiversity, 

ecosystem services). Long-term monitoring of these field trials is especially important 

with long-lived species, as genotypes that do best in the short term may exhibit higher 

relative mortality in the long term or vice-versa (e.g., Sthultz et al. 2009). 

Just as agricultural field trials have been essential to identify superior genetic lines 

to maximize crop yield under current and future conditions, the same basic field trial 

approach is being used with commercially important forest trees (e.g., Marris 2009; 

O’Neill et al. 2008a,b). As an example of the success of this approach, O’Neill et al. 

(2008a) developed a model using a network of provenance trials to match seed source 
USDA Forest Service RMRS-GTR-411. 2020. 67

populations of lodgepole pine with planting sites where they will perform best under 

future climatic conditions. Based on their findings, it is now reforestation policy in British 

Columbia to plant with trees adapted to a 2 °C hotter environment than the local stock 

(Marris 2009; O’Neill et al. 2008b; Pedlar et al. 2011), and these recommendations are 

being further refined to obtain even better matches with projected future environments 

(O’Neill et al. 2017).

Research in the United States on foundation tree species has greatly benefited 

from this field trial/common garden approach, exemplified in a new research facility, 

the Southwest Experimental Garden Array (SEGA). The provenance trial approach 

specifically involves collecting genotypes from each of several different populations 

throughout a plant’s range and reciprocally transplanting them in common garden test 

sites to quantify their performance in different environments. Northern Arizona University 

(NAU) has undertaken a long-term program to develop SEGA as a research platform to 

study and provide solutions for utilizing genetics-based approaches to mitigate climate 

change, invasive species, and other global challenges. This research platform (fig. 7) 

allows collaborating scientists and institutions (e.g., USGS, USFS, NAU) to quantify the 

ecological and evolutionary responses of species to changing climatic conditions using 

emerging technologies. 

SEGA was initiated in 2012 with $4.5 million in seed funding from the National 

Science Foundation (NSF) and NAU to create a network of 10 gardens along an elevation 

gradient in northern Arizona. The SEGA model is flexible, and a growing network of 

satellite sites are being added by users that expand upon and increase the resolution of 

the core sites. Because data are archived to NSF standards and are available online, this 

array is ideally suited to the experimental study of gene by environment interactions that 

can best be addressed using a common garden network. Study organisms span from soil 

microbes, grasses, forbs, shrubs, and trees. 

Figure 7—The Southwest Experimental 

Garden Array (SEGA) spans a steep 

elevational gradient of 1,400 m in 

northern Arizona that encompasses 

desert to aspen-mixed conifer 

vegetation types. Each site has 

irrigation to help establish plants 

and experimentally eliminate 

the effect of soil moisture with 

increasing elevation if desired. 

A weather station at each site, 

sensor arrays, live video, and other 

instrumentation allow researchers 

to monitor their plots remotely 

from their home institutions. 

Illustration by Paul Heinrich from 

the SEGA website http://www.sega.

nau.edu.
68 USDA Forest Service RMRS-GTR-411. 2020.

Because temperature and moisture predictably change with elevation, these common 

gardens arrayed along an elevation gradient reflect climatic differences from desert to 

aspen-mixed conifer forest biomes that mimic the effects of climate change. Thus, this 

elevation gradient allows researchers to simulate the effects of climate change. For example, 

by planting the same genotypes of plants in multiple gardens, especially at lower elevations 

where it is hotter and drier, plants can be challenged with expected future climatic 

conditions. Researchers can then identify genotypes that show adaptive potential consistent 

with survival in a changing environment for use in restoration. Similarly, by adding 

supplemental water at a site or removing water with the use of rainout shelters, researchers 

can quantify the combined effects of temperature and moisture gradients with even greater 

control, or quantify just the effects of temperature alone. Experimental separation of 

temperature and moisture effects is especially important since increasing temperature has 

been demonstrated in virtually all climate models, while precipitation predictions are more 

variable as some areas might get wetter or have a change in the distribution of precipitation. 

In combination, this experimental design allows researchers to identify optimal 

populations and genotypes that will perform best on adjacent restoration sites. 

Experimental designs also allow more complicated studies to include the examination 

of nurse plant effects and mycorrhizal mutualists (e.g., Gehring et al. 2017), the planting 

of co-adapted communities (i.e., multiple species co-evolved at the same site) planted 

and tested together versus randomized plantings of species from different sites (Grady 

et al. 2017), and planting of genotypes to promote biodiversity, community stability, and 

network structure (Keith et al. 2017). 

SEGA enables a new generation of genetics-based climate change research that 

allows researchers from diverse disciplines to quantify the ecological and evolutionary 

impacts of climate change on (1) foundation plant species, (2) their associated 

communities, (3) native-exotic species interactions, and (4) the ecosystem processes 

that emerge from these interactions. Additionally, the Northern Arizona Research 

Greenhouses support plant propagation for use in SEGA, and the Environmental Genetics 

and Genomics center (EnGGen) allows researchers to genetically characterize their target 

species and identify candidate genes for functional traits of special interest (e.g., drought 

and temperature tolerance, productivity). These integrated facilities greatly enhance the 

research potential of SEGA. 

In short, SEGA and related facilities allow unprecedented opportunities to address 

issues important in riparian systems, and other vegetation types, that can enhance 

restoration success for future climates and other global challenges. SEGA emphasizes 

interdisciplinary-interagency collaborations, which can help shorten the turn-around time 

between discovery and implementation by land managers.

SEGA sites have provided important findings that can be used by restoration 

biologists and managers to select high performing tree genotypes for a given restoration 

site under future conditions. For example, using trees collected from 14 source 

populations of Fremont cottonwood and grown at a Bureau of Reclamation riparian 

restoration site on the Lower Colorado River (fig. 8), Grady et al. (2011) showed that 

different genotypes and populations of Fremont cottonwood predictably varied in 

their performance when grown at the lower edge of their distribution where record 

temperatures can exceed 50 °C. From these studies, five major findings emerged:

(1) Transfer distance has an important impact on annual above-ground net primary 
USDA Forest Service RMRS-GTR-411. 2020. 69

productivity (ANPP). A transfer distance of 6.5 °C means that a high-elevation, cool-site 

population was transferred to the warmer, low-elevation field trial, and the greater the 

transfer distance, the greater the decline in ANPP (fig. 8). ANPP is important because it 

reflects the growth rate of individual tree genotypes, but it is also positively correlated 

with survivorship (Grady et al. 2015) and biodiversity (Ikeda et al. 2014; Stone et al. 

2010; Swaty et al. 2004). Comparing the same genotypes from different transfer distances 

further revealed that local populations had genetics-based leaf economic traits that made 

them adapted to the local site (Grady et al. 2013). 

(2) Lower elevation source populations are currently better adapted to temperatures that 

can exceed 50 °C than higher elevation populations where temperatures are cooler. This 

finding of local adaptation is supported by the fact that populations like the low-elevation 

field site on the Lower Colorado River (i.e., sites that have a long, hot growing season) 

achieve higher ANPP than populations from higher mountain populations (i.e., sites that are 

genetically adapted to frost and a shorter growing season; fig. 8). However, with ongoing 

and projected climate change, local stock is expected to become maladapted.

(3) The slope of the regression line shows the sensitivity of cottonwoods from the 

source populations used at the planting site to temperature (mean annual maximum 

temperature, MAMT) (figs. 8, 9). Foundation species characterized by high sensitivity 

(i.e., with a steep regression slope) are key candidates for conservation/restoration 

strategies because of their sensitivity to climate change. Less sensitive species with a 

shallow regression slope might receive lower priority (Grady et al. 2011). In other words, 

populations of species with a more shallow regression slope can be transferred greater 

distances without suffering a loss in productivity compared with populations of species with 

a steep slope. 

(4) The trait-based variance of growth is illustrated by standard error vertical bars in 

figure 8, which represents a population’s potential to adapt to climate change. Thus, for 

any population shown in this figure, if the predicted change in climate is less than the 

variance, then the population has the potential to adapt (evolve) to predicted changes. 

Such rapid evolution has been demonstrated in pinyon pine (Sthultz et al. 2009). After the 

record drought in 2002 that resulted in landscape-level mortality, the genetic composition 

of the survivors was significantly different than the populations prior to the drought; 

similar findings have been found with other plants in response to rapid climate shifts (e.g., 

Franks et al. 2007) and introduced mammalian herbivores (Smith et al. 2015). 

Populations with high genetic variation in climate change related functional traits 

allow rapid evolution to new conditions. However, if predicted climate changes exceed 

the genetic variation in a population’s climate change related functional traits, then that 

population is likely to die out once conditions exceed the ability of all members to cope 

with the hotter environment. In other words, the pace of climate change leaves little room 

for adaptation beyond existing genetic variation. Thus, in regions of the world where 

global change is highest, it would be a grave mistake to depend upon rapid evolution to 

solve this problem; once the trait variation in a population is exceeded (e.g., temperature 

or drought tolerance), the local population will likely be extirpated. Genotypes already 

adapted to a hotter environment represent the logical source of stock for long-term 

restoration in an altered environment. Thus, if the temperature of the planting site is 

projected to increase by 2 °C, source populations already adapted to sites that are already 

2 °C warmer than the planting site would be good sources of propagation stock.
70 USDA Forest Service RMRS-GTR-411. 2020.

Figure 8—At a low elevation site at the lower 

edge of the distribution of P. fremontii, 

mean Aboveground Net Primary 

Productivity (ANPP) per population is 

plotted as a function of the temperature 

transfer distance of the source population 

to the restoration site. A transfer distance 

of 0 °C indicates a source population from 

a hot site similar to the hot planting site at 

low elevation. A transfer distance of 6.5 °C 

indicates a cooler, high elevation mountain 

population transferred to the hotter, 

lower desert planting site and a mismatch 

in environmental conditions that each 

population is locally adapted to (adapted 

from Grady et al. 2011).

Figure 9—At a high elevation site nearer the upper, 

cold edge of the distribution of P. fremontii, 

mean Aboveground Net Primary Productivity 

(ANPP) per population is plotted as a function 

of the temperature transfer distance of the 

source population to the restoration site. A 

transfer distance of 0 °C indicates a source 

population from a cooler site similar to the 

cooler planting site at high elevation. A 

transfer distance of -10 °C indicates a hotter, 

low elevation population transferred to the 

cooler, higher elevation planting site and a 

mismatch in environmental conditions that 

each population is locally adapted to (adapted 

from Grady et al. 2015).

The problem of identifying which populations to use in restoration is especially grave 

at the lower edge of the species’ distribution, because there are no other populations that 

are currently known to exist in even hotter environments. In other words, managers may 

have few options for finding individual genotypes or populations that are already adapted 

to what the environment will become in the future. This suggests that efforts to restore 

cottonwoods at the lower edge of their distribution are likely to fail with further climate 

change that exceeds the physiological tolerances of the species that would be reflected 

in widespread mortality and recruitment failure. In such cases, researchers and managers 

should use field trials to identify and experimentally deploy the most functionally 
USDA Forest Service RMRS-GTR-411. 2020. 71

equivalent species that is adapted to an even hotter environment to potentially fill the 

functional role of cottonwood as a foundation species (e.g., perhaps mesquite). Although 

not a preferred solution, this may be the best option available to achieve some restoration 

goals of supporting community biodiversity and providing ecosystem services. 

(5) While the options for lower elevation edge populations facing climate change may 

be limited due to a lack of genetically appropriate stock, higher elevation populations 

have more promising options, because there are numerous populations at lower elevations 

that are already adapted to a hotter environment. For example, in another field trial study 

conducted near the upper end of the elevational distribution of Fremont cottonwood, 

Grady et al. (2015) demonstrated local adaptation in which high-elevation populations 

outperformed low-elevation populations in the high-elevation garden (fig. 9). With 

projected climate changes, these high-elevation populations would also be expected to 

become maladapted to the new environment. Because numerous populations live at lower 

elevations with a hotter environment, managers have the opportunity to select among 

these populations to find the one(s) that are the best candidates to plant at a higherelevation site undergoing climate change, by considering both transfer distance that 

equates with projected warming as well as similarities in site characteristics (e.g., soils). 

Using the information in this graph, we can select source populations for a predicted 

level of climate change. For example, if temperature is predicted to increase by 3 °C at the 

high-elevation planting site, using the empirical data from this graph, populations from 

lower elevation sites that have evolved for a 3 °C hotter environment should do best in 

the hotter future environment of the high-elevation planting site (see -3 °C populations 

indicating a hotter to cooler temperature transfer distance of 3 °C). Given our findings, 

we recommend a maximum 3 °C transfer distance to balance success between current 

and future climates. While trees from greater transfer distance would be better adapted to 

even hotter environments, trees would suffer from productivity losses in the current cooler 

environment. However, transfer distances of ≤ 3 °C would suffer minimal productivity 

losses in the current environment and would perform best in a 3 °C warmer future 

environment. If climate change is projected to increase by 6 °C, a phased or stepwise 

approach over time is recommended, in which increasingly more temperature tolerant 

genotypes and populations are planted as the climate continues to change.

In combination, these findings provide powerful tools to assist managers in the 

selection of genotypes and source populations that can survive a given level of climate 

change. Based upon the above and other studies, we recommend planting a combination 

of local stock for the current environment and stock from sites that are 1-3 °C hotter than 

the planting site, which would perform best in predicted future climates. The combination 

of these genotypes and populations provides important genetic variation that selection can 

act upon as climate change and other stressors determine who wins and who dies. Such 

strategies are especially important in regions where global changes are great such as the 

arid Southwest. 

Insistence on using only local stock—where the mismatch of the local populations 

and genotypes with the new environment are great as in figure 6—greatly diminishes the 

probability of successful long-term restoration. We emphasize that field trials should be 

embedded in the sites where extensive restoration is proposed because they standardize 

soil conditions, community species pools, and other factors that could confound planting 

strategies based upon findings from sites farther removed from the actual restoration site.
72 USDA Forest Service RMRS-GTR-411. 2020.

The Importance of Maintaining Genetic Diversity for Associated 

Communities

We emphasize that restoration plantings must maintain high genetic diversity both for 

the foundation species as well as their associated communities. Genetic variation provides 

the essential building blocks for natural selection to act upon (Fisher 1930), thereby 

allowing species to evolve in response to unforeseen challenges including future climate 

change, new invasive species, and currently unknown biotic and abiotic effects. 

For example, Sthultz et al. (2009) documented rapid evolution in pinyon pine, Pinus 

edulis, in response to record droughts, in which drought-tolerant genotypes suffered only 

21 percent mortality, whereas drought intolerant genotypes from the same population 

suffered 68 percent mortality resulting in a rapid shift in the genetic structure of the 

population. Without such genetic variation naturally occurring in the population, the 

impacts of drought could have been much more severe, resulting in a major bottleneck 

event and potential local extinction had all genotypes been drought intolerant. These and 

other studies clearly demonstrate the importance of maintaining high genetic variation in 

foundation species populations, even when the roles of genetics in community diversity 

and the environmental conditions likely to favor certain foundation species traits are 

poorly understood. 

For example, Orians and Fritz (1996) showed that under one set of environmental 

conditions, one group of willows was found to be resistant to insect herbivory, but under 

another set of environmental conditions, another group of willows was most resistant to 

insect herbivores. Similarly, antagonistic pleiotropy refers to a single gene controlling 

many traits, with at least one trait conferring positive fitness effects and at least one 

resulting in negative impacts to the organism (Yanchuk et al. 2011). For example, a 

plant may exhibit suboptimal performance under normal environmental conditions, but 

outperform the “optimal” genotypes under conditions of extreme stress. Planting only the 

most productive genotypes tested under “normal” conditions could result in large losses 

under drought conditions. In other words, high genetic variation is an insurance policy 

against unknown future biotic and abiotic conditions.

In addition to high genetic diversity being crucial for the long-term survival of 

foundation and other species, many studies have also demonstrated that high genetic 

variation in foundation species promotes greater biodiversity in their associated 

communities (e.g., Ferrier et al. 2012; Crutsinger et al. 2006; Wimp et al. 2004). A 

primary reason for this relationship between genetic diversity in the foundation species 

and the diversity of the associated community is that different genotypes support different 

communities. Because many plant traits such as phenology, phytochemistry, growth, 

and morphology have a strong genetic component and can covary and interact with each 

other (Endler 1995), the combination of all these traits is referred to as the multivariate 

phenotype (Holeski et al. 2012). As this multivariate phenotype differs greatly among 

individual genotypes, it is not surprising that different organisms and communities are 

found on different plant genotypes and that interspecific indirect genetic effects (IIGEs) 

among species in a community further lead to differences in community assembly on 

different plant genotypes and evolution (Allan et al. 2012). 
USDA Forest Service RMRS-GTR-411. 2020. 73

Because these multivariate plant traits and interactions are genetically based and 

passed to their offspring, the communities they support represent heritable traits that 

can be quantified (Keith et al. 2010; Shuster et al. 2006). Numerous studies of plants 

around the world from the tropics to the tundra and desert to alpine environments exhibit 

community heritability in which related individuals tend to support the same communities 

(review by Whitham et al. 2012). Numerous studies with cottonwoods have demonstrated 

the importance of individual genotypes supporting different communities of diverse 

organisms including mycorrhizal mutualists, decomposing soil fungi and bacteria, lichens, 

twig endophytes, and insects that are closely tied to the multivariate phenotypes of 

individual plant genotypes (Lamit et al. 2015). 

Studies of other systems such as conifers and eucalypts have shown similar patterns. 

Gehring et al. (2014) showed that different genotypes of pinyon pine (Pinus edulis) 

supported different mycorrhizal communities, and Barbour et al. (2009) showed that 

different genotypes of blue gum (Eucalyptus globulus) supported different communities 

of arthropods and fungi. These genetic effects even extend to beavers (Bailey et al. 

2004), elk (Bailey et al. 2007), nesting birds (Martinsen and Whitham 1994) and seedeating birds and mammals (Christensen and Whitham 1993). Thus, if maintaining high 

biodiversity is desirable, managers can select for a combination of different genotypes and 

populations that support the greatest biodiversity. 

Another important principle is the concept of managing foundation species for 

both genetic and community connectivity. Genetic connectivity considers the degree 

to which populations of foundation species share genes and associated alleles across 

their distribution. Similarly, community connectivity considers the degree to which 

populations of foundation species share communities across their distribution, which is a 

concept supported by the genotypic-community associations described above. Thus, it is 

critical to understand the biotic and abiotic factors that either facilitate or inhibit genetic 

connectivity in foundation species and how it may change over time in conjunction with 

environmental change. 

Cushman et al. (2014) and Bothwell et al. (2017) recently demonstrated the utility 

of this approach via the identification of genetic corridors in two widely distributed 

foundation species, P. fremontii and P. angustifolia. Like many cottonwood species, both 

P. fremontii and P. angustifolia support large and diverse communities of organisms 

whose connectivity may depend, in part, on the maintenance of genetic connectivity in 

their respective foundation species. Although such studies are in their infancy, “managing 

for connectivity” in both foundation species and their associated communities constitutes 

a “best approach” for preserving biodiversity into the future.

In combination, maintaining high genetic diversity in restoration plantings is crucial 

to both the survival of the foundation species and the diverse communities they support. 

These findings illustrate why it is so important for restoration biologists to maintain high 

genetic diversity in plantings, not only to support evolutionary potential in the trees, but 

also to support a diverse community of organisms. Experimental field trials (common 

gardens) represent a robust tool that allows researchers to quantify these relationships 

and develop a truly “adaptive management” strategy that is based on both ecological and 

evolutionary principles.
74 USDA Forest Service RMRS-GTR-411. 2020.

Summary

Despite their intuitive appeal, ecological restoration policies encouraging the use of 

locally derived stock may be misguided in the face of climate change, invasive species, 

altered flood regimes, fire, and other biotic and abiotic challenges. Rapidly changing 

environmental conditions are likely to favor plant traits that are distinct from the historical 

conditions in which local species have evolved. Failure to recognize these crucial 

environmental differences can lead to long-term failure of restoration projects. While 

the focus of this chapter has been on riparian restoration, global change is a universal 

challenge and the same principles presented here apply to all ecosystems facing similar 

challenges worldwide. 

We propose the following stepwise restoration approach to mitigate the impacts of 

global change: 

(1) Focus first on regions that are recognized as undergoing rapid environmental 

changes such as riparian habitats in the American Southwest. 

(2) Identify the foundation plant species that are recognized as the drivers of their 

respective ecosystems (e.g., cottonwoods, willows, and other dominant plants on the 

landscape). This is relatively easy to do as there is a wealth of published data and 

knowledge about the species that have the greatest functional roles in defining their 

respective ecosystems. As we continue to become more knowledgeable about these 

ecosystems in identifying hidden players such as endophytes, fungal mutualists, and 

pathogens (e.g., Bailey et al. 2005) that also play foundational roles, they can be added 

into more sophisticated studies. 

(3) On Federal, State, and private lands where change is imminent or has already 

occurred, such as the Little Colorado River that is inundated by invasive tamarisk 

(Tamarix spp.) and camelthorn (Alhagi maurorum), field trials should be established to 

determine which populations and genotypes can survive in these altered environments. 

These plantings will be most effective if they comprise the range of genotypes present 

throughout the species’ range, especially from hotter, drier climates that have similar 

hydrology, soils, and invasive species. 

(4) Based on these field trials, a diverse set of adapted genotypes and source 

populations can be identified and propagated for use on adjacent lands that can best 

survive both current and future environmental conditions. 

(5) Because different genotypes support different communities, individual genotypes 

and source populations can also be selected that support high biodiversity, community 

stability, and connectivity. Thus, as supporting high biodiversity (especially of sensitive 

and listed species) becomes increasingly important, managers can move from restoring 

at the level of species to restoring at the genotype level. They do this by using specific 

plant genotypes that are preferred by community members of special interest such as 

the southwestern willow flycatcher (Bangert et al. 2013). However, managers must be 

careful to avoid selecting only optimal genotypes for a specific purpose. As noted above, 

sometimes the suboptimal genotypes under one set of conditions outperform the optimal 

genotypes under another set of conditions.

The Southwest Experimental Garden Array funded by the National Science 

Foundation was specifically established to achieve these goals of incorporating genetics 

approaches in land management with the major vegetation types on the Federal, State, 
USDA Forest Service RMRS-GTR-411. 2020. 75

and private lands that the array encompasses. We emphasize that such genetic approaches 

have long been used in agricultural crops in which yield is the major desired outcome. 

The goals of wildland restoration generally emphasize creating native habitat to support 

high biodiversity are different than agricultural fields. But the genetic principles are the 

same, and land management can be improved using such methods. In the face of great 

environmental challenges, genetics-based approaches are cost effective and represent a 

key step forward to help mitigate the impacts of global environmental change.

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Chapter 5. The Watershed Continuum: A Conceptual 

Model of Fluvial-Riparian Ecosystems 

Lawrence E. Stevens, R. Roy Johnson, and Christopher Estes

Mountains of music swell in the rivers, hills of music billow in the creeks, and meadows 

of music murmur in the rills that ripple over the rocks. Altogether, it is a symphony of 

multitudinous melodies. All this is the music of waters ... sounds that span the diapason 

from tempest to tinkling raindrop, from cataract to bubbling fountain.

John Wesley Powell 1895: 394, 397

Introduction

Terrestrial fluvial-riparian ecosystems (FREs) are riverine landscapes that integrate 

aquatic, riparian, and upland domains within watersheds, linking physical, biological, 

and cultural-economic processes (Tockner et al. 2002). FRE characteristics and processes 

intergrade through the entire watershed, from the river’s headwaters to its terminus into 

an endorheic basin or the sea, and can extend far out into the submarine environment 

(e.g., Canals et al. 2009; Vannote et al. 1980; Ward et al. 2018; fig. 10). FREs include 

all sources of water that contribute to the basin’s riverine ecosystem, including springs, 

surface runoff, lakes, and atmospheric sources, such as humidity and fog. Only an average 

of 2,120 km3

 (0.0002 percent) of the world’s water exists in river systems at any given 

time (Shiklomanov 1993). 

While rivers process only a tiny fraction of the Earth’s fresh water and occupy only 

a minute proportion of the Earth’s terrestrial surface, FREs are highly productive and 

ecologically interactive, often supporting diverse, densely packed biotic assemblages 

across short to vast time scales (Behrensmeyer et al. 1992; Sabo and Hagen 2012). 

Elevated biodiversity in FREs is linked to, and influenced by, many factors, including 

hydrogeomorphological, ecotonal, and shifting habitat mosaic effects (Gregory et al. 

1991; Naiman and Décamps 1997; Naiman et al. 1987, 2005). Humans rely on FREs, 

and our species’ evolutionary history and modern demography clearly demonstrate that 

reliance. As human domination of the Earth has progressed, rivers have been subjected 

to a host of anthropogenic alterations, including resource extraction, groundwater 

withdrawal, flow diversion and regulation, water quality degradation, and introduction 

of nonnative species. The need to integrate understanding of groundwater, fluvial, and 

lacustrine interactions within basins, and to sustainably manage rivers, has never been 

more urgent (Famiglietti 2014).

The natural dimensions and human impacts on FREs have stimulated a rich history 

of basic research, generating a vast literature, and prompting initiation of many local to 

international riparian research and stewardship organizations. Diverse ecohydrogeological 

models have been synthesized, tested, and predictively applied to FREs stewardship. 

Although sometimes biased by the expertise of the authors, these syntheses have 

advanced the collective understanding of FRE ecology. However, much remains to be 

learned. With the growing anthropogenic pressures on rivers and associated riparian 

ecosystems, we require an expansive, hierarchically organized conceptual framework to 

generate advanced understanding and management of FREs. 
USDA Forest Service RMRS-GTR-411. 2020. 81

Figure 10—The hydrological cycle, 

emphasizing stream-order 

transition and lateral processes 

within watersheds between 

constrained, alluvial, and lentic 

reaches. Numbers are percent of 

Earth’s water in different settings 

(Shiklomanov 1993). Illustration by 

Lawrence E. Stevens.

Here, we introduce the watershed continuum model (WCM) to expand 

interdisciplinary collaboration in ecological resource watershed system science and to 

improve science-based natural resource stewardship. The WCM describes matter, energy, 

and socio-values subsidy exchange within an entire FRE basin through the physical, 

ecological, and cultural processes that influence ecosystem geomorphology, biota, and 

society, across four-dimensional temporal and spatial scales. FRE connectivity extends 

from groundwater emergence and surface-derived flow, across the watershed, to the 

mouth, into the receiving basin or sea, and into the atmosphere. This integrative approach 

has been emphasized by Annear et al. (2004, 2009), Hynes (1975), Stanford (1998), 

Ward (1989), and others to interrelate physical hydrology, geochemistry, geomorphology, 

sedimentology, and aquatic and riparian domain linkage. We further advance this 

framework by linking aquatic and biological domains across stream order within 

watersheds (sensu Horton 1945; Strahler 1957) and over time. 

We restrict our discussion in this chapter to natural, unmanipulated FRE processes 

and characteristics, independent of the many anthropogenic impacts affecting FRE 
82 USDA Forest Service RMRS-GTR-411. 2020.

ecology. Anthropogenic impacts on rivers are widespread and merit much more scientific 

and societal attention, but we focus on natural processes and components to clarify and 

describe the interdisciplinary integration of hydrogeology, geomorphology, ecosystem 

ecology, and evolution. These components collectively interact within the watershed 

to generate FRE processes, assemblage structure, and stewardship decision-making. It 

is impossible here to do full justice to the broad array of FRE characteristics, themes, 

and models that contribute to this complex view; rather, we seek to provide a synthesis 

of patterns and insights, and to identify information gaps that affect FRE ecology and 

evolution. Collectively, the WCM couples multidirectional, trans-temporal material, 

and energy interchange from uplands to fluvial habitats throughout the basin, ultimately 

influencing the evolution of both aquatic and riparian FRE biota, as well as human 

cultures. 

Although our focus is global, we draw on examples from western North America, 

especially the Colorado River basin in the American Southwest where we have done 

much of our work. We begin this chapter by summarizing basic fluvial-riparian concepts, 

and then we describe the physical, biological, and ecological elements that provide the 

basis for the WCM. Next, we review and discuss biologically focused FRE models using 

illustrated schema. Finally, we discuss issues that remain understudied or unresolved. 

Given the focus of this publication, and because integration of aquatic and terrestrial 

domains across scales within FRE ecology is complex and incomplete, we add additional 

emphasis on the riparian domain. We recommend further considerations and actions and 

discuss conceptual model dimensions and limitations.

We present the WCM to enhance ecological resource watershed system science and 

improve natural resources management. Full predictive capacity of the WCM is not likely 

to be soon forthcoming, because some components have not been measured, numerically 

modeled, or assessed in relation to process interactions. Nonetheless, we welcome 

insight and collaboration from all who are interested in expanding and refining this 

integration, and we hope to stimulate further research needed to advance FRE ecology 

and stewardship.

Physical FRE Elements and Processes

Overview

Many supporting models have been proposed to describe the array of physical 

processes associated with FREs. FREs are terrestrial dendritic surface-water flow paths 

transporting matter and energy downslope through their channels within watersheds, with 

flow contributed by groundwater and multiple surface water sources. Riverine processes 

and geomorphology vary in relation to the geologic setting, gradient, flow, channel bed 

material, and sediment loads (Hey et al. 1982; Leopold and Mattuck 1953; Li et al. 1976; 

Schumm 1985; Tinkler and Wohl 1998; Wohl 2010). 

We initiate this discussion with a brief overview of physical flow and 

hydrogeomorphological models (HGMs). Early work in fluvial hydrogeomorphology 

focused on classification and process-based modeling at microsite to basin-wide spatial 

scales, and on understanding and describing the interrelationships between basin 

characteristics and climate-based flow on channel form, function, and development. These 
USDA Forest Service RMRS-GTR-411. 2020. 83

HGM topics remain under intensive study and debate, particularly in relation to physical 

FRE management of rivers (e.g., Leopold et al. 1964; Leopold and Maddock 1953; 

Lewin 1978; Morisawa 1968; Schumm 1985). More recent emphases have shifted to 

tributary- and reach-based HGMs, with growing recognition that channel geomorphology 

and function depend on the extent of geologic constraint (Bellmore and Baxter 2014; 

Montgomery 1999; Sabo and Hagen 2012). In some situations, bio-ecological processes 

may reciprocally influence FRE geomorphology (e.g., Beschta and Ripple 2006). 

We follow this literature review with a description of how these elements relate in an 

integrated watershed approach. 

Fluvial Hydrogeomorphology

While trophic cascades can influence fluvial geomorphology, study of “bottomup” physical processes has stimulated a long history of HGMs and much ongoing 

effort to improve understanding and description of FREs. The initial foci of HGMs was 

on understanding channel landform organization and distribution and the processes 

that generate them (e.g., Buffington and Montgomery 2013; Knighton 1998; Rosgen 

1994, 2008). HGMs were primarily considered in unconstrained alluvial rivers, but 

Rosgen (1996), Tinkler and Wohl (1998), Wohl (2010), and others extended channel 

formation description to constrained and bedrock-dominated settings. Additional focus 

has been placed on the extent of groundwater-dependence (Eamus and Froend 2006). 

The Committee on Riparian Zone Functioning and Strategies for Management (2002), 

Giller and Malmquist (1998), Malanson (1993), and Naiman and Décamps (1997) 

also emphasized the importance of lentic habitat (river-related lakes and wetlands) 

relationships to rivers. That discussion was expanded by the Instream Flow Council 

(Annear et al. 2004, 2009) to include variability in lateral and vertical flow in relation to 

instream flow assessment and management strategies, particularly for salmonid fish.

Discipline-based or regional specificity has limited the applicability of some HGMs. 

For example, Bennett and Simon (2004) focused on engineering-based analyses of 

riparian ecology, without considering the implications of evolutionary ecology, as 

discussed by Steiger et al. (2005) and Stromberg et al. (2004). Harvey and Gooseff 

(2015) recognized this difficulty and emphasized integration of physical and ecological 

models. However, increased interdisciplinary dialogue is needed to further integrate 

hydrogeomorphic models with fluvial engineering, aquatic and riparian ecosystem 

ecology, as well as socio-cultural, economic, and evolutionary models (e.g., Lubinski 

1993).

The process domain concept (Bellmore and Baxter 2014; Montgomery 1999) shifted 

attention to FRE structure and process from the spatial scale of the basin to that of 

reaches. Alluvial reaches often support broader riparian zones and filter, store, and process 

organic matter from upstream (Nilsson and Svedmark 2002). In contrast, constrained 

reaches function more as conduits for material transport. At a coarse scale, Bellmore and 

Baxter (2014) compared aquatic-riparian organic matter dynamics in five confined versus 

five alluvial reaches, reporting that the former had twice the allochthonous (extrinsic) 

organic input, but had a reduced capacity to retain and process that material. As an 

across-scale approach, the process domain concept posits that spatial variation among 

geophysical processes shapes the disturbance regime, ecosystem structure, and ecosystem 

dynamics. This approach resolves some of the limitations of the river continuum 
84 USDA Forest Service RMRS-GTR-411. 2020.

concept (Vannote et al. 1980; below) that were attributed by Minshall et al. (1983) to the 

influences of location-specific lithology and geomorphology.

FRE structure is strongly dominated by physical processes operating in a bottomup fashion to shape riverine landforms and the habitats on which aquatic and riparian 

assemblages develop (figs. 10, 11). The geologic context and climate of the watershed 

control basin position, geomorphology, topography (elevation), configuration, and the 

supporting aquifers are physical state variables. Collectively, those physical state variables 

regulate flow, hydrography, water quality, and sediment transport. They also generate the 

FRE templates that are organized through hierarchically flow-linked reaches (Fisher et al. 

1998; Montgomery 1999; Vannote et al. 1980). Ecologically, nearly all terrestrial FREs 

exhibit four-dimensional subsidy exchange, including: (1) upslope eolian and zoochorous 

transport processes; (2) capillary rise of groundwater through fine sediments; (3) lateral 

and vertical transport through the watershed, often with subsurface or tidal influences at 

the confluence of tributaries, in parallel flow with head from adjacent rivers, in lowermost 

reaches, and into the atmosphere; and (4) artesian groundwater or, in very low-gradient 

reaches, downstream tributary flooding that initiates upstream-directed flow (Ward 1989; 

figs. 10, 11).

Figure 11—Conceptual fluvial-riparian 

ecosystems model depicting interactions 

among independent and dependent 

physical and ecological variables 

and processes, across stream order (0 

headwaters to X- mouth) and time (T1 to 

TX). Black arrow points indicate relative 

impacts of tributaries of different sizes.
USDA Forest Service RMRS-GTR-411. 2020. 85

A comprehensive FRE conceptual model must integrate these dimensions over time 

among aquatic, riparian, and upland domains within the watershed. It also must link 

independent and dependent physical, biological, and interacting ecosystem processes. 

All of these processes respond dynamically toward equilibration of fluvial boundary 

conditions, matter transport, and energy dynamics. Consequently, we depict FREs as 

primarily driven by physical factors. This emphasizes the foundational role of physical 

processes with dependent biotic composition, structure, function, and trophic interactions 

(figs. 10, 11). Although little-discussed in the RCC, lower stream order FRE changes 

often occur in a punctuated, stepped, or reach-boundary fashion, as flowing water emerges 

from groundwater and FREs receive tributary contributions. In contrast, higher order 

streams terminating at the confluence with the sea (but not those terminating in endorheic 

basins) generally change more gradually, both spatially and over time.

Fluvial Hydrogeology Model Elements

Overview

Hydrogeological influences on FREs vary in relation to latitude, elevation, season, 

stream order, and the derivation of reach source waters (e.g., precipitation and runoff, 

groundwater, springs, or lakes). This variation exists on scales ranging from microsite to 

local stream cross sections, among reaches, across entire basins, and through different 

climate/humidity regimes (figs. 10, 11). Impacts also vary as a result of differing 

regional and global tectonic histories. Hydrogeological FRE models have developed 

incrementally over the past century, building on each other without any major conceptchanging epiphanies. Careful experimentation and analyses have gradually expanded 

the logic, empirical background, and predictive capabilities of groundwater and surface 

water models. However, additional research is needed to incorporate ecological and 

evolutionary concepts into fluvial ecohydrology.

Basin Geography

The geologic setting in which groundwater and surface water basins exist control 

FRE form and function. Tectonics, geography, sources, elevation range, basin structure, 

climate, parent rock stratigraphy, glacial rebound status, structural geology, and aspect all 

influence rivers to varying degrees (figs. 10, 11). Tectonic controls on river development 

were reviewed by Holbrook and Schumm (1999), but coarse-scale configuration of 

groundwater and surface water basins under different tectonic regimes needs further 

elucidation. Large, high order rivers typically interconnect intracontinental basins through 

periods of tectonic quiescence. Such periods of tectonic inactivity can persist across 

evolutionary and geologic time scales. Integration of large river systems across complex 

landscapes has involved 10+ million years in the Nile, Colorado, Rio Grande, and 

Mississippi River basins (Mack et al. 2006; Timmons and Karlstrom 2012; Wickert et al. 

2013; Woodward et al. 2007). 

Orogenies that dominate continental margins often produce abundant low to moderate 

stream order basins. In contrast, extensional terrains in East Africa and western North 

America generate streams of moderate order from high elevations that discharge into 

isolated, endorheic basins. While aquifers that develop in these landscapes often are 

assumed to be constrained by surface catchment boundaries (e.g., Springer et al. 2016), 
86 USDA Forest Service RMRS-GTR-411. 2020.

such is not always the case. For example, groundwater passes more than 200 km beneath 

several endorheic basins in Nevada before emerging in the lower White River (Winograd 

and Thordarson 1975).

Tributary Effects

Tributary influences on mainstream rivers vary in relation to stream order and the 

extent of differences in hydrography, water quality, sediment loading, and biota (Benda 

et al. 2004b; Dye 2010; Rice et al. 2001; Thorpe and DeLong 1994; Ward and Stanford 

1983). Habitat complexity at tributary confluences increases ecological productivity 

and biodiversity, and it sustains biogeographic habitat connectivity in four dimensions 

(Naiman et al. 1993; Rice et al. 2008; Sabo and Hagen 2012; Stevens and Ayers 2002; 

Thorpe and DeLong 1994). Tributary impacts on mainstream water quality are likely 

greatest when flows of the two are similar; however, differences in biota may follow 

the opposite pattern: aquatic macroinvertebrates in a small, springfed tributary may 

substantially differ from that in the large, adjacent mainstream. In figure 11, we depict the 

magnitude of tributary influences as dark triangles of varying size, but we also note that 

the impacts of large or influential tributaries occur across mainstream order.

Source Waters

Along with precipitation, snowmelt, and lake source contributions, springs play a 

far more important role in FRE ecology than has previously been recognized. Many of 

the world’s major rivers arise from discrete springs, springfed marshes, or groundwaterfed lakes (Junghans 2016). Rivers with spring-sourced baseflow include the Amazon, 

Colorado, Mississipi, Rhine, Volga, Murray, and many others. For example, the 

Mississippi River heads in springfed Nicollet Creek and two tributaries of Elk Lake 

in Minnesota, as well as springfed Lake Itasca. The Missouri River heads at Browers 

Spring on the Montana-Idaho border. One of its major tributaries, the Yellowstone River, 

heads in the Absaroka Mountains in northwestern Wyoming, passes through Yellowstone 

Lake (which is partially springfed by biota-rich thermal vents; Morgan et al. 2007), 

and eventually joins the Missouri River. Both the Nile and Colorado Rivers head in 

springfed helocrene marshes. Lake Tahoe in the Sierra Nevada Mountains of California 

and Nevada receives about 0.05 km3

/year from groundwater inflow (0.033 percent of its 

volume; Thedol 1997). That lake is the source of the Truckee River, which terminates in 

endorheic (and also partially springfed) Pyramid Lake in Nevada’s Great Basin Desert. 

Both subaerial and subaqueous springs contribute to river flows, although subaqueous 

springs are difficult to quantify and study (Springer and Stevens 2009). Headwater springs 

often are distinctive habitats with unique water quality that may influence mainstream 

processes, such as imprinting among larval fish (Teears 2016; but see Muller-Schwarze 

2006), and they often provide headwater baseflow for rivers in non-ice-dominated 

regions. 

FRE hydrologic models often assume that river baseflows arise from diffuse 

emergence of groundwater. Such a generalized assumption is questionable, because 

geologic structure usually brings groundwater to the surface at discrete points (springs), 

generating focused point-sources of groundwater. In 2016, the senior author found that 

the headwaters sources of the Fiume Tagliamento in northeastern Italy, the last undammed 

river in the Alps, arose from a suite of high-elevation rheocrene and hillslope springs, 

with some summertime flow augmented by snowfield runoff. This was despite reports 
USDA Forest Service RMRS-GTR-411. 2020. 87

of its source flow arising from diffuse groundwater (Tockner et al. 2003). In addition, 

rivers often support infiltration, adding to groundwater supplies. However, the assumption 

that diffuse groundwater largely provides the baseflow sources of rivers has hampered 

integration of knowledge of groundwater-surface water interactions in FRE stewardship, 

obscured understanding of the contributions of springs to surface water, and biased river 

science and policy (e.g., U.S. Army Corps of Engineers and Environmental Protection 

Agency 2015).

Reaches and Stream Order

River drainage networks are often subdivided into segments and reaches, which are 

important organizational units of FREs. We define a river segment (sensu Stevens et al. 

1997b) as one or a group of river reaches that are collectively subject to an abrupt change 

(usually) in one or more ecosystem characteristics (e.g., water temperature, geochemistry, 

suspended sediment load, or gamma diversity species assemblage). Such changes often 

are introduced by a tributary, thereby affecting downstream ecology (Bruns et al. 1984; 

Rice et al. 2008). Reaches lie within segments and are distinguished geomorphologically 

on the basis of differences in parent rock geology, shoreline erodibility, slope (gradient), 

and thalweg position. Schmidt and Graf (1990) analyzed several hundred cross sections 

to define the geomorphic reaches of the Colorado River in Grand Canyon in relation to 

bedrock geology. Thus, both segment and reach boundaries commonly occur at tributary 

confluences with the mainstream. Division of watersheds into sub-basins at various spatial 

scales (e.g., the hydrologic unit classification system used in the United States) may not 

well reflect reach geomorphology or segment configuration. Agreement on definition of 

these terms remains an important issue for further progress in FRE ecology.

FRE complexity is defined, in part, by stream order, which increases when two stream 

channels of the same magnitude meet (Horton 1945; Strahler 1957) (fig. 11). We regard 

springs as “zero order” streams, and springbrooks are typically first order channels. Stream 

order increases erratically downstream: the largest rivers can exceed a stream order of 10. 

For example, with more than 1,000 tributaries more than 1,000 km long, the Amazon River 

is considered to be a 12th order stream. FREs lacustrine reaches can occur at any order 

within a basin, through which limnological processes like thermal or chemical stratification 

can influence downstream flow and geochemistry. 

Higher stream order FRE characteristics and processes vary primarily in relation 

to the extent of geologic constraint, stream order, and reach geometry (figs. 10, 11). In 

alluvial rivers, reach characteristics are generally influenced by stream order, exhibiting 

distinctive, sinuous meanders, or braiding. Shales and other soft bedrock strata allow 

reach width to expand in unconstrained river channels, resulting in less variation in flow 

dynamics, reduced responses to aspect and microhabitat effects, and reduced species 

extinction probability. Such streams usually have increased solar energy, productivity, 

and colonization potential. In contrast, in geologically constrained reaches, the bedrock 

geology and geologic structure exert stronger influences over reach geometry. Narrow 

reaches often form in harder bedrock, where their channels sustain greater variation in flow, 

stronger responses to aspect, reduced solar energy budgets and productivity, and decreased 

colonization potential coupled with increased species extinction probability. Riparian zones 

in narrow, constrained reaches typically exist in a state of perpetual succession, a process 

suspended by high disturbance intensity (Campbell and Green 1968). 
88 USDA Forest Service RMRS-GTR-411. 2020.

Large rivers are typically high order streams, and the large quantity of water they 

transport exert unique ecological influences on their ecology, buffering changes in water 

temperature, geochemistry, and equilibration timing. Johnson et al. (1995) questioned 

whether the RCC that Vannote et al. (1980) applied to large (high order) rivers, due to 

the temporally and spatially nested hierarchical organization, the potentially increasing 

influences of physical bottom-up ecological controls, increased equilibration time, and the 

inadequacy of understanding and modeling physical process interactions. 

Water Quality

River water quality varies across lithology, latitude, elevation, humidity province, 

season, and stream order within basins, and among reaches, and springs and lakes can 

influence river waters. Water quality characteristics transition over stream order and 

are important determinants of macrophyte composition and life history cues for aquatic 

macroinvertebrates, fish, and amphibians,. In turn, they influence food web linkage (but see 

Heino et al. 2015) and riparian groundwater quality. Surface flow geochemistry generally 

dominates higher stream orders, with river water quality change occurring at tributary 

confluences (Dye 2010), and to a lesser extent in side channels and shallow, low-velocity 

shoreline habitats. Limnologically influenced water quality dominates lake-sourced rivers, 

but we know of little research on natural downriver responses to such alteration. River 

water trends toward a universal quality across stream order, creating relative similar 

geochemistry among the world’s major rivers at their mouths. However, the contributions 

and evolution of FRE water quality depends in large measure on subbasin geology and the 

relative contributions of tributaries (e.g., Giller and Malmqvist 1998; Kabede et al. 2005).

As zero order streams, springs often exhibit strikingly different temperature and 

geochemical characteristics than those in the adjacent higher order streams with which 

they are confluent (e.g., Lowe and Likens 2005). The ecological transition from headwater 

springs into first order streams is highly individualistic, often occurring at a chemically and 

thermally discrete distance from the source (Morrison et al. 2013). The quantity and quality 

of riverside or in-stream springs, as well as seasonal flow changes driven by precipitation, 

also can affect stream channel geomorphology. For example, limestone (travertine)-

depositing springs shape stream geomorphology by precipitation of calcium carbonate 

(CaCO3), forming dams and creating pools similar to those behind beaver dams (Cantonati 

et al. 2016). Such springs have received global attention as important aquatic ecosystems, 

and often add much dissolved load to rivers. They also affect downstream channel 

landforms, morphology, and interstitial pore space in mainstream bed sediments. 

Geothermal springs, such as those at Yellowstone National Park and Mammoth Hot 

Springs, often deposit travertine (Sorey 1991). CaCO3 precipitation rates in Fossil Creek 

in central Arizona exceeded 0.11 g/L of discharging flow, actively depositing travertine 

that shapes channel geomorphology (Malusa et al. 2003). Travertine deposition appears to 

be facilitated by algal growth, a deposition process that quickly oxidizes organic carbon, 

either through microbial activity or through as-yet-unrecognized chemical processes, 

enhancing the role of travertine deposition in CO2 release. Larger springs can dominate or 

strongly influence or dominate riverine water quality. For example, Silver Springs provides 

nearly 13 m3

/sec of baseflow to the Silver River in Florida (Odum 1957), and the newly 

discovered Shanay-Timpishka River is a large geothermal river arising on the floor of the 

Amazon basin (Ruzo 2016).
USDA Forest Service RMRS-GTR-411. 2020. 89

Hydrography

Flow, stage, and flow variability are the primary physical factors affecting FRE 

geomorphology and ecological processes and components (e.g., Junk et al. 1989; Tockner 

et al. 2000; Topping et al. 2003). Flows of most temperate and many tropical rivers vary 

seasonally, and factors that regulate flow include drainage area, geographic position, 

humidity province, underlying lithology and groundwater hydrology, vegetation cover, 

seasonality, and regional climate. Two useful analyses for understanding individual FREs 

are annual hygrographs (plots of the average daily flow and variation across days of 

the year) and flow duration curves (plots of the annual frequency distribution of flows). 

These plots vary in relation to climate, basin structure, and location within the basin. 

Such analyses are most robust when data represent long time series, clearly displaying 

interannual variability among wet and dry years. Failure to incorporate a sufficient 

duration of flow data and the range of variation can lead to serious mis-management of 

water. For example, the Colorado River was famously over-allocated among basin States 

in 1922, based on an insufficiently long flow time series that overemphasized wet-year 

data.

Physical disturbance strongly influences FRE geomorphology and biotic assemblages, 

but how does disturbance intensity vary across the channel and across stream order? 

Magilligan (1992) used HEC-2 flood modeling to describe variation in channel boundary 

shear stress and unit stream power on an array of stream channels across 2- to 500-year 

floods. She included analyses of channel cross section and distance downstream though 

basins with varying lithology, noting at least three-fold variation in flood power through 

the basin due to valley width, which is largely controlled by lithology. Wide valleys with 

broad, alluvial channels were subject to lower flood power, because of reduced percent of 

channel conveyance during large magnitude floods and decreased rates of depth change. 

In contrast, such events generated increased flood power in narrow valleys with 

constrained channels, a pattern influenced both by basin size and by local controls, 

such as dams. Magilligan estimated minimum shear stress and unit stream power 

for “catastrophic” floods in humid, alluvial channels of 300 W/m2. Interestingly, she 

suggested that maximum flood impacts on channel geomorphology occur at discrete 

points in the reach or drainage and that such points are likely to shift over time (headward 

movement of such points seems most likely). Her insights suggest that geomorphic 

evolution of a drainage network occurs most intensively at a highly localized scale, 

affecting reach-based disturbance dynamics and FRE biotic assemblages.

Antecedent high flows exert lasting impacts on FRE structure and ecology. Foster 

et al. (1998) highlighted the importance of antecedent imprints from large, infrequent 

disturbances on ecosystem structure and function. Parsons et al. (2005) reported that 

large, infrequent floods following volcanic eruptions, wildfires, or hurricanes increased 

heterogeneity in fluvial riparian habitat, creating multiple trajectories for vegetation 

renewal along the Sabie River in Kruger National Park in South Africa and Mozambique. 

Thus, river geomorphology may continue to respond to flood impacts from the distant 

past—impacts that often are difficult to discern.

The frequency, duration, magnitude, and timing of low flows and dewatering events 

also are critical determinants of FRE ecology. If flows are sufficiently low, the river may 

become a series of pools, with limited connectivity, or may entirely desiccate. A full 
90 USDA Forest Service RMRS-GTR-411. 2020.

dewatering event can eliminate FRE macroinvertebrate and fish species, and prolonged 

dewatering alters a host of geomorphic and ecological processes and characteristics. 

However, long-term flow data on droughts and their ecological effects are limited 

and difficult to correlate. These data challenges sometimes can be addressed through 

dendrochronology. Long-term flow modeling of dendrochronology within basins among 

a wide array of arid to humid environments around the world have provided insight into 

drought frequency and duration (e.g., Akkemik et al. 2004; Case and MacDonald 2003; 

D’Arrigo et al. 2009; Melo et al. 2012; Therrell and Bialecki 2015). Such studies have 

the advantage not only of evaluating wet years and extreme precipitation events, but also 

of drought characteristics that strongly affect water resource supplies management. The 

studies highlight the need for adaptive strategies to cope with climate change (e.g., U.S. 

Bureau of Reclamation 2012). 

The impacts of natural, regular, short-term stage fluctuations in rivers are generally 

poorly known, but they are of great consequence in rivers impounded for hydroelectric 

power production (e.g., Kennedy et al. 2016). Natural semi-daily tidal bores are a 

common phenomenon in the lowermost reaches of low-gradient rivers that reach the 

sea, and less regular stage fluctuations may extend upstream into low-gradient streams 

that open into wind-influenced lakes. Daily variation in flow stage in such settings may 

desiccate or freeze macrophytes, or macroinvertebrate habitats, and eggs. It may also 

interrupt aquatic and riparian faunal feeding and other behaviors, leading to reduced 

or fluctuating primary and secondary consumer production. Given the frequency of 

hydropower-driven flow fluctuation impacts on regulated rivers, understanding the effects 

of natural fluctuating flows is an important topic for future FRE and WCM research. 

Sedimentology

The erosion, and deposition, of bed, suspended, dissolved loads, and flotsam 

are related to watershed geology, aquifer properties, flow dynamics, and channel 

configuration. In addition, ice processes, upland wildfire, forest pest insect outbreaks, 

and overgrazing affect fluvial sedimentation and nutrient transport (e.g., Bormann and 

Likens 1979). Cumulatively, including anthropogenic materials, the world’s rivers deposit 

about 20 billion metric tonnes of solid material into the sea each year (Gray and Simões 

2008). However, rather than being solely a function of basin area, sediment deposition 

is disproportionately the result of discharge from thousands of small, relatively highgradient rivers (i.e., drainage areas < 10,000 km2

) that open directly into the ocean 

(Milliman and Syvinski 1992). Large rivers deposit proportionally less sediment due 

to subaqueous storage in deltas. Alluvial reaches often have relatively uniform bed 

materials and channel landform configuration, and often are closer to equilibrium than 

are constrained channels. Parent rock lithology, geologic structure, and less predictable 

gradients and bed loads control reach characteristics and channel geometry in constrained 

channels (e.g., Hey 1982; Schmidt and Graf 1990; Tinkler and Wohl 1998; Vogel 1981). 

Models of sediment deposition and erosion are diverse (reviewed by Merritt et al. 

2003, among others), and can provide adequate two-dimensional prediction of suspended 

sediment transport through channels with varying bed roughness, channel steepness, and 

sediment transport. Some two-dimensional models accurately predict flow and sediment 

transport for some rivers (e.g., the Glen and Grand Canyon reaches of the Colorado River) 

(Rueda 2015). However, most rivers have insufficient historical flow and hydrographic 
USDA Forest Service RMRS-GTR-411. 2020. 91

resolution to permit high-precision modeling (Alley et al. 2013). Variation in turbulence, 

sheer stress, transport capacity, and bed and suspended loads results in more highly 

variable channel landform and FRE habitat variation among bedrock-defined reaches. 

In addition, three-dimensional flow modeling is needed to relate discharge and sediment 

transport to sedimentation and channel landform development.

FRE suspended sediment transport is a power function of flow, increasing and 

varying dramatically during floods (e.g., Leopold et al. 1964; Topping et al. 2003, 2013). 

Suspended and bed load grain sizes vary from clay, silt, sand, and pea gravel to boulders 

and bedrock, typically with grain size negatively related to elevation and stream order. 

Fluvial sediment grain sizes are non-randomly erodible, with clay and cobble-to-larger 

sediments far less erodible than are silt or sand (Hjulstrøm 1939). Suspended sediments 

are actively redistributed by high flows and strongly influence FRE channel landforms 

and the availability and quality of germination and maturation microhabitats needed for 

riparian plant and faunal assemblages. Deposition of fine-grained sediments tends to 

occur on the falling limb of flood hydrographs and, in constrained reaches, at discrete 

microsites, commonly in relation to release of channel constrictions (e.g., Schmidt and 

Graf 1990; Topping et al. 2013). 

FREs also transport extensive allochthonous (upland- and upstream-derived) and 

autochthonous dissolved and fine-to-coarse organic matter, depositing large quantities 

that may affect flow, channel geomorphology, and riparian habitat development (Stevens 

1997; Tockner et al. 2003). A vast literature exists on the quantity, distribution, and 

ecological roles of woody debris (reviewed by Stevens 1997). For example, coarse woody 

debris provides refugia for aquatic macroinvertebrate larvae during upper Colorado River 

basin floods. 

Channel Geometry

Geomorphologists have identified an array of river channel types based on the extent 

of geologic constraint: geologically unconfined (alluvial) and highly erodible channels 

and mobile beds differ geomorphically and ecologically from constrained (fixed shoreline 

to canyon-bound) channels with relatively immobile (low erodibility) margins and beds 

(e.g., Rosgen 1996; Tinkler and Wohl 1999) (figs. 10, 11). Alluvial streams often have 

relatively broad, meandering, lower gradient, shallow channels with highly mobile beds, 

and sometimes with anastomose channels. In contrast, geologically constrained rivers 

often have narrow, low-sinuosity channels with higher slopes, deep channels, and largely 

immobile beds. The primary variables of interest in understanding development of fluvial 

channels include: channel slope, the stage-to-discharge relationship, the hydrograph, bed 

materials, channel roughness, and the history and frequency of catastrophic disturbances 

(e.g., ice damming, earthquakes, volcanism and lava flows, and mass wasting slope 

failures) (Leopold et al. 1964).

Many channel classification systems have been proposed (reviewed by Buffington 

and Montgomery 2013), providing a robust description of channel types and ranging from 

non-symmetrical spring channels (e.g. Griffiths 2008) and bedrock channels (Tinkler 

and Wohl 1998), to a wide array of alluvial, highly sinuous, braided, or anastomosing 

channels (e.g., Rosgen 1996). From a stream ecology perspective, rivers can be divided 

into rhithron (headwaters) and potomon (lowland) zones. 

Ward (1994) identified three low-order stream segment types in European headwater 
92 USDA Forest Service RMRS-GTR-411. 2020.

alpine habitats. Kryal (icemelt) reaches were characterized by glacially sourced flows with 

low temperatures, variable discharge, and turbidity; limited and largely allochthonous 

food sources; and limited benthic assemblages. Rhithral (channel stream) reaches 

were dominated by snowmelt hydrograph, varying temperature, and more complex 

and variable stream invertebrate assemblages. Krenal (springbrook) segments were 

steadier-flow, groundwater-dependent reaches, with high water clarity and water quality, 

dominated in his study by calcium carbonate, and supporting a richer assemblage of 

benthic macroinvertebrate species derived from multiple sources. Rhithron and potomon 

segments can be either constrained or alluvial. However, few studies have quantified or 

modeled channel geometry across stream order within basins in different tectonic settings 

or across latitude and additional comparative studies are needed to reveal broader scales 

of channel geomorphological organization and the implications on FRE ecology. 

Springs that emerge along river channels can increase the frequency of bank failure 

and redirect mainstream current direction (L.E. Stevens, unpublished observations). 

Hillslope springs along the edges of smaller perennial or many ephemeral FREs may exert 

feedback on channel geomorphology by increasing dense wetland plant cover or woody 

phreatophytic shrubs and trees (plants rooted in the water table or on the phreatic surface) 

along adjacent low-order mainstream channels. Such vegetation may be sufficiently 

extensive to reduce stream power during floods, causing local deposition and geomorphic 

alteration of channel geometry.

Fluvial Landforms

Channel landforms include pools, eddy deposits, debris and cobble islands, side 

channels, paleo- and contemporary terraces, and other features, whose development can 

be mapped but not yet reliably modeled. Predicting fluvial landform development and 

responses to flow events remains limited, due in part to divergence among turbulence 

physics equations. Post-flood flow subsidence sequentially exposes an array of draining 

shoreline habitats and generates terrace development along alluvial, sand-dominated 

channels. As flow stage declines, lateral channels change from being through-flow 

features to being open at the downstream end, and then becoming isolated pools. This 

annual flood cycle allows a progression of vegetation, invertebrates, and fish assemblages 

changes (e.g., Stella et al. 2011). In constrained, sand-dominated segments, return current 

channels are distinctive side channel features that form in recirculating zones and mouth 

upstream (Schmidt and Graf 1990). Under low flows, these features can undergo fluvial 

hydrarch succession into marshes and, with prolonged dewatering, into terrestrial habitats 

(Nilsson 1984; Stevens et al. 1995).

Riparian vegetation is often strongly zoned in response to stage (fluvial groundwater 

availability) and the return frequency and magnitude of flooding or ice scour disturbances. 

Although definition of riparian terrace structure and habitat has been attempted several 

times in various hydrological and ecological contexts over the past century (Carothers et 

al. 1979; Hupp and Osterkamp 1988, 1996; Johnson 1991; Nilsson 1983; U.S. Natural 

Resources Conservation Service 2013), there appears to be little consensus in the river 

science community about terrace lexicon and classification. For example, the phrases 

“first overbank terrace,” “geolittoral zone,” and “hydrologically active zone” all apply to 

what we refer to (below) as the “lower riparian zone.” Such lack of agreement on basic 

terminology impedes understanding of FRE landform and vegetation interactions. It also 
USDA Forest Service RMRS-GTR-411. 2020. 93

hinders interdisciplinary collaboration. Here and in the figures, we describe and define 

physical and biological terminology and concepts to help clarify, assess, and descriptively 

model the WCM as the basis for proposing this FRE model. In general, we prefer the 

simplest, most intuitive landform descriptions, and therefore we propose the following 

stage-terrace definitions: 

The aquatic zone (AQZ) is the wetted channel, downslope form the lowest stage 

elevation.

The hydrologically active zone (HAZ) is inundated and scoured one to many times/

year, and it extends upward from the baseflow water’s edge up to the bankfull stage. This 

zone, when vegetated, usually is occupied by wetland and flood-tolerant plant species.

The lower riparian zone (LRZ) is the first overbank terrace. It is usually flooded or 

scoured every 1-2 years and in the arid regions often is occupied by a combination of 

drought-tolerant deciduous species (e.g., clonal Phragmites australis) and shallow-rooted 

phreatophytes (e.g., Acer negundo, Alnus, Platanus, some Fraxinus and Quercus species, 

and various Salicaceae in the American Southwest) as well as deeply rooted, nonnative 

Tamarix and Eleagnus. In mesic regions of eastern North America, the LRZ often is 

occupied by riparian understory herbs and graminoids, as well as Acer, Platanus, Quercus, 

Populus, Salix, and other bottomland hardwoods, Larix, and other species. Patagonian 

LRZs often support Salix (Stevens, unpublished data). Australian LRZs also are occupied 

by a wide array of herb and woody species, depending on the region (Lukacs et al. 2008). 

The middle riparian zone (MRZ), sometimes referred to as the second overbank 

terrace, is usually flooded or scoured every 2-10 years and in arid regions often is 

occupied both by moderately deeply rooted phreatophytes, and opportunistic and weedy 

upland species. In mesic regions, it often is dominated by upland species.

The upper riparian zone (URZ) is only flooded or scoured by large, rare (> 10 years) 

events, and in arid regions it is generally occupied by deeply rooted phreatophytes and 

both facultative and non-facultative upland-riparian plant species. In mesic regions, it 

usually is dominated by upland species.

The upland zone (UPL) occupies non-flooded/non-scoured stages above the river 

channel, and it is dominated by non-phreatophytic plant species.

FRE Soils

Organic riparian soils develop in relation to flood and saturation frequency, as 

well as organic litter deposition and decomposition. Those processes vary by humidity 

province, stage elevation (scour disturbance intensity and frequency), and microhabitat 

type. Because pedogenesis requires considerable time, riparian soil development is 

negatively related to flood and ice scour return frequency and may increase non-linearly 

with distance and elevation above the active floodplain. Soils differentially influence plant 

functional group distribution (e.g., Andrew et al. 2014). With sufficient time, geologic and 

climate conditions, geomorphology, and vegetation, riparian soils can develop into deep, 

fertile loams and serve as a sink for carbon fixation.

Riparian substrata, soils, and microhabitats affect germination niche availability 

and quality (Grubb 1977). Day et al. (1988) reported that soil fertility (as grain size and 

nutrient status) and flood scour decreased germination and growth of fluvial wetland 

plant species. Bagstad et al. (2006) examined soil-vegetation relationships along the San 

Pedro River in southeastern Arizona, reporting that electrical conductivity, silt content, 
94 USDA Forest Service RMRS-GTR-411. 2020.

organic matter content, and biologically available phosphate all increased with patch 

age, as determined by the age of associated plants, although silt content more likely was 

a function of the geomorphic setting. Stevens (1989) experimentally demonstrated that 

germination and ecesis probabilities of common southwestern United States riparian plant 

species increased in relation to the concentration of silt in sandy shoreline sediments, due 

to reduced desiccation, susceptibility to flood scour, and increased nutrient availability. 

Jacobs et al. (2007) and Stella et al. (2011) reported that complex interactions among 

upland and river environmental variables, soils, microsite geomorphology, and large 

animal behavior all affected N and P processing in mesic and semiarid African riparian 

zones. 

Soil salinity can be a strong organizing factor in riparian and lacustrine plant 

assemblages, influencing plant assemblage composition and structure. Brotherson (1987) 

identified five zones along the shore of saline Utah Lake in central Utah. The lowest 

elevation zone was occupied by halophilic saltgrass (Distichlis spicata) and dysclimax 

weeds, followed upslope (in order of increasing elevation) by saltgrass-alkaligrass, 

saltgrass-forb, saltgrass, and spikerush assemblages. Soluble salt concentration and soil 

pH decreased downslope, while moisture and organic matter increased. Forb distribution 

was correlated with micronutrient concentration. 

Fluvial Microclimate

Fluvial climate is influenced by global- to local-scale phenomena, the latter including 

nocturnal cool air subsidence and upriver mountain valley wind patterns (e.g., Draught 

and Rubin 2006; Whiteman 1990), as well as local microclimate interactions. However, 

we know of few meso-scale studies of river influences on basin, reach, or local microclimates. Stevens (2012) and Yard et al. (2005), respectively, reported that riparian and 

in-stream interception of photosynthetically active radiation (PAR) varied temporally and 

by reach in the deep, narrow Colorado River in Grand Canyon, influencing in-canyon air 

temperature, relative humidity, aquatic and riparian production. Although not yet studied, 

variation in PAR flux also may influence slope failure frequency, cliff retreat, and canyon 

landform evolution. 

At the local microclimate scale, Stanitski-Martin (1999) demonstrated that flow and 

channel aspect influenced cross-sectional fluvial solar energy flux, creating seasonally 

discrete patterns of diel temperature and relative humidity on the Colorado River in 

Arizona. In particular, she noted a discrete belt of humid air that developed over the river 

surface. This belt of humid air is evident because it conveys sound more effectively than 

less humid air farther above the surface. The humid air belt increased in thickness and 

extended into the riparian zone, particularly during nocturnal hours, and it often shrank 

during daytime. Occasionally visible as fog, it may positively influence riparian biotic 

productivity, particularly in arid regions. Such river-based microclimate patterns also 

affect riparian trophic dynamics. For example, increased riparian humidity was positively 

associated with avian species diversity in the Murray-Darling River basin in Australia 

(Selwood et al. 2016). Microclimate gradients and soil moisture availability contribute to 

the “riparian effect” in which the lowest elevation occurrences of aridlands upland shrub 

and tree species occur in ephemeral riparian channels.

Microsite variation in PAR flux and microclimate also may be a common 

phenomenon at tributary confluences, particularly in temperate constrained or canyon-
USDA Forest Service RMRS-GTR-411. 2020. 95

bound rivers, affecting FRE aquatic and terrestrial productivity. The confluences of 

east (E) or west (W)-flowing tributaries in temperate rivers that flow north (N) or south 

(S) may not receive any direct solar radiation during the winter months, particularly in 

deep canyons at latitudes greater than 23° (Stevens 2012). In contrast, in the northern 

hemisphere, the confluences of N-flowing tributaries on mainstream rivers that flow E or 

W are likely to receive a shaft of light throughout the year, potentially greatly increasing 

PAR, increasing ambient temperature, and reducing the duration of winter freezing, all 

of which add to productivity and habitat diversity at ecologically interactive tributary 

mouths. The opposite pattern (a south-flowing tributary confluent with an east- or westflowing canyon-bound mainstream) should obtain in the southern hemisphere. Such 

solar radiation patterns clarify why tributary confluences can serve as biological hotspots 

(Thorpe and DeLong 1994).

Dynamic Geomorphological Equilibrium

Rare coarse-scale events, such as lava flows, peak flows, tectonic events, ice dam 

failures, fire, and other types of physical catastrophic events exert formative and longlasting influences on channel establishment and development. Antecedent events, 

including the failure of natural dams or large paleofloods, are long-term drivers of 

fluvial geomorphology (Foster et al. 1998; Parsons et al. 2005). FRE channels undergo 

geomorphic transitions after destabilizing flows, moving toward spatially uniform energy 

dissipation, as described in Morisawa’s (1968) dynamic equilibration model (DEM). 

The DEM is a cornerstone of our WCM, in which stream channel geomorphology and 

ecosystem development trend toward equilibrium but never reach that condition because 

destabilizing flooding and other disturbances are frequent, vary in intensity, and are 

erratic.

Variation in high flow return frequency (FRF) in rivers with erodible banks affects 

FRE channel geometry and geomorphology by creating or maintaining shoreline habitats, 

point bars, gravel and cobble islands, and terraces, all of which can support riparian 

assemblages. River landforms are maintained or readjusted through hydrographic time 

and change at rates ranging from instantaneous to the geological lifespan of the watershed 

(107 years or more), depending on FRE and the stage elevation of the landform. However, 

and in part due to the power function of suspended sediment load to flow rate (e.g., Gray 

and Simões 2008), discrete FRF intervals (particularly the bankfull 1.5-2.5 year interval) 

generate new channel bedforms and scour or reset riparian vegetation on hydrologically 

active and lower riparian terraces (Campbell and Green 1968). Channel geometry and 

landforms respond over long periods, with rare, large, antecedent floods exerting longlasting impacts on channel landform distribution.

Physical Elements and Processes: Conclusions

Much widely divergent, interdisciplinary expertise is required to understand rivers as 

ecosystems, and the incomplete integration of physical science elements with ecological 

disciplines remains an impediment to FRE ecology and stewardship. The next steps in 

refining and testing HGM concepts are likely to involve intensified study. This includes 

improved modeling of groundwater-surface water interactions and interactions among 

reaches (including lentic and frozen landscape transitional reaches) among watersheds 

in relation to tectonic history and under changing climates. In addition, more refined 
96 USDA Forest Service RMRS-GTR-411. 2020.

communications among FRE physical and ecological researchers may help clarify 

future science directions. This dialogue requires improved long-term monitoring data on 

hydrographic, hydrologic, and other physical and biological variables. Such emphases 

will enhance understanding of channel landform development, as well as the interactive 

spatial and temporal scales of hydrogeomorphological and ecological influences on FRE 

development.

Bio-Ecological FRE Concept Modeling

Historical Synthesis 

Building on the abiotic processes discussion above, we review and discuss biological 

elements and processes of FRE ecology and evolution. In addition, we use a description 

of the details of figures 11 and 12. We examine continuity and other models of FRE 

processes and characteristics within the watershed across stream order and hierarchical 

frameworks of stream habitat classification (e.g., Frissell et al. 1986). Many river 

ecosystem models have been developed over the past several decades, including recent, 

highly integrative approaches. While many of these models have substantially advanced 

understanding of FRE structure and function, none are mutually exclusive, and many 

broadly but incompletely overlap.

Hutchinson (1967) and Hynes (1970)

While G.E. Hutchinson (e.g., 1967) and his students and colleagues set the stage for 

much of modern ecology and limnology, the first comprehensive synthesis of flowing 

water ecology was presented by Hynes (1970). Hynes’ Ecology of Running Waters 

was a pioneering compendium focused largely on biological elements and ecological 

processes associated with streams and smaller rivers. Hynes (1970) described and related 

many discrete biological studies of flowing water, and his subsequent work included 

consideration of the watershed spatial scale and groundwater influences, advancing both 

basic and applied research. Hynes’ research included the role of the watershed, lakes, and 

ecological risk assessment in FRE ecology and management (e.g., Hynes 1974, 1975; 

Stanford 1998; Ward 1989). However, Hynes focused less attention on non-flowing 

watershed components, like the importance of the riparian zone or a unified FRE model. 

Wurtsbaugh et al. (2015) reported that focus on Hynsian (lotic) versus Hutchinsonian 

(lentic) freshwater limnology created differences in interpretation about the roles of 

habitat and biotic factors in fish population biology and freshwater aquatic research. 

Nonetheless, both Hutchinson’s and Hynes’ contributions initiated a rush of integrative 

thinking about FRE ecology that followed their 1970s work and continues today. 

The River Continuum Concept (RCC)

Beginning in the 1970s, the RCC was developed in the 1970s by Cummins (1976) 

and Meehan et al. (1977), and was formalized by Vannote et al. (1980), integrating 

aquatic and riparian domains of FREs. The RCC “... described the entire fluvial system as 

a continuously integrated series of physical gradients and associated biotic adjustments as 

the river flows from headwater to mouth” (Annear et al. 2004: 9-10). The RCC described 

“... a series of responses within the constituent populations resulting in a continuum of 

biotic adjustments and consistent pattern of loading, transport, utilization and storage of 
USDA Forest Service RMRS-GTR-411. 2020. 97

Figure 12—Expanded detail of linkage of the 

FRE, contrasting allochthonous (uplands 

and tributary) vs. autochthonous 

(mainstream) ecosystem energy inputs 

with aquatic vs. riparian domain 

interactions. Arrows indicate common 

energy pathways among trophic levels in 

the four FRE arenas. Not all interactions 

occur in every FRE, and other trophic 

interactions that are not depicted 

here may exist in some FREs. Inset A 

depicts spatial change in reach-based 

FRE structure and function in response 

to watershed changes. For example, 

upland fire can result in sediment, 

ash, and nutrient loading through 

tributaries, processes that may diminish 

FRE productivity and ecological role 

in the watershed. Similarly, reduction 

in precipitation or groundwater 

contribution through climate change 

or aquifer depletion may reduce 

mainstream and riparian function.

organic matter along the length of a river” (Vannote et al. 1980:130). The focus of the 

RCC was on lotic (riverine) processes supporting the aquatic ecology of invertebrates and 

fish and did not relate streams to groundwater and lentic sources or to potential hyporheic 

refugia (Palmer et al. 1992). Quickly recognized as a useful overall FRE model (Merriam 

1984), the RCC was broadly supported by studies of low-medium order streams, 

primarily in mesic regions (e.g., Minshall 1988, Minshall et al. 1983, 1985). Applicability 

of the RCC to higher stream orders has been questioned (e.g., Sedell et al. 1989).

The RCC described perennial streams (AQZ) bordered by hydroriparian (HRZ) 

to URZ ecosystems from headwaters to mouth (Johnson et al. 1984) and generally 

supporting large, important fish populations, particularly salmonids. Although 

concentrating on aquatic ecosystems, it emphasized interactions between aquatic and 

riparian domains. The RCC considers the riparian zone as a dependent contributor 

of organic matter, shade, and fish food, sourced from upstream reaches and adjacent 

uplands. Riparian zones provide important ecotonal interfaces between aquatic and 

terrestrial domains (reviewed in Committee on Riparian Zone Functioning and Strategies 
98 USDA Forest Service RMRS-GTR-411. 2020.

for Management 2002). In addition to autochthonous primary productivity, FRE 

aquatic domains receive essential contributions of nutrients and bio-available leaf litter, 

wood, detritus, and dissolved organic carbon, as well as recalcitrant carbon from the 

adjacent riparian zone. The bio-available portions of those contributions are processed 

by stream microbes and macroinvertebrates, providing food resources for fish. In turn, 

the adjacent riparian ecosystem depends on subsurface water from the hyporheic zone, 

other groundwater, and surface flow events including overbank flooding. As originally 

described, the RCC did not explicitly extend into the hyporheic zone, but a plethora of FRE 

macroinvertebrates, such as Plecoptera (stoneflies) and amphipods, can occur there (Palmer 

et al. 1992), some being found in 10 m-deep wells as far as 2 km from the channel on the 

floodplain of the Flathead River, Montana (Stanford 1998; Stanford and Ward 1988; Annear 

et al. 2004). 

The RCC quickly became the leading paradigm explaining FRE ecology, and it 

stimulated a vast wave of research, assessment, metrics development. It also encouraged 

policy relating water quality, aquatic macroinvertebrate feeding guilds, and fish populations 

as management indicators of stream health (Annear et al. 2004, 2009). This emphasis 

has included Karr’s (1991, 1999) development of the index of biotic integrity (IBI), 

which assembles multiple layers of information to provide a quantifiable metric of stream 

ecosystem health to guide management and stream rehabilitation. Reviews of the success of 

IBI and other metrics by Merritt et al. (2008), and of the success in understanding gradient 

responses of aquatic macroinvertebrates by Heino et al. (2015), are somewhat reserved 

because of the growing recognition that these metrics and relationships often only weakly 

describe patterns of macroinvertebrate biodiversity and distribution. Interactions occur 

among physical and biotic variables, not to mention anthropogenic factors, operating at 

microsite to watershed scales. These interactions influence the ecological integrity of FRE 

aquatic habitats in sometimes incomprehensible ways.

The RCC champions the concept of ecological connectivity, which was recognized in 

the 1980s as “... maintenance of longitudinal, lateral, and vertical pathways for biological, 

hydrological, and physical processes” (Annear at al. 2004:215). However, spatial focus 

on watershed and river flow and geomorphology in the RCC has somewhat neglected 

the temporal dimension, which is integral to FRE development, function, and dynamic 

seasonal and interannual geomorphic adjustment (Morisawa 1968). Although connectivity 

is recognized as important to perennial streams and adjacent riparian zones, it also is 

applicable to ephemeral and intermittent streams and their adjacent riparian zones, as well 

as river source lakes, lentic zones, and groundwater sources, including springs (National 

Research Council 2002). 

Subsequent to Hynes’ (1970) work and formulation of the RCC, many FRE syntheses 

have been undertaken, including comprehensive edited volumes and reviews by: Annear et 

al. (2004); Bouwman et al. (2013); Fisher et al. (1998); Humphries et al. (2014); Johnson 

and Jones (1977); Johnson et al. (1985); Karr (1991, 1999); Karr and Chu (1999); Malanson 

(1993); Minshall et al. (1983, 1985, 1988); Naiman et al. (1998, 2005); National Research 

Council (2002); Sedell et al. 1989; Stanford (1998); and Thorp et al. (2006, 2008). While 

a full review of these syntheses is beyond the scope of this document, we note that each 

review has emphasized particular aspects of FRE ecology, somewhat to largely overlapping 

previous reviews, but sometimes shifting focus away from conceptual integration.
USDA Forest Service RMRS-GTR-411. 2020. 99

Contributing Ecological Models and Syntheses

During and since the initial formulation of the RCC, a number of important additional 

models have been proposed. These contributions appear to us and others (e.g., Sedell 

et al.1989; Ward 1989) to refine FRE ecology as critically definitive axioms, corollary 

processes, and addenda to the RCC, although several offer alternative perspectives on 

FRE ecology.

The nutrient spiraling concept was formulated by Newbold et al. (1981; reviewed 

in Ensign and Doyle 2006) to describe the helical ecological pathways of material 

and nutrient processing through FREs. Autochthonous (endogenic) and allochthonous 

(exogenic) production is processed, released, and taken back up as water flows through 

uplands and tributaries into riparian zones and mainstream rivers. Nutrient spirals 

lengthen with stream order and with higher flows, thus, spiraling is a process that directly 

links upstream to downstream reaches, in addition to lateral shoreline habitats.

Although developed to describe the impacts of impoundment on regulated rivers, 

Ward and Stanford’s (1983, 1995) serial discontinuity concept (SDC) illuminated the 

issue in many natural rivers of the roles and impacts of natural dams that form lacustrine 

reaches and affect FRE channel geomorphology, flow, and population dynamics, both 

upstream and downstream from the dam. Lacustrine reaches can occur anywhere in a 

basin as a result of tectonism, lava dams, slope failure, or glacier development, and the 

natural dams may persist for short to long durations. Lake Victoria in Uganda formed 

as a result of tectonic rifting, interrupting the flow of the Kagera and other Nile River 

headwater streams. Lago de Nicaragua (L. Cocibolca), the largest lake in Central 

America, formed as a result of tectonic uplift in the lower Tipitapa and San Juan River 

basins. Prominent examples of North American lava flow dams include the formation 

and collapse of basalt dams; large to enormous paleo-impoundments in lower Grand 

Canyon (Crow et al. 2008; Dalyrmple and Hamblin 1998; Fenton et al. 2004); and in 

the Verde River basin in central Arizona (Elston et al. 1974). Costa and Schuster (1987) 

claimed that natural impoundments created by landslides, glacial ice, and glacial moraines 

constitute substantial threats to human life and property. They identified six types of slope 

failure dams around the world, ranging from relatively common single events that halfimpounded a valley, to rare, simultaneous impoundment of multiple valleys, creating 

several to many natural lakes. The largest slope failure dams in the Colorado River basin 

in the American Southwest occurred in the Surprise Valley region in the middle of Grand 

Canyon (Rogers and Pyles 1980). Due to several triggering mechanisms among several 

events, nearly 15 km of the north rim failed (at least once catastrophically), slid into the 

canyon, and repeatedly dammed the river, causing it to shift its deeply incised course. 

Other large dam-forming landslides are known throughout Grand Canyon and elsewhere 

in the region, including a 0.3 km3

 dam-forming landslide in the Virgin River Canyon 

downstream from Zion National Park (Castelton et al. 2016; Hereford et al. 1995). 

Enormous Pleistocene glacial dam outburst floods from Lake Missoula swept through 

the lower Columbia River drainage, creating scablands geomorphology (Benito and 

O’Connor 2003; Bretz 1923).Ice damming floods are well known on fjord rivers as well 

(e.g., Reeburgh and Neburt 1987). 

Like anthropogenic dams, natural impoundments change river seasonal water quality 

and flow, hydrography, stage relations, velocity, habitat quality and distribution, and FRE 
100 USDA Forest Service RMRS-GTR-411. 2020.

biogeography. The SDC posits that the location and size of a dam resets and influences 

downstream recovery of the FRE “normal” conditions through tributary contributions 

of flow, water quality, and biota (Dye 2010; Rice et al. 2008). Furthermore, the extent 

of river recovery from natural impoundment varies in relation to the number, size, and 

flow characteristics of downstream tributaries. Like large, influential tributaries, natural 

dams are not controlled by stream order (they can occur anywhere within a basin), and 

a large and persistent natural dam may exert long-term direct and antecedent impacts on 

FRE geomorphology and ecology. Like anthropogenic impoundments, natural dams also 

seasonally influence upstream FRE habitat quality and connectivity.

Also related to Ward and Stanford’s (1983) SDC, the discontinuity impacts of 

larger tributaries often are abrupt and generate multi-reach alterations of mainstream 

FRE ecology, as incorporated in the link discontinuity concept (LDC; Rice et al. 2001). 

The LDC posits that rivers are networks of tributary confluence nodes linked by the 

mainstream. This perspective was advanced in the network dynamics hypothesis (Benda 

et al. 2004b), which integrates drainage networks of channel and confluences, and 

proposes that the complexity of the overall basin shapes tributary contributions to the 

mainstream, resulting in minor to overwhelming influences on mainstream characteristics 

(Benda et al. 2004a,b; Clay et al. 2015).

The flood pulse concept (FPC; Junk et al. 1989; Tockner et al. 2000) added the 

importance of high flow pulses to the FRE function. Floods regularly restructure river 

landscapes, remove encroaching channel margin vegetation, transport and deposit large 

woody debris, trigger river biota life history events, and open new habitat for germination 

and establishment of fluvial and riparian biota. Floods vary in frequency, duration, timing, 

and magnitude, sometimes on catastrophic spatial scales (e.g., regular seasonal flooding), 

accounting for the state of suspended succession recognized as a common characteristic 

of natural riparian vegetation by Campbell and Green (1968).

Ward (1989) clarified four dimensions of spatial and temporal scale operating in most 

lotic ecosystems, including: (1) the “longitudinal” dimension up- and downstream through 

rivers; (2) across-channel, riparian-aquatic domain interactions; (3) vertical interactions 

with hyporheic habitats and groundwater; and (4) a broad temporal dimension (fig. 11). 

Dynamic interactions among all of these dimensions contribute to the individuality of 

character of FREs.

Ecosystem research and stewardship require detailed and long-term understanding 

of geologic, hydrographic, biota, and land use history within a basin. White (1979) 

emphasized this in relation to forest management, and Décamps et al. (1988) and Petts et al. 

(1989) described the relevance of such understanding in western European river basins. 

Following Stanford (1998) and Ward et al. (1989, 1998, 2002), Malard et al. (2002) 

expanded discussion of the biodiversity and roles of FRE hyporheic exchange through 

microbial and macroinvertebrate distribution and activity (figs. 10, 11).

Focusing specifically on river riparian zones but related to Ward’s (1989) 

considerations, Nilsson and Svedmark (2002) recognized that four major processes or 

characteristics interactively function in FREs: (1) The flow regime (hydrograph) regulates 

FRE ecological and geomorphological processes, including riparian succession. (2) The 

channel provides a corridor for inorganic and organic transport, primarily downstream 

but also upstream, and including dispersal of propagules. (3) The riparian zone functions 

as a filter and boundary between upland and riverine processes. Naiman et al. (1993) and 
USDA Forest Service RMRS-GTR-411. 2020. 101

Nilsson and Svedmark (2002) also recognized (4) that riparian zones are particularly rich in 

biodiversity and species interactions, and play important roles in watershed biodiversity. 

From advances in nearshore marine ecology patch dynamics concept (e.g., Pringle 

et al. 1988; Townsend 1989), Thorp and DeLong (1994) proposed and refined the river 

productivity model (RPM), which posits that production, as well as decomposition, 

recruitment, and other important river processes are distributed nonrandomly within the 

channel. These processes occur at specific points or in specific zones in the channel, such 

as at tributary confluences, along shorelines, or in certain depositional settings. Thus, 

their RPM proposes that river ecosystems are mosaics of microhabitats with differing 

ecological functions. They and their colleagues subsequently expanded this description 

to more fully describe the complexity of river food webs (Thorp et al. 2006, 2008), and 

much subsequent research has proceeded from this area of inquiry.

Fisher et al. (1998) proposed the telescoping ecosystem model, in which the FRE 

includes nested, concentrically positioned subsystems of the stream, hyporheic zone, 

channel margins, and riparian zone. Collectively, these function like a telescope, 

extending or retracting in relation to flood disturbance. Their model implies stepped 

transitions at tributary confluences and reach boundaries across stream order.

Following the lead of Jacobs et al. (2013) and Meybeck (1982), Bouwman et al. (2013) 

emphasized the importance of integrated biogeochemistry in FRE conceptualization. 

Rather than regarding a river simply as a single thread of flow or as a coupled mainstream 

and floodplain landscape, they considered all parts of the flow system (river mainstreams, 

floodplains, lakes, wetlands, etc.) as a biogeochemical retention and processing network. 

The river wave concept (Humphries et al. 2014) posited that most phenomena within 

at least the FRE aquatic domain can be viewed in a wave context. For example, river flow 

can be characterized as a wave that varies in frequency, length, and shape, and travels 

lengthwise and laterally through the channel. Wave position determines or regulates 

production and transport of organic material, with allochthonous input occurring primarily 

in troughs and ascending limb of the wave, and with autochthonous production occurring 

on the crests and declining limbs of the wave. 

Muehlbauer et al. (2014) defined the biological stream width as the distance of travel 

of resource subsidies from the FRE aquatic domain into the surrounding upland terrain. 

They used meta-analysis to model the spatial extent of this “stream signature,” reporting 

that the 50 percent stream signature (the point at which subsidy resources are half the 

aquatic domain maximum) lay only 1.5 m from the stream edge, but that 10 percent of 

the signature can extend more than 0.5 km into the adjacent uplands. Thus, the biological 

stream width often is much larger than that defined hydrogeomorphologically.

While expansive in scope, the RCC is based on orderly, gradual downstream 

connectivity within a perennial alluvial channel network, and it does not well describe 

ephemeral and intermittent streams-riparian ecosystems or groundwater-surface water 

interactions (Stanford 1998). Ephemeral FREs are colloquially known as dry washes, 

arroyos, wadis, and other names throughout the world, and comprise more than half of the 

global stream channel network (Datry et al. 2014). Such stream ecosystems are becoming 

increasingly abundant as rivers are dewatered by human activities and subjected to a 

drying climate. A well-documented example was the conversion of the Santa Cruz River in 

Tucson, Arizona from a perennial to an ephemeral stream through groundwater pumping 

(Webb et al. 2014). 
102 USDA Forest Service RMRS-GTR-411. 2020.

Flooding releases CO2 sequestered by burial of organic matter and invertebrates, such 

as clams (Butman et al. in review; Smith et al. 2016), and that process is interrupted in 

ephemeral streams. Benthic invertebrates that shred, graze, or collect organic debris often 

are absent or rare in ephemeral streams, reducing decomposition rates, and their roles are 

often replaced by microbes and physical molar action when the streams flood. 

Terrestrially, ephemeral versus intermittent riparian zones are bordered by distinctive 

suites of xeroriparian (dry riparian) to mesoriparian perennial plant species that provide cover 

and food resources (Johnson et al. 1984). Analysis of an ephemeral stream near Peshawar 

University in Pakistan revealed that deeply rooted woody perennial shrubs occurred in the 

wash, and weeds dominated the bed following winter rains, with drought resistant species 

occurring on terraces (Chaghtai and Khattak 1983). Aquatic productivity and trophic 

energetics of aridland ephemeral streams are reduced and interrupted during dry seasons 

(e.g., Jenkins and Boulton 2003), interrupting spatial and temporal “flow-through” RCC 

processes, but ephemeral channels commonly provide essential wildlife habitat connectivity. 

They perform as punctuated, rapidly functioning biogeochemical reactors (Larned et al. 

2010). More study of ephemeral stream ecosystems continues to be warranted.

Biological Processes and Characteristics

Overview

Climatological, hydrogeological, and geomorphic physical processes generate the 

microscale-to-watershed template on which the microclimate and biological functions 

of FREs generate or develop nutrient dynamics, assemblage composition and structure, 

and trophic interactions (Merritt et al. 2010; Naiman et al. 1993, 1998, 2005). Here we 

describe and illustrate these processes and interrelationships through the figures presented 

below, with figure 11 as the template, figure 12 adding detail to figure 11, and subsequent 

figures to illuminate key ecological processes and interactions between linked aquatic 

and terrestrial domains. These biological processes and elements sometimes reciprocally 

influence physical factors, such as sedimentology, channel geometry, and microclimate 

(Merritt 2013; Montgomery et al. 2003; Pollen et al. 2004), as well as FRE characteristics, 

such as ecological resiliency. We discuss these biological processes and elements, placing 

additional emphasis on riparian vegetation formation and dynamics.

Nutrient and Organic Matter Production and Release

FREs export allochthonous (upland and upstream) and autochthonous fine to coarse 

woody debris, other organic matter, and nutrients downstream through ecological spirals 

(Newbold et al. 1981) (figs. 11, 12). Although the dominant direction of FRE matter 

transport is downslope and downstream, important returns of nutrients also can occur 

through upstream aerial or zoochorous transport of sediments, nutrients, and propagules 

(e.g., spawning salmon, wind-blown seeds, or adult forms of aquatic insects; review 

by Cederholm et al. 1999). N, P, and C are differentially stored in alluvial reaches but 

are generally exported from constrained reaches. Fluvial nutrient dynamics models are 

needed to relate nutrient fixation, storage, retention, transport, and recycling, but such 

models remain a largely overlooked dimension in FRE ecology (Bouwman et al. 2013).
USDA Forest Service RMRS-GTR-411. 2020. 103

Biodiversity

Elevated biodiversity in FREs is universally recognized. Nearly 7.5 percent of the 

nearly 1.5 million described species on Earth are aquatic and occur in freshwater habitats, 

and many more aquatic and riparian invertebrate species remain undescribed (Collon et 

al. 2014). In addition, many other species occur in adjacent wetland and riparian habitats. 

The high biodiversity and proportion of gamma diversity in FREs can be attributed to 

a suite of factors (Naiman et al. 1993). Gravity transports propagules downslope from 

the uplands and also downstream into different reaches and segments (Johansson et al. 

1996; Stevens 2012). The dendritic network of a channel system provides a corridor for 

movement or range shifts for many species. As ecotones, FREs bring different suites of 

species together (Décamps and Tabacchi 1994; Gregory et al. 1991; Naiman and Décamps 

1997). Also, rivers provide shifting mosaics of highly productive habitat, allowing 

different colonization opportunities over space and time (Naiman et al. 2005). Stevens and 

Ayers (2002) reported that nearly half of 1,400 vascular plant species in Grand Canyon 

occurred within 200 m of the Colorado River elevation, of which one-third were riparian 

and 10 percent were springs-dependent species. In addition, more than two-thirds of the 

regional fauna obligatorily or facultatively used riparian habitat. Thus, although the area 

of FRE habitat is trivial, it supports a vast proportion of regional biodiversity, particularly 

among aridland river basins (Jansson et al. 2007; Stevens 2012).

Nilsson et al. (1989) studied riparian plant species richness (S) along rivers in 

northwestern European rivers, reporting peak S at middle stream order. Substrate 

heterogeneity and fineness were the primary factors influencing total richness, with a 

unimodal peak in S at intermediate levels of substrate fineness. Renöfält et al. (2005) 

compared riparian-to-upland plant S in the Vindel River basin in northern Sweden, 

reporting that S was related to local, river-related processes and corridor-based dispersal. 

However, unlike Nilsson et al. (1989), they reported a monotonic decrease in S from the 

headwaters to the coast, and high floristic similarity between the uplands and the riparian 

zone. Differences between these studies highlight the roles of stream order and regional 

variation in upland plant diversity.

In an aridland comparison of riparian and upland vegetation, Bloss and Brotherson 

(1979) described plant assemblage composition in a Sonoran Desert valley near New 

River, Arizona. They reported that vegetation either intergraded or abruptly shifted 

from upland slopes into the river channel, based on the erosional structure of terraces. 

Leguminous shrubs and trees occurred differentially on middle portions of the moisture 

gradient. Although plant diversity was highest on upland slopes, diversity was positively 

related to soil moisture, and floodplain species exhibited the broadest niche widths. They 

concluded that disturbance as well as moisture availability influenced the transition from 

desert to riparian plant diversity.

Differences among the above and other studies highlight the need for comparative 

analyses of FRE plant diversity, composition, and structure across latitude and among 

humidity provinces. 

Trophic Energetics and Structure

Sunlight is the primary source of terrestrial FRE energy, not only generating the 

hydrologic cycle, but also powering photosynthesis (e.g., Stevens 2012; Yard et al. 2005). 
104 USDA Forest Service RMRS-GTR-411. 2020.

The proportional contribution of autochthonous versus allochthonous production and 

nutrients varies between aquatic and riparian domains in relation to the physical setting of 

the watershed, stream order, season, and reach-based channel geometry and connectivity 

(Fisher and Likens 1973), changing over time in a successional fashion if not disturbed by 

flooding (Fisher 1983; Fisher et al. 1982) (figs. 11, 12). High levels of intrinsic production 

may occur in clear headwater springs and low- to mid-order streams. Lower elevations 

in the basin may be warmer (or sometimes cooler if subsidence occurs) and often have 

longer growing seasons, but aquatic productivity may be reduced in higher order streams 

where organic and inorganic particles reduce PAR (e.g., Yard et al. 2005). Stevens (2012) 

also emphasized the PAR-limiting influences of cliff shading in large temperate river 

canyons.

FRE trophic structure varies among aquatic, riparian, and upland domains, and 

across the chemical and ecohydrological gradients occurring within them, as described 

above and illustrated in figures 11 and 12. Aquatic food-web structure is regulated by 

temperature (Glazier 2012), flow, sediment load, seasonality, shading, and many other 

factors. For example, Mustonen et al. (2016) used 12 experimental channels to examine 

the effects of flow and suspended sediment on primary producer, macroinvertebrate, and/

or fungal production and decomposition. Flow and sediment impacts on the response 

variables were largely independent, but interaction effects were antagonistic (e.g., flow 

stimulated algal production, while sediment loading reduced production). While their 

results are largely intuitive, such studies help clarify and quantify the fundamental driving 

features of FRE ecology.

Gawne et al. (2007) tested predictions of the RCC, RPM, and FPC models on 

river metabolism through analysis of the ecological roles of microbial and macrophyte 

assemblages in three lowland tributaries of the Murray River in Australia. They 

concluded that all three models were supported to some extent, but the extent, causes, 

and consequences of aquatic primary production were varied, and no individual model 

fully explained the patterns observed. In contrast to the aquatic domain, riparian foodweb dynamics are more diverse due to the more open nature of riparian-to-upland 

interactions and feedbacks. In a seminal paper, Carothers et al. (1974) reported that the 

highest diversity of breeding (primarily neotropical) birds in central Arizona occurred 

in cottonwood-willow riparian habitats bordered by agricultural field in central Arizona, 

with riparian vegetation structure affecting not only bird assemblage composition but also 

social organization. Although not yet studied to our knowledge, the noise levels generated 

by rivers in steep canyons also is likely to influence FRE avifaunal assemblages (e.g., 

McClure et al. 2013).

Trophic cascades are regularly observed in fish-dominated ecosystems and in 

some low-order fishless systems (e.g., Blinn 2008) but are limited in FREs by physical 

processes (e.g., hydrology, sediment transport, ice impacts), where average sheer stress/

unit area appears to be negatively related to stream order (Magilligan, 1992). However, 

turbidity generally increases with stream order, reducing downstream PAR availability 

and primary through tertiary aquatic production (e.g., Yard et al. 2005). Complex trophic 

relationships also develop in riparian zones, directly and indirectly influencing primary 

producer structure and composition. For example, leaf beetles, grasshoppers, beaver, 

and ungulates all can strongly influence riparian vegetation composition, structure, and 

decomposition/soil formation (e.g., Bailey and Whitham 2006; Sacchi and Price 1988), 
USDA Forest Service RMRS-GTR-411. 2020. 105

and such impacts can be modified by both secondary consumers (e.g., ants; Schweitzer 

et al. 2005) or top predators (e.g., wolves; Beschta and Ripple 2006, but see Marshall et 

al. 2012). In one example of complex riparian interactions, manna (honeydew) produced 

by the host-specific cicadellid tamarisk leafhoppers (Opsius stactogalus) stimulated 

soil fungal growth, which in turn killed germinating seedlings of both the host plant 

and other riparian plant species beneath Tamarix canopies (Simieon and Stevens 2015). 

Whether such multi-trophic-level interactions constitute an evolved process remains to 

be determined, but proximally it ensures Tamarix stand persistence through inhibitory 

(rather than facilitation or tolerance) successional mechanisms (sensu Connell and Slayter 

1977). Just one of many complex riparian trophic interactions, this interaction is favored 

in relatively constant environmental conditions, and is most influential on the broad, lowgradient floodplains of higher order streams.

Trophic cascades also arise from interactions among other aquatic and terrestrial FRE 

taxa. For example, predatory aquatic invertebrates, such as hellgrammites (Megaloptera: 

Corydalidae), can function as top predators in clearwater streams, foraging actively at 

night when fish may be less able to detect prey. Amphibians can alter algal composition, 

algal production, and organic matter dynamics in small streams, and they can function 

as important aquatic and riparian predators (e.g., the giant aquatic salamander 

Cryptobranchus alleganiensis), thereby influencing aquatic-to-terrestrial ecosystem 

energy transfer (Whiles et al. 2006). The dual nature of amphibian life cycles may mean 

that the loss of an FRE amphibian species equates to the loss of two functional species, 

effects that may be greatest in neotropical stream ecosystems.

 Finer-scale illustration of FRE trophic structure (fig. 12) depicts the complex foodweb interactions among aquatic and terrestrial domains within reaches, which produce 

and receive ecosystem energy from autochthonous as well as from upslope, upstream 

(including tributary), and groundwater-derived allochthonous sources (e.g., Townsend 

et al. 2000). Four relatively discrete component triangles are illustrated in figure 12: 

autochthonous mainstream and riparian zone, and allochthonous tributary and upland plus 

tributary riparian triangles. These four triangles are depicted as being ecologically open 

and potentially interacting with each other (dashed lines). The most common inter-trophic 

interactions among the four components occur vertically (primary producers through 

tertiary consumers), but many other complex interactions exist among trophic levels 

among these four triangles. For example, across-triangle interactions include such foodweb interactions as crocodilians feeding on riparian or upland ungulates and predators in 

tropical rivers. 

Environmental variation within reaches and across spatial and temporal scales 

influences the relative contributions of the four FRE component triangles (fig. 12, inset 

A). For example, under a drying climate, decreased runoff may reduce or eliminate 

mainstream and riparian nutrient contributions. Conversely, increased wildfire frequency 

and severity may at least temporarily mask or inflate the influence of upland and tributary 

nutrient contributions to the watershed. Thus, relationships among component triangles 

are expected to vary over time and therefore are not expected to reach equilibrium.

FRE Biogeography

FRE biogeography involves colonization, recruitment, and population establishment 

overland by volant and other highly vagile species, as well as passive dispersal through 
106 USDA Forest Service RMRS-GTR-411. 2020.

gravity, aerial drift, or zoochorous transport of propagules through both overland 

and dendritic stream corridors (fig. 11). Regardless of the pathway, FRE population 

persistence and assemblage resilience is predicated on the ability of a species to remain 

in or disperse-recover their position in the watershed. Therefore, persistence of all 

FRE species requires some form of upstream dispersal, with eviction or extirpation the 

inevitable consequence of failed in situ or headwater recruitment strategies. 

FRE dispersal by riparian plants is achieved through hydrochory, anemochory, 

or zoochory. FRE propagules of seeds, rootstocks, eggs, or larvae drift downstream 

through hydrochory (Merritt and Wohl 2002). But propagules must be blown upstream 

or be transported by animals if that species is to persist in the FRE. Larval aquatic 

macroinvertebrates may drift downstream, while adult aquatic insects often fly or are 

blown upstream as aerial drift. Dragonflies, salmonids, and many other fish taxa migrate 

upstream to spawn, against the dominant flow direction, and some fish transport larval 

unionid mussel larvae upstream. Migratory western North American warblers and other 

passerine birds intensively use aridland riparian habitat as stop-over habitat during 

migration (Carlisle et al. 2009; Skagen et al. 2005; Stevens et al. 1977), a pattern not 

strongly evident in mesic eastern North America (Kelly and Hutto 2005). 

However, western North American songbirds generally migrate northward along 

broad fronts, rather than using FREs as navigation corridors, although “specific 

populations are likely restricted to narrower migration routes,” such as riparian nesting 

species along FRE corridors (Carlisle et al. 200; RRJ, unpublished research). Frontbased bird migrations also occur among some western North American shorebirds, but 

many waterbird species appear to follow FRE corridors, particularly through complex 

landscapes (e.g., Stevens et al. 1997a). In addition, many non-volant vertebrate species 

follow river corridors as dendritic pathways, although terrestrial faunal movements can 

be thwarted by steep cliffs, perilous crossings, and anthropogenic landscape interruptions 

(Stevens 2012). 

Colonization and extirpation frequencies vary by taxon and life history strategy and 

on the basis of physiological, life history, mobility, and reproductive strategies, which 

collectively influence species dominance in a given FRE reach. In relation to insular 

biogeographic theory (MacArthur and Wilson 1967), colonization probability in FREs is 

enhanced by productivity through habitat “hospitability” to colonizing taxa, while local 

extinction (extirpation) probability is more related to disturbance, particularly from scour 

(flooding, ice scour), low flows, and random or erratic events.

Disturbance

Ecological disturbance is defined as events that kill mature individuals within 

habitats. It exerts direct and indirect controls over sessile species richness, composition, 

and structure, as described in the dynamic equilibrium concept (DEC) (Huston 1979, 

1994; Laliberté et al. 2013; Lehman and Tilman 2000) (figs. 11 and 13). Floods, 

channel inundation or desiccation, glacial movements, ice scour, slope failure, and 

other disturbances affect sessile riparian species richness by removal or killing of adult 

organisms and resetting of environmental conditions (Connell 1978; Sousa 1984). 

Scour impacts arise from unit stream power, applied as sheer across the channel 

surface (Bendix 1992). Scour impacts vary in relation to latitude, lithology, and stream 

order in alluvial versus constrained reaches, and across stage, differentially affecting 
USDA Forest Service RMRS-GTR-411. 2020. 107

Figure 13—Plant life history 

groups (Grime 1977) in 

relation to disturbance 

intensity and productivity 

gradients (Huston 1979, 

1994). C—competitive 

species, R—ruderal species, 

S—stress tolerators. Black 

areas support no plant 

species (NS = 0), gray areas 

have very low NS, and white 

cells have some to high 

NS. Highest “hat above S” 

[middle square]=maximum 

plant species richness.

lower terraces (Magilligan 1992). Due to shear stress and channel margin erosion, scour 

mortality of vegetation may differentially predominate lower terraces, while drowning 

mortality may be more likely in the latter. Stevens and Waring (1985) reported great 

variation in flood-related mortality among guilds of riparian plant species and across 

stage elevation in the Colorado River in Grand Canyon, Arizona. However, vegetation 

assemblage responses to flooding differ among terraces, reaches, and rivers. Pettit et al. 

(2001) reported reduced germination density, plant species richness, and riparian tree size 

in the intensively flood-scoured Blackwood River in southwestern Australia; however, 

they reported neutral germination responses and little reduction in species richness in 

response to flooding along the Ord River in northwestern Australia.

The impacts of ice formation and breakup on FREs are pronounced and are gaining 

attention in temperate and boreal rivers (Prowse and Culp 2003). At higher latitudes and 

elevations, ice formation and “shoving” routinely scour shorelines and bed surfaces and 

may dam channels, uplifting and redepositing fine to coarse substrata, including boulders 

and coarse woody debris. They may also alter channel geometry. In addition, processes 

such as melting black (benthic river) ice or surface ice breakup and scour can alter benthic 

and riparian composition complexity and structure (Scrimgeour et al. 1994).

Productivity

Riparian productivity is positively associated with biodiversity through intrinsic 

mechanisms, such as organism size distribution, niche specialization, assemblage history, 

as well as interaction with disturbance gradients (Fukami and Morin 2003; Hooper et al. 
108 USDA Forest Service RMRS-GTR-411. 2020.

2005; Huston 1979, 1994; Marquard et al. 2009; Tilman et al. 2001) (figs. 11 and 13). 

Reach-based FRE productivity varies by stage (moisture and soil), aspect, and microclimate, as well as macro-scale climate, latitude, and elevation gradients. In lieu of a 

way to describe it, we refer to the colonization potential of a microhabitat as “ecological 

hospitability,” referring to the productivity and receptivity of a site to colonization. Low 

levels of productivity and high levels of disturbance reduce the richness of sessile species 

through resource limitation and reduced survival. Competition limits species richness at 

low levels of disturbance and high levels of productivity (Connell 1978; Huston 1979) 

(fig. 13).

Disturbance-Productivity Interactions

FRE riparian zones exhibit steep ecological gradients in disturbance and productivity, 

particularly in arid regions. However, riparian habitats have both the highest potential 

disturbance and productivity nearest the water’s edge. Gradients decrease with distance 

from, and elevation above, the shoreline (figs. 11 and 13). Stevens (1989) and Pollock 

et al. (1998) both reported support for the DEC (Huston 1979) for riparian plant species 

richness along western North American streams. However, Reice (1985) experimentally 

tested disturbance intensity on stream invertebrates, reporting no support for Connell’s 

(1978) intermediate disturbance hypothesis. Although high levels of disturbance limited 

diversity in his study, competitive exclusion did not appear to reduce species richness at 

low levels of disturbance. 

The intermediate disturbance hypothesis (reviewed by Wilkinson 1999), as well as the 

insular biogeography model of MacArthur and Wilson (1967), were developed for sessile 

taxa (such as plants and corals) but not for vagile species, some individuals of which can 

actively avoid disturbance events (e.g., stream abandonment behavior by the giant water 

bug, Abedus herberti) (Lytle 1999). However, Townsend (2003) found support for the 

intermediate disturbance hypothesis for benthic macroinvertebrates among 54 streams 

with varying histories of flood disturbance, reporting that bed disturbance accounted 

for the most variation in both sessile and vagile taxon richness. Nonetheless, the role of 

disturbance and the explanatory power of these FRE biodiversity models vary among 

aquatic and riparian taxa. 

Thus, while efforts have been made to distinguish the impacts of disturbance and 

productivity gradients in FREs, spatial autocorrelation prevents clear separation of the 

individual impacts of these two gradients on the structure of riparian vegetation or other 

sessile taxa (e.g., ant hives). Steady flow systems, such as zero order headwater hillslope 

springs or seeps that flow into highly disturbed channels, provide a more refined study 

context than do FREs in which to distinguish disturbance from productivity impacts. The 

impacts of other gradients (e.g., nutrient availability) on FRE biodiversity and ecological 

functions can also be more accurately assessed. 

Life History Strategies

Life history strategies among aquatic and terrestrial plant and animal species display 

complex responses to the environmental gradients and gradient interactions in FREs. A 

central focus of FRE ecology has been classification of guilds—groups of species with 

similar life history traits, particularly recruitment habitats and interactions among riparian 

plants (e.g., Johnson et al. 1984)—and feeding or habitat niche use among animals. For 
USDA Forest Service RMRS-GTR-411. 2020. 109

example, Hough-Snee et al. (2015) conducted an analysis of woody riparian plant life 

history traits in relation to environmental gradients and assemblage distributions in the 

Columbia and Missouri River basins in western North America. They identified five 

guilds of riparian plants based on rooting depth, canopy height, and resilience to flood 

disturbance, traits that may sort assemblages under a changing climate.

Grime (1977) identified three main plant life history strategies: short-lived, r-selected 

ruderal species; long-lived, K-selected competitors; and usually long-lived stress 

tolerators. Plant species characterized by these life history strategies vary spatially 

in relation to disturbance and productivity gradients, but disturbance-productivity 

interactions and competition for nutrients create conflicts within the Grime (1977) model 

(Craine 2005; Grace 1990; Huston 1979, 1994; Tilman 1988; Walker and Peet 1985). 

Plant species employ different strategies at different life history stages. For example, 

Salicaceae and Tamarix along southwestern United States streams have an exploitative, 

ruderal seedling establishment phase, but they also have more competitive and stresstolerant mature phases (Stevens 1989). Nonetheless, Grime’s (1977) three life history 

strategies may help explain some of Huston’s (1979, 1994) dynamic equilibrium 

predictions: competitors are likely to dominate low disturbance environments, while 

ruderals are likely to dominate more highly disturbed habitats, and stress tolerators are 

likely to dominate in low-productivity settings (fig. 13). As a consequence, at least part 

of the reason that Huston’s DEM predicts higher species richness at intermediate levels 

of both disturbance and productivity is that those gradient positions support all three of 

Grime’s life history strategies.

Feeding guild and habitat use also have dominated classification and applied ecology 

of aquatic FRE macroinvertebrates. The RCC focused on the downstream transition 

of secondary production in relation to stream order (although with less emphasis on 

predators and drifting terrestrial invertebrates), stimulating a rigorous national effort 

to classify macroinvertebrates as bio-indicators of stream ecosystem health (e.g., Karr 

1991; U.S. Environmental Protection Agency 2016; but see Heino et al. 2015 and Merritt 

et al. 2008). Water acceptable for human needs may support readily identifiable aquatic 

macroinvertebrate assemblages.However, U.S. water quality and quantity regulations fail 

to acknowledge that natural non-potable waters (like those commonly occurring in arid 

regions) support many common, important endemic and some endangered aquatic and 

wetland taxa and assemblages (e.g., Blinn 2008; Norment 2014).

While predatory invertebrates, amphibians, reptiles, birds, and mammals occasionally 

exert top-down trophic cascade influences on aquatic ecosystems, fish often are the most 

influential aquatic FRE species. Fish commonly affect lower trophic levels through both 

herbivory and trophic cascades. However, most conceptual modeling studies of river 

fish ecology involve applied species- and river-specific studies, usually directed toward 

flow regulation, pollution, harvest potential, and other anthropogenic impacts on socioeconomically important fisheries. Continental-scale fluvial fish ecology studies (e.g., 

Dudgeon 2000) and global modeling across continents, river basins, and stream orders 

remain relatively rare. Ibañez et al. (2009) conducted a comparative study of river fish 

feeding guild structure in Africa, Europe, and North and South America, reporting general 

support for RCC-related hypotheses that overall species richness and the proportion of 

omnivorous species increased over stream order, while the proportion of invertebrate 

feeders declined. They also reported that piscivore and herbivore/detritivore taxa were 
110 USDA Forest Service RMRS-GTR-411. 2020.

relatively depauperate in smaller North American and European streams. 

In an effort similar to that of Grime (1977), Bennett (2015) and Mims et al. (2010) 

used multivariate statistics to reveal three fish life history strategies among North 

American freshwater fish species: (1) equilibrium fish species with low fecundity and high 

juvenile survivorship (corresponding to Grime’s competitive species), (2) opportunistic 

species with early maturation and low juvenile survivorship (corresponding to Grime’s 

ruderal species), and (3) periodic species with late maturation, high fecundity, and low 

juvenile survivorship (somewhat corresponding to Grime’s stress-tolerant species). Such 

convergence of life history strategies among terrestrial primary producers and aquatic 

consumers suggests that similar biotic and abiotic constraints analogously shape FRE 

guilds within trophic levels in both aquatic and terrestrial domains. 

An ecologically intriguing group of species occupy FRE torrent habitats. Torrent 

species include organisms as diverse as: aquatic mosses; various riparian plants (e.g., 

Asteraceae: Brickellia longifolia, Poaceae: Phragmites australis, and Tamaricaceae: 

Tamarix pentandra); various insect taxa (some plecopteran stone flies; hemipteran 

belostomatids (Lytle 1999); gerrid Merobates and Trepobates small water striders; some 

dipteran tipulid crane flies and simuliid buffalo gnats; some trichopteran web-spinning 

caddisflies); immature anguillid eels and many salmonid, cyprinid, and other fish taxa; 

and some bird species (e.g., South American torrent duck [Merganetta armata], New 

Zealand blue duck [Hymenolaimus malacorhynchos], and American Dipper [Cinclus 

mexicanus]). Such taxa are specifically adapted for life in high velocity aquatic settings, 

some even occupying madicolous habitats (shallow cascading flows of white water).

Riparian Vegetation

Overview

Plant species vary enormously in flood and drought tolerance physiology, as well as 

architecture, reproductive strategies, recruitment strategies, and distribution (Malanson 

1993; Reichenbacher 1984). As with benthic invertebrates, the wide array of plant 

adaptive traits has stimulated many attempts to classify and subdivide upland and riparian 

assemblages into functional guilds (e.g., Cody 1991; Grime 1977; Hook 1984; Johnson et 

al. 1984) or mapping units. Such efforts may aid in modeling or management, but lifestage differences and the substantial residual noise in such models reveal the diversity 

of life history strategies required to cope with FRE spatio-temporal non-equilibrium 

environmental variability.

Global FRE Riparian Vegetation Distribution

The Holdridge (1947) diagram of global vegetation is based on upland gradients of 

potential evapotranspiration and annual precipitation across humidity provinces, and it 

is scaled to latitude and elevation (e.g., Lugo et al. 1999) (fig. 14). Riparian and wetland 

habitats are not considered in that model. They pose an ecologically and evolutionarily 

important contrast to upland vegetation: abundant moisture availability in riparian 

habitats greatly reduces the constraints imposed by evapotranspiration and precipitation, 

collapsing the two primary axes of the Holdridge diagram to differences among humidity 

provinces.
USDA Forest Service RMRS-GTR-411. 2020. 111

Figure 14—Simplified Holdridge 

(1947) global upland vegetation 

structure in relation to gradients 

of aridity, precipitation, 

evapotranspiration, latitude, and 

elevation.

Figure 15—Modified Holdridge 

(1947) diagram describing the 

distribution and structure of 

riparian vegetation. Moisture 

availability limitations are 

reduced or eliminated by 

proximity to groundwater 

or surface water. Potential 

riparian vegetation structure 

is depicted in relation 

to humidity province, 

temperature regime, and 

latitude. Interactions among 

elevation, latitude, and stage 

are not depicted here.

Based on our observations across elevation throughout the New World, riparian 

vegetation appears to be predominately influenced by latitude, elevation, geomorphology, 

disturbance, groundwater and surface water sources, soil moisture, Holdridge humidity 

provinces. Riparian vegetation is also influenced to some extent by water quality (e.g., 

elevated aridity and soil-water salinity impose limits to vegetation at lowest elevations). 

Temperate regions, especially those in arid and semiarid regions in the New World 

Northern Hemisphere, often are dominated by galleries of usually deciduous forests (fig. 

15). The spatial range of gallery riparian forests is remarkably broad, extending from 

20-50° and across elevation at lower latitudes from 0.02-2.5 km. Riparian shrub and 
112 USDA Forest Service RMRS-GTR-411. 2020.

woodland vegetation extend even more broadly across latitude and elevation. In humid 

to superhumid regions, adjacent upland forest vegetation often dominates riparian and 

ephemeral tributary habitats. 

In arid to superarid regions, ephemeral channel xeroriparian habitats sustain 

temporally varying fluvial groundwater moisture availability, generating reduced but often 

compositionally diverse channel vegetation. Subsurface fluvial soil moisture availability 

also often extends across the floodplain, producing aridland riparian zonation and a 

transition of vegetation—potentially from gallery riparian forest through xeroriparian 

woodlands and shrublands, to facultatively riparian upland vegetation, to true upland 

vegetation. Riparian zonation is less apparent but nonetheless evident in mesic habitats 

(e.g., Hook 1984), with riparian dominance by bottomland tree species, such as Taxodium 

distichum (bald cypress), Nyssa aquatica (water tupelo), Larix decidua (larch), Platanus 

spp. (sycamores), and several Salicaceae species.

Water Uptake and Flood Tolerance Physiology

Osmotic control on water uptake capacity differs markedly between upland versus 

phreatophytic or wetland plant species, affecting stand development, composition, habitat 

structure, and zonation, particularly in arid regions (Carothers et al. 1979; Woodbury 

1959). Upland species in arid lands often lack the ability to restrict water uptake, and 

consequently, they quickly become waterlogged and drown when inundated for even short 

periods. However, arid land phreatophyte species have the ability to regulate water uptake 

and many taxa can persist for extended periods of time when inundated (Kozlowski 1984; 

Stevens and Waring 1985; Warren and Turner 1975). While the evolutionary directionality 

of this physiological divergence apparently has yet to be explored, we hypothesize 

that selection has repeatedly and independently favored loss of osmotic control among 

xerophytic taxa as a derived trait. At any rate, the substantial physiological differences 

between these two groups of species strongly sort contemporary aridlands FRE plant 

assemblage composition across riparian terraces in relation to inundation and the depth to 

groundwater. 

Concomitant with upland-riparian variation in water uptake strategies, the diel range 

of variation in xylem water potential (Ψ) from midnight to midday varies widely among 

riparian species, and it is based on soil moisture and riparian groundwater depth. Fluvial 

wetland species exhibit a high, narrow range of Ψ, while middle and upper riparian zone 

species have progressively lower and broader ranges of Ψ. These trends were abundantly 

evident among 18 common riparian zone species along the Colorado River in Grand 

Canyon, sampled at 13 sites with known stage-discharge relationships (fig. 16). This 

variation in species Ψ ranges is largely responsible for the stage-elevation zonation 

commonly observed in southwestern aridland riparian plant assemblages. 

Zonation

In relation to the above physiological issues, aridland riparian vegetation, as well as 

lacustrine and lentic aquatic macrophytic vegetation, are characterized by zonation: bands 

of discrete plant assemblages situated co-parallel to the mainstream channel existing in 

well-defined stage zones (e.g., Bayley 1995; Brotherson 1987; Carothers et al. 1979; 

Friedman et al. 2006; Johnson 1991; Spence 1982; Stevens 1989; Woodbury et al. 1959) 

(fig. 17). For example, Friedman et al. (2006) examined the relationship between flow 
USDA Forest Service RMRS-GTR-411. 2020. 113

and riparian vegetation along 58 km of the San Miguel River in southwestern Colorado. 

They found that riparian plant assemblages were arrayed along the hydrologic gradient, 

with species like coyote willow (Salix exigua) often occupying terraces with flood return 

frequencies of less than 2.2 years, and river birch (Betula occidentalis) occupying terraces 

just upslope. They also reported that proportional cover decreased upstream, where 

upland processes such as landslides dominated the channel.

Figure 16—General conceptual model of individual riparian plant species colonization and stand replacement dynamics, with 

germination, establishment, growth, reproduction, and propagule dispersal occurring in relation to FRE hydrography. Em, Eg, 

Er

, and ratios thereof are life history energetics (E) of plant growth (g), maintenance (m), and reproduction (r), respectively.

Figure 17—Range of stage elevations of 

17 common obligate and facultative 

riparian plant species at 13 sites along 

the Colorado River in northern Arizona 

in relation to the full operating range of 

xylem water potential (MPa) for each 

species from midday (low) to night 

(high). HRZ—hydropriparian zone, 

LRZ—lower riparian zone, MRZ—

middle riparian zone, UPL—upland 

zone, URZ—upper riparian zone. 

Plant species abbreviations: Alca—

Alhagi maurorum, Arlu—Artemesia 

ludoviciana, Baem—Baccharis emoryi,

Basa—Baccharis sarothroides, Basl—

Baccharis salicifolia, Brln—Brickellia 

longifolia, Chsp—Chlorocantha 

spinosa, Coca—Conyza canadensis,

Cyda—Cynodon dactylon, Eqar—

Equisetum arvense, Eqhy—Equisetum 

ferrissii, Melil—Melilotus spp., Phau—

Phragmites australis, Prgl—Prosopis 

glandulosa, Saex—Salix exigua, Segr—

Senegalia greggii, Tara—Tamarix spp. 
114 USDA Forest Service RMRS-GTR-411. 2020.

In addition to xylem water potential, riparian zonation also arises from variation in 

phenotypic plasticity of plant species in response to geomorphology and the disturbance 

regime, soil texture, and soil water and nutrient availability, as well as root architecture 

and rooting depth. Stevens (1989) grew 12-28 seedlings of 18 common southwestern 

riparian plant species for 1 month and measured the dry biomass of above-ground and root 

growth in fine, silty, nutrient-rich pre-dam Colorado River soil (the best naturally-available 

substratum) versus coarse, sandy, nutrient-poor post-dam soils (fig. 17). We used those data 

to calculate the phenotypic plasticity index (PPIi) for each species I as the ratio of average 

dry post-dam belowground growth to average dry above-ground dry growth, as:

where mi_post_dbg and mi_post_dbg are the average dry biomass of below- or above-ground 

growth of Ni_post seedlings of species i grown in (suboptimal) post-dam fine-medium sand, 

compared to that average for seedlings of species i grown in optimal pre-dam fine silty 

sand. 

For example, seedlings of Baccharis salicifolia seepwillow, a weedy LRZ shrub with 

high phenotypic plasticity, increased its relative allocation from aboveground leaf and 

stem growth by 5.5-fold to roots when grown in nutrient-poor post-dam sand. 

We also report species-specific PPI responses to reproductive strategy, seed size, 

and seed longevity measured under field conditions. Ruderal Baccharis salicifolia and 

some nonnative species (e.g., Tamarix spp.) generally had higher PPI than did long-lived 

K-selected tree species, such as Fraxinus pennsylvanica (velvet ash), Populus fremontii 

(Fremont cottonwood), or Salix gooddingii (Goodding’s willow). K-selected Prosopis

glandulosa (honey mesquite), which can live more than 800 years in the URZ in Grand 

Canyon (R. Hereford, U.S. Geological Survey, personal communication), has a large, 

long-lived seed but also a moderately high phenotypic plasticity, conferring upon it an 

adaptive advantage in unpredictable habitats. Collectively, this analysis indicates that 

no single plant life history trait explains the success of any plant species in the riparian 

environment. It suggests that compensatory options exist within the suite of a plant 

species life history traits to permit survival in non-equilibrium riparian ecosystems. 

Riparian vegetation zonation also is common in mesic and humid-superhumid 

regions, but it is often less distinguishable than in arid regions. In humid environments, 

many upland species are tolerant of waterlogging and may have similar among-species 

variation in root architecture, plasticity, and Ψ range. For example, Dowe (2008) 

identified 263 plant species in three superhumid northern Australian watersheds, of 

which only 23 species (8.7 percent) were obligate riparian species. Oreliana et al. (2012) 

reviewed and synthesized water use models for groundwater-dependent plant species, 

suggesting that more research was needed to clarify riparian water use, and to differentiate 

water use from saturated versus unsaturated zones.
USDA Forest Service RMRS-GTR-411. 2020. 115

Riparian Plant Recruitment and Stand Replacement

The niche-box model (NBM) (Merritt et al. 2010) classified guilds of riparian plants 

in relation to similarities among life history traits. The NBM incorporates and compares 

many autecological elements for each plant species to improve prediction of vegetation 

assemblage development in relation to hydrography and riparian conditions. While 

successfully grouping some species, the large amount of variation in the niche-box model 

multivariate plots reminds us about the tremendous variation in life history strategies among 

riparian plants, variance that is highly adaptive but which does not readily lend itself to 

simple classification.

We describe and illustrate the life cycle of an individual riparian plant species to clarify 

autoecological life history, energetic constraints, and stand replacement potential (fig. 18). 

HAZ-MRZ plant recruitment often takes place on the descending arm of the hydrograph 

after the most recent flood or ice scour (e.g., Fenner et al. 1984; Rood et al. 2007). Freshly 

scoured, moist, fine-grained substrata are exposed and serve as potential regeneration niche 

sites for germination and establishment (Grubb 1977; Harper 1977). The characteristics of 

such “safe germination site” vary by species and season and are the most critical phase of 

successful recruitment. Safe sites for seed-reproducing phreatophytes with small or shortlived seeds generally require open, moist, silt-rich sediment deposits, and those species 

typically have high seedling mortality and extensive self-thinning (Type III survivorship 

curve; Deevey 1947). Common phreatophytes (e.g., wetland herbs, such as Carex, Juncus, 

Phragmites, Schoenoplectus, Typha, and some clonal Salix) often disperse through 

hydrochory as rhizomes or rootable stems, and sometimes through zoochory (e.g., beaver 

dispersal of rootable stems). Plant species with larger seeds (greater maternal investment; 

e.g., Prosopis) may germinate in shady habitats and display either Type I or II survivorship, 

vigorous growth, and less ecotypically plastic architectural responses (Stevens 1989).

Figure 18—Phenotypic 

plasticity index of 

18 common riparian 

plant species in the 

American Southwest 

(see text for calculation). 

Reproductive strategy: 

K selected, r selected, 

or rK intermediate. Seed 

size (relative): S small, L 

large, M medium. Seed 

longevity: L long-lived (> 

2 years), M medium-lived 

(0.2-2 years), S shortlived (< 1 month).
116 USDA Forest Service RMRS-GTR-411. 2020.

Following germination, establishment takes place as stems and roots grow at speciesand microsite-specific rates (Stevens 1989) (fig. 18). Energy for growth (EG) is that 

remaining after the individual expends energy on maintenance and respiration (EM and ER, 

respectively), which are relatively greater in more stressful habitats. Subsequent survival 

to recruitment age varies by species, location (i.e., stage elevation), and environmental 

conditions, including disturbance, nutrient availability, competition, herbivory, and 

other factors. Annual riparian species, living between flood spates, face a boom-orbust fate. They maximize EG, ER, and the energy devoted to propagule production (ES), 

perhaps by reducing emphasis on EM or maximizing ecotypic plasticity. The persistence 

of annual riparian species within the reach requires relatively high levels of maternal 

investment in seed production, perhaps coupled with low seed longevity (i.e., < 1 year) 

and either continuous recolonization from upstream sources or in situ propagule retention, 

recruitment dynamics that have been little studied. In contrast, biennial and perennial 

plant species can adaptively balance EG, EM and ER in relation to ES. Individuals may defer 

ES costs during periods of unusual stress (e.g., exceptional flow years), allowing those 

individuals to reallocate ES energy to EG and EM and even defer reproduction if necessary.

Riparian Succession

Vegetation succession (predictable change over time) occurs through three modes: 

facilitation, inhibition, and tolerance (Connell and Slayter 1977). Riparian successional 

trajectories may differ between Holdridge humidity provinces and in relation to stream 

order, fluvial hydrodynamics (disturbance frequency), geomorphic setting, grainsize 

distribution, depth to water table, and biological effects—conditions that collectively 

create the mosaic template on which vegetation develops (Stanford et al. 2005). However, 

variation in successional trajectories has been little studied across stream order. At one 

extreme, zero and first order streams have insufficient stream power to prevent wetland 

(if open) or woody (if forested) vegetation from colonizing stream terraces. In such lowdisturbance settings, riparian vegetation may entirely overwhelm the FRE, eliminating 

surface water (e.g., Kodrick-Brown and Brown 2007). However, at higher stream orders 

(larger, more highly disturbed riparian settings), the trunks of trees and driftwood piles 

may resist scour and affect channel geometry by stalling flow and depositing sediment, 

even to the extent of creating mid-channel islands (e.g., Tockner et al. 2003).

Geomorphic setting exerts dominant influences over riparian plant successional 

processes and modes within reaches, but its role often has been obscured in floodplain 

studies. Succession in middle- to higher-order unaltered rivers may be suspended in 

hydrologically active to middle riparian zones because annual-biennial flooding or ice 

scour resets the riparian zone to an unvegetated state (Campbell and Green 1968; Prowse 

and Culp 2003) (fig. 19). Lower riparian terraces are highly productive, but the frequency 

of scour favors species capable of sweepstakes colonization through germination, rapid 

exploitation of available space and resources, and intense competition before the next 

scouring event, and therefore favor r- (ruderal) over K- (competitive) selected species. 

Rapid growth of or colonization by woody phreatophytes may reciprocally force channel 

narrowing or alter meandering in some alluvial reaches (e.g. Johnson 1994; Hupp and 

Osterkamp 1996). However, flow events shape channel geometry in more constrained 

reaches, forcing riparian vegetation there to respond to, rather than reshaping, bar 

configuration (e.g., Birkeland 1996).
USDA Forest Service RMRS-GTR-411. 2020. 117

Figure 19—General model of 

temporal development 

or succession of aridland 

riparian plant zonation in 

relation to stage elevation 

and log10 (disturbance 

return period). Black 

lines show tendency of 

vegetation spatial and 

temporal distribution 

in mesic (humid) 

environments. HAZ—

hydrologically active 

zone, LRZ—lower riparian 

zone, MRZ—middle 

riparian zone, UPL—

upland zone, URZ—upper 

riparian zone.

Clonality in lower riparian zone species adds a further level of complexity to 

succession: Clonal Phragmites and Arundo grasses strongly resist scour and can deter 

sediment deposition during high flows, effectively armoring channels (e.g., cane breaks 

along the Lower Colorado, Rio Grande, and Mississippi Rivers). The root stocks of these 

and some woody phreatophyic clonal or rhizomatous species are highly flood tolerant and 

can persist belowground during scouring events, vigorously resprouting and growing out 

following disturbance, exploiting space and nutrients and quickly resuming dominance. 

Clonality also means that the genet (genetic individual) can move over time, as one portion 

of the root mass may be scoured away while another portion survives and regrows. Similar 

rapid colonization can occur on higher terraces after larger scouring events, but more time 

between disturbance events can allow other successional modes to occur.

Several recent studies have identified analogous successional patterns in riparian 

habitats. Naiman et al. (2005) identified four general stages of riparian succession: 

establishment, competitive stem exclusion, understory initiation, and maturity. Whether 

and how this progression occurs on all terraces remains unclear, as do the extent to which 

such patterns occur across latitude and humidity provinces. Egger et al. (2015) modeled 

terrace-based riparian succession in the Kootenai and Flathead Rivers in northwestern 

United States and southwestern Canada. They reported that individual species occupied 

similar hydraulic environments in different reaches, and that following impoundment 

(cessation of flood and ice scour), initial colonization and dominance by cottonwood 

(Populus) gave way over 60-150 year time scales to dominance by spruce (Abies). 

Although the mode was not identified, the latter transition likely occurred through 
118 USDA Forest Service RMRS-GTR-411. 2020.

facilitation as the conifers establishing under deciduous canopy shade, coupled with 

tolerance as the conifers out-lived the co-occurring cottonwoods. In those temperate 

latitudes, the two taxa apparently share similar flood tolerance and root depth-togroundwater relationships. 

Geerling et al. (2006) and Metz et al. (2016) reported three alternative successional 

trajectories on floodplains of the lower Allier River in France over 60 years of analysis. 

These alternative trajectories included: (1) progression (successional transformation 

from water to exposed sediment to colonization, first by pioneer species, to transitional 

grasslands and shrublands to riparian forests); (2) retrogression (a reversed trajectory 

toward bare soils or open water); or (3) stable (unchanging and stand-replacing patch 

conditions). They did not relate those three Markovian alternative states to the speciesbased successional modes of Connell and Slayter (1977), but such an analysis would 

likely be integrative. 

Surrounding upland assemblages strongly interact with upper riparian terrace 

vegetation. Hence, the mature (equilibrium or climax) vegetation stage in the above 

studies becomes a mixture of downslope-colonizing upland species on upper riparian 

terraces and, in arid regions, dominance by deeply rooted, long-lived phreatophytes that 

maintain root connections to the river water table (e.g., Woodbury 1959, Carothers et al. 

1979).

Riparian plant succession also can be directed by biological processes. Mycorrhizal 

succession has been identified along Montana rivers, with arbuscular mycorrhizae 

succeeding to ectomycorrhizae at decadal or longer time scales (Piotrowsi et al. 2008). 

The trajectory of riparian plant succession also can be altered by selective herbivory, 

particularly by ecosystem engineering rodents like beaver (Bailey and Whitham 2006), 

herbivorous insects (Simieon and Stevens 2015), plant diseases, and even some bird 

species (Stevens 1989). Overall, these studies indicate that reduction or elimination of 

flooding disturbance promotes as-yet-poorly-described stage-specific FRE succession. 

FRE Evolutionary Ecology

FREs have been persistent habitats over evolutionary time, as demonstrated by the 

broad array of paleo-landforms and fossil deposits throughout the world (Behrensmeyer 

et al. 1992) and the enormous biodiversity and influence of rivers on upland biota. 

Evolutionary isolation and gene flow restriction can occur at microsite, reach- and amongbasin scales, as well as between aquatic and riparian domains. Headwater elevation 

may influence along- and across-channel colonization (e.g., Vences et al. 2009). Stevens 

(2012) reported that the 32 km long, steep, canyon-bound Muav Gorge reach separated 

the Colorado River ecosystem in Grand Canyon into two basins: an isolated eastern basin 

that contains most of the river corrido’s endemic plant and animal species, and the more 

open western basin dominated by Mohave Desert species and containing fewer endemic 

species. 

Temporally, vicariance or large antecedent events, such as long-term drainage basin 

integration or natural or anthropogenic impoundment events, may contribute to isolation, 

restricting gene flow sufficiently to allow gene fixation. In addition, vegetation change in 

response to a drying climate across elevation in deep canyons may restrict across-basin 

colonization. An example is the well-known case of upland population isolation between 
USDA Forest Service RMRS-GTR-411. 2020. 119

Kaibab and tassel-eared squirrels (Sciurus aberti kaibabensis versus S. a. aberti) on the 

North and South Rims of Grand Canyon, respectively, following repeated Pleistocene 

inter-glacial habitat separation events (Jones and Wettstein 1997).

At a finer scale, microhabitat isolation is particularly pronounced in large, deep 

canyons, resulting in isolation and formation of endemic species (Stevens 2012). 

Environmentally constant and harsh, in-canyon springs, caves, north- and south-facing 

slopes, and rim edges are settings in Grand Canyon that foster development of endemism. 

Aridland springs in particular are renowned as isolated hotspots of endemism due to 

their unique water quality and environmental constancy (Kreamer et al. 2015; Stevens 

and Meretsky 2008). FRE endemism in deep canyons may develop sequentially as 

microhabitats and associated species become increasingly isolated and rare during drying 

climate phases. Climate recovery can subsequently allow re-expansion of habitats and 

associated species—and in speciation events through adaptive radiation that increases 

biodiversity, as has occurred with hydrobiid springsnails (Hershler and Liu 2008) and 

cyprinodontid pupfish (Martin and Wainwright 2013).

Local genetic adaptation is likely an important characteristic of r- and intermediate 

rK-selected riparian species. For example, following flooding, LES monitored a 10 m2

patch of Lower Colorado River riparian zone habitat that, after 1983 flood subsidence, 

supported germination of > 104

 seedlings of woody riparian species, including Tamarix, 

Salix, and Populus, but on which only a single individual Tamarix seedling survived to 

reproductive age. Such intensive selection results in highly individualistic, site-specific 

local adaptation. Subsequent reproductive mixing of those traits is likely to enhance 

species-level fitness by generating elevated genetic heterozygosity. Consequently, localscale endemism tends to be relatively rare among North American riparian plant species, 

although this may not be the case in tropical rivers due to higher productivity (e.g., Dowe 

2008; Harrison and Grace 2007). 

FRE Ecology Research Recommendations and Conclusions

Fluvial-riparian ecosystems are hierarchically and dynamically influenced by physical 

and biotic processes that vary spatially over stream order and time within the watershed, 

approaching but rarely achieving equilibrium conditions in channel geometry, fluid and 

matter transport, ecosystem energy dynamics and structure, and ecological developmental 

state. A wide array of conceptual models has been used to describe aspects of FRE 

ecology and responses to natural and anthropogenic perturbations. Most models have 

focused on single or a reduced suite of variables at site-specific, within-reach, or other 

incomplete watershed scales, and most often of anthropogenically altered streams. In the 

WCM, we emphasize the importance of understanding temporal and spatial scaling across 

the entire basin.

Despite much progress, a wide array of important ecohydrological processes, 

questions, and issues remain to be addressed or more fully investigated and integrated 

into the WCM. Not presented in prioritized order, this list of additional research topics 

includes but is not limited to: (1) the roles of self-similarity across reach and stream 

order spatial and temporal scales; (2) groundwater-surface water interactions and 

connectivity; (3) the significance, extent, and roles of groundwater and headwater 

springs as zero order streams in FRE ecology (sensu Gomi et al. 2002; Lowe and 

Likens 2005); (4) interrelationships among lentic and lotic habitats; (5) the distribution, 
120 USDA Forest Service RMRS-GTR-411. 2020.

importance and ecology of ephemeral streams and xeroriparian ecosystem ecology; 

(6) the interrelationships between interconnected ephemeral and xeroriparian riparian 

ecosystems; (7) the roles and importance of aspect, gradient, and photosynthetically active 

radiation in canyon-bound stream segments (e.g., Stevens 2012; Yard et al. 2005); (8) the 

ecological importance and roles of river microclimate; (9) the multi-dimensional roles 

of flooding, ice, and glacial effects, especially in boreal and high-elevation rivers; (10) 

multidirectional material and gene flow in dendritic pathways, which are dominated by 

downslope gravity and hydrochorous material, nutrient, and propagule transport, but that 

also have ecologically important upriver eolian and zoochorous transport mechanisms; 

(11) the role of plant physiology in riparian vegetation distribution; (12) the biogeographic 

significance of rivers as corridors, barriers/filters, and refugial systems (Stevens 2012); 

(13) stream order-driven and across-channel spatial impacts on biodiversity; (14) 

population and successional models among FRE biota and trophic levels; (15) FRE 

ecosystem genetics and evolutionary processes, including the development of endemism 

across latitude, longitude, and among tectonic landscapes; (16) climate change influences 

on FRE form and function; and (17) the role of noise on riparian songbird assemblage 

composition and structure. 

Adequately incorporating the above research topics, and more fully constructing and 

testing the WCM, will require another generation of research, including collaborative 

discussion among hydrogeological, ecological, and socio-cultural disciplines (e.g., Fisher 

1997). Such data and integration efforts are needed to improve understanding, modeling, 

and stewardship of FREs at local, regional, and global spatial and temporal scales.

River ecosystems are extraordinarily complex and vital to life on Earth. Informative 

and elegant as they are, the FRE models proposed to date remain incomplete. There 

also are many challenges associated with inadequacy of the physical and biological 

data required to calibrate and refine existing models and to develop new models. Here 

we attempted to summarize and illustrate the state of knowledge for FREs, and we 

point out additional elements that need further investigation and better integration. 

However, FREs cannot be readily, adequately, or usefully reduced to a suite of equations 

or simple illustrations. For example, some cultures commonly view rivers as living 

beings, supporting divine spirits. Integrating indigenous traditional ecological concepts 

and knowledge into improved stewardship has rarely been attempted. We suggest that 

improved comprehension of FREs may require consideration of other socio-cultural 

dimensions. 

John Wesley Powell, whose quote opens this chapter, suggested that rivers can be 

understood as music. Rather than a Hutchinsonian stage on which the ecological play 

is enacted in the theater of evolution, a river ecosystem might better be considered as a 

time-transitive orchestral composition. Such a symphony might be composed of themes 

brought in by each major tributary and integrated in its watershed and geographic 

setting. In form, the symphony might be a fugue- like integration of variably self-similar 

subthemes amplified across stream order, with rhythms reflecting reach geomorphology, 

ecological character, and tempo. Each movement in the symphony might be a time-step 

in watershed development: the opening sonata describing the basin’s geologic origination, 

the second movement portraying the FRE in its natural state, the third depicting the 

FRE under the terms of contemporary anthropogenic influences, and the final movement 
USDA Forest Service RMRS-GTR-411. 2020. 121

recapitulating the river’s ultimate stratigraphic and evolutionary contributions. Given the 

precepts of Morisawa’s (1968) dynamic equilibrium model, we do not expect this fluvial 

symphony to be either particularly orderly or melodic.

Many rivers, large and small, have come and gone on Earth. Seen in cross section, 

400-million-year-old Devonian river channels in upper Grand Canyon that intersect the 

modern Colorado River still serve as conduits for groundwater flow and perhaps direct the 

course of contemporary tributary incision (Stevens 2013). Thus, the spatial and temporal 

impacts of large paleo-FREs may last for hundreds of millions of years.

Would such a symphonic model help move river and watershed science forward? 

Each river’s symphony can help remind us of our species’ evolution and history, the limits 

of our capacity to understand multi-dimensional reality. Rivers inspire our core sense 

of wonder, our fear of, and our deep need for flowing and lentic freshwater, elements 

that have stimulated abundant fundamental, integrative, and applied FRE science. 

Composition of symphonies for major rivers would be a worthwhile artistic endeavor 

that might more broadly engage the public and the artistic community in environmental 

stewardship. But like all models, even the most elaborate river symphonies would fall 

short of fully representing these remarkable, important, and dynamic ecosystems. 

A rich array of research topics and endeavors awaits future students of FRE ecology, 

and scientific research, artistry, and public involvement all continue to be urgently 

needed. Improved stewardship of the world’s rivers to sustain vital ecological functions is 

essential to sustaining life and socio-economic well-being. This effort remains a critical 

challenge and responsibility for all of humanity. We welcome comments and suggestions 

on this synthesis, and we hope this chapter helps to stimulate the research, synthesis, and 

communications needed to reach this goal. 

Acknowledgments

We benefited greatly from discussions about this chapter with Robert Briedenthal 

and many fine river scientists we each have known over the years. We thank Courtney 

McDaniel, Tierney Schipper, and particularly Kenneth Kingsley for reviewing and editing 

the final manuscript; however, oversights and shortcomings of this manuscript remain the 

responsibility of the authors.

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2005. Influence of topographic complexity on solar insolation estimates for the Colorado River, 

Grand Canyon, AZ. Ecological Modeling. 183: 157-172.
138 USDA Forest Service RMRS-GTR-411. 2020.

Chapter 6. It’s Not All Bad News—Riparian Areas in the 

Anthropocene 

Kenneth J. Kingsley

Introduction

An abundance of information, some of which is included in other chapters in these 

volumes, documents and laments the historic loss of or changes to riparian ecosystems. 

Historic impacts to most riparian organisms, ranging in size from cottonwood trees to 

microbes, are fairly well documented. Causes of these changes are also well documented 

or speculated upon based on the best available information. It is clear that the world, and 

specifically the riparian world, has changed because of human activities. It appears likely 

that changes will continue to occur as human activities continue, and that our impacts are 

sufficient to cause the planet to enter a new era of geology, atmospheric chemistry, and 

biology termed the Anthropocene (Crutzen and Stoermer 2008). Most of the changes that 

have occurred historically are generally considered “bad” because they involve a loss of 

biological communities that we usually consider “good.” A small industry has developed 

involved in the protection of remaining riparian areas and wetlands and restoration or 

creation of new ones to compensate for historic losses and solve other problems. 

This industry has met with mixed results, but complete restoration or creation of 

fully functional self-maintaining wetlands has probably not been achieved and may 

not be entirely desirable (Karpiscak and others 2004; Willott 2004). This industry is 

largely supported by funding by the U.S. Army Corps of Engineers and Environmental 

Protection Agency for restoration or wastewater treatment projects, although some may 

be supported by private, local, or regional sources. Much effort is currently being directed 

toward intentionally “restoring” riparian communities, or at least the most conspicuous 

desirable indicators of those communities, and protecting those remaining riparian 

communities that appear to be somewhat intact. These efforts continue to lead to a greater 

understanding of the nature of riparian communities. Many components from microbial to 

landscape and planetary are necessary to make healthy ecosystems. 

Historically throughout the southwestern United States, rainwater sank into 

permeable ground or ran off impermeable or saturated surfaces into natural channels. 

These confluenced with increasingly larger streams until eventually the water sank into 

the streambed or ran to a sea or natural lake. Riparian communities developed along 

these natural drainage ways. The natural drainage scheme has been diverted, obstructed, 

reoriented, eliminated, or otherwise altered by human activities for thousands of years. In 

the most recent period of the Anthropocene, the permeability of the watershed substrate 

has been changed by overgrazing of watersheds, fires, invasion of alien plant species, 

trampling by livestock and people, and paving. Drainage ways have been channeled, 

blocked, driven underground, piped, dammed, and diverted. Especially in urban areas, 

what comes down from the sky no longer follows natural paths but travels through a maze 

of structures designed (or not) to put it somewhere else. 

One could lament the loss of the natural riparian communities that were once part 

of the natural hydrologic pathways. One could also learn to appreciate some of the 
USDA Forest Service RMRS-GTR-411. 2020. 139

ways people and nature have combined to create new and different (and, frankly, often 

inferior) riparian communities. Here, my purpose is to look at some examples of riparian 

communities that have been developed, intentionally or not, in association with human 

attempts to deal with the hydrologic cycle. Most of what follows is based on personal 

observations over several decades, and may not be documented or supported by scientific 

or other literature. When appropriate references are available, they are given.

Wastewater Treatment Plants—A Source of Water

Example 1: Santa Cruz River Downstream From Nogales and Tucson

The Santa Cruz River was historically an interrupted perennial stream (Stromberg and 

others 2012; Webb et. al. 2014) that began in the Canelo Hills of southern Arizona, passed 

briefly into Mexico, and then made a northward turn to run some 200 miles to join the Gila 

River. Today, the same generalization can be made, but historically wet reaches have become 

dry, except during brief spates, and historically dry reaches are now perennially wet. Detailed 

descriptions of the historic river and anthropogenic changes to it are given in many sources, 

most notably the recent volume by Webb and others (2014) that includes a compilation of 

historic records, including photographs and descriptions by early settlers and naturalists. 

More recent, frequently updated descriptions of conditions can be found on the website of 

the Sonoran Institute (https://sonoraninstitute.org). Groundwater pumping and watershed 

changes are primary reasons for the drying of the historically wet reaches and treated urban 

sewage effluent is the sole reason for some reaches to be perennially wet today. 

Over the past two decades, a growing body of literature examining the currently wet 

reaches has developed and is expected to expand as additional studies are conducted. 

Much of this literature is in the category of “gray literature,” consisting of agency and 

consulting reports, synopses, or abstracts of meeting presentations, theses, dissertations, 

and web-based information. A review of this literature and description of conditions in 

2013 was done by Pima County (2013). Updates of many current research projects and 

conditions are available from the Sonoran Institute website (https://sonoraninstitute.org) or 

by following leads on that website. It is clear that a new and dynamic river has emerged, 

differing greatly from the former in flow regime, water quality, substrate qualities, bank 

morphology, wildlife and plant uses and distribution, and human uses. Any description of 

“current” conditions will be of transitory value, and the situation will change, sometimes 

quite rapidly. This brief section describes conditions as they were primarily during the first 

decade of the 21st century. Several restoration projects of varying scales and approaches are 

in progress at this time (2018) and more are likely to be developed. 

In 2018, there were two perennially wet reaches, one carrying effluent from the Nogales 

International Treatment Plant in Nogales, Arizona (Upper Santa Cruz), the other carrying 

effluent from treatment plants serving the Tucson area (Lower Santa Cruz). 

Effluent input began on the Upper Santa Cruz in 1951 and on the Lower Santa Cruz 

during the 1970s, when the urban populations were less than half of present, and when 

treated effluent discharge into rivers was considered an expedient way to get rid of a 

problem. There were few regulations and little attention directed toward the discharge. 

Water flow in these reaches is subject to daily and seasonal fluctuations, depending on water 

use in the areas served by the treatment plants. Longer-term changes occur as the urban 
140 USDA Forest Service RMRS-GTR-411. 2020.

populations grow, demand for effluent for other uses varies, and sewage treatment processes 

develop. The amount of effluent in the Upper Santa Cruz was greatly reduced in 2012, 

changed in quality in 2013, and is expected to be practically eliminated as new technology 

is applied in Mexico. Added to input are occasional floods following precipitation events in 

which urban stormwater runoff and drainage from undeveloped portions of the watershed 

are added to the effluent flow. 

Water is removed from these reaches by penetration into the substrate, diversion for 

irrigation, and evapotranspiration by vegetation. The depth to groundwater has been lowered 

so much by groundwater withdrawal that there is little or no immediate connection between 

the deep aquifer and surface water, and the zone of water availability to plants is limited to 

the immediate vicinity of flow channels. Water availability to plants and subsurface recharge 

may be reduced by buildup of a clogging coating of relatively impermeable material 

(schmutzdecke) caused by physical, chemical, and biological processes. This material may 

be removed at unpredictable intervals by flooding caused by summer monsoon storms. 

If there is no flooding, then riparian vegetation may die. This is an “inherently unstable” 

situation, which may result in rapid loss of riparian communities if no human intervention 

occurs to break up the clogging layer (Treese and others 2009). Innovations at the Nogales 

treatment plant in 2009, and subsequently, have led to an improvement in water quality, 

reduction of schmutzdecke, and return of native Gila topminnows (Poeciliopsis occidentalis) 

to a portion of the currently wet Upper Santa Cruz (U.S. National Park Service 2015).

Water quality is different from that of natural water sources and varies considerably 

with distance from the inlets, natural input events, changes in human population and water 

use, and sewage processing technology. A change in water quality for the Lower Santa Cruz 

followed the 2014-15 activation of new technology that is part of the Regional Optimization 

Master Plan. This resulted in a decrease of schmutzdecke and an increase of diversity of 

aquatic macroinvertebrates and fish including Gila topminnows (Sonoran Institute 2017). 

Prior to this change, water quality was such that many species of aquatic animals and plants 

that formerly lived in wet reaches of the river could not survive, whereas others, including 

nonnative invasives, thrived. 

Propagules of native species that formerly inhabited the area may not be available, 

either because the currently wet areas never had them, or because populations have been 

extirpated from the watershed. Temperature is higher at effluent inlets than downstream. 

Nitrogen compounds are high near effluent inlets and attenuate downstream. Ammonia 

levels may be toxic to fish, amphibians, and invertebrates for a portion of the stream. Benthic 

macroinvertebrate communities vary with distance from inlets and concomitant changes in 

several parameters of water quality (Boyle and Fraleigh 2003), with taxa that are tolerant 

of pollutants present near the inlets and diminishing downstream, and intolerant taxa absent 

near inlets and becoming established downstream. Benthic macroinvertebrate communities 

are greatly limited by substrate condition (mostly sand and schmutzdecke, little gravel, few 

cobbles, very few boulders). They are also limited by flow amounts and distributions, which 

vary through a wide range hourly, daily, and seasonally. 

Microbial communities, including potentially pathogenic microbes, may be greatly 

influenced by introduction of propagules and living organisms in effluent and by variation 

in conditions downstream from inlets (Duran and Spencer 2004). Unique mixtures of 

chemicals, including biologically active compounds such as pharmaceuticals and endocrine 

disruptors, occur and may impact the plants and animals in and along the stream. 

Riparian vegetation, including cottonwood-willow forests, have developed in patches 
USDA Forest Service RMRS-GTR-411. 2020. 141

that formerly had only xeroriparian scrub, and a few areas (near Tubac and Tumacácori 

National Historical Park) of previously developed riparian forest and woodland have been 

nourished and enhanced by the effluent. Through much of the currently wet reach of the 

Lower Santa Cruz, the banks have been deeply incised and are now lined with soil cement to 

prevent bank erosion and flooding. Although the newly developed riparian areas have many 

characteristics of historic pristine conditions, they also differ in that they have a mix of native 

and nonnative plants and animals and lack many species that were historically found in the 

wet reaches of the river. 

For the most part, riparian trees are smaller in stature under the new conditions than 

those found in more natural riparian areas, but that may be because they are younger. 

Native trees include Fremont’s cottonwood (Populus fremontii), Goodding’s willow (Salix 

gooddingii), velvet mesquite (Prosopis velutina), and blue palo verde (Parkinsonia florida). 

Nonnative trees include “Chilean” mesquite (Prosopis sp.) and possible hybrids with natives, 

Jerusalem thorn (Parkinsonia aculeata, apparently nonnative in this situation), athel tamarisk 

(Tamarix aphylla), and saltcedar (Tamarix ramosissima). The mix of species in any given 

reach may be unique and may change over time. 

Because the depth to a zone of permanent available water is greater than the depth to 

which roots can reach, if the flow is reduced or changes channels slightly, riparian vegetation 

may be cut off from water and die. Changes in effluent amounts released by treatment plants 

may have profound impacts on vegetation. Minor channel changes caused by floodwaters 

rearranging the substrate can result in loss of blocks of riparian vegetation. Floodwaters 

cannot spread across a broad floodplain but are confined within soil cement walls through 

much of the urban area. This confinement increases the velocity of floods and damage to 

river bottomland vegetation when floods occur. 

Many species of native birds have discovered and adapted to the newly established 

riparian areas. Some of the areas have become quite popular with local birders and appear on 

bird guidebooks and websites. A list of species found has been growing over many years, and 

migrating species that are outside their normal range are frequently spotted. In a yearlong 

study of birds using five different sites along the effluent-dominated Lower Santa Cruz, 

133 species were recorded (SWCA 2000). There were differences between sites, between 

survey dates on each site, and between seasons. Most of the differences between sites can 

be attributed to the relative composition of vegetation associations at each site. Where 

trees, especially Fremont’s cottonwood and Goodding’s willow, were dominant, native bird 

species typically found in native riparian habitats were present. However, because blocks 

of trees were narrow and most of the trees young, many species of birds typically found in 

large blocks of native riparian habitat were missing (e.g. yellow-billed cuckoo Coccyzus 

americanus) or rare (e.g. summer tanager Piranga rubra). Twenty-seven species of wading 

birds, shorebirds, and waterfowl were documented. Many migrating birds were observed, 

typically only once or twice during the year. 

Other wildlife is known to use the area, including nonnative American bullfrog 

(Lithobates catesbianus), but no native frogs. Native Sonoran mud turtle (Kinosternon 

sonoriense) and nonnative spiny softshell turtle (Apalone spinifera) and most native reptiles 

appropriate for riparian habitats are present. Some mammals, both native (coyote, woodrat, 

cotton rat, bats) and nonnative (domestic and feral dogs and cats and cattle) were seen. 

Native fish (Gila topminnow) were found for the first time in decades in 2017 (Sonoran 

Institute 2017), but several nonnatives have been long been present, including mosquitofish 

(Gambusia affinis). Presence of bullfrogs and mosquitofish probably precludes or at least 

complicates potential restoration of populations of natives. 
142 USDA Forest Service RMRS-GTR-411. 2020.

Figure 21—Riparian habitat and multi-use trail, along the historically wet, currently dry reach of the Lower Santa Cruz River in 

Tucson, which is now part of a linear park with multi-use trail. Photo by Amy Gaiennie.

Figure 20—Riparian habitat: effluent-dominated reach of Santa Cruz River in Tucson, historically dry, currently wet. Note narrow 

width of riparian vegetation and soil cement bank. Photo by author.
USDA Forest Service RMRS-GTR-411. 2020. 143

The existing conditions of the Lower Santa Cruz River provide a valuable resource for 

the human community. A paved multi-use trail runs through developed parkland along both 

sides of the river for much of its course through urban Tucson, and it is very heavily used by 

pedestrians and bicyclists (figs. 20 and 21). The larger trees, especially athel tamarisk and 

Fremont’s cottonwood, are frequently used as campsites by homeless people. As mentioned, 

bird watchers find the area of interest. Long-term residents understand and appreciate the 

changes that have occurred, but many young people and visitors may think the current 

conditions are normal and natural and appreciate their protection as parkland. 

Example 2: An Artificial, Intentional Riparian Site—Sweetwater Wetlands

An active domain of wetland creation is constructed wetlands for processing of sewage 

wastewater (Kadlac and Wallace 2009). An example is Sweetwater Wetlands, Tucson, 

Arizona, which was created to process water that had been used to backwash filters at 

the Roger Road Wastewater Treatment Plant and put that water into the reclaimed water 

system. Considerable planning went into the effort to create an aesthetically pleasing 

wetland that would allow plants to process the water through a process of filtration, settling, 

and biological activity. Early on in the planning process, an expert on mosquito control was 

hired to help design a wetland that would not be plagued by mosquitoes. Unfortunately, 

mosquito control recommendations were scrapped in the later stages of planning because 

effective mosquito control was not consistent with effective water processing, the hoped-for 

wildlife and aesthetic values, and the project budget. 

Sweetwater Wetlands was created, planted with native wetland plants including 

Fremont cottonwood and cattails (Typha latifolia) and put into operation. It included ponds 

with islands for nesting birds (fig. 22), dense cattail stands for various stages of water 

Figure 22—One of the ponds at Sweetwater Wetlands, February 2014. Originally used to process filter backflush water as part 

of an older sewage treatment system, this and adjacent ponds are currently filled with reclaimed water to provide wildlife 

habitat and a recreational amenity for the community. Photo by author. 
144 USDA Forest Service RMRS-GTR-411. 2020.

processing, a complex of several ponds surrounded by a walkway and piers for wildlife 

viewing, interpretive signs, restrooms, a public parking lot, and a system of aquifer recharge 

ponds. Educational programs were developed and thousands of schoolchildren have learned 

about wastewater processing, wetlands, and ecology on field trips. Informative brochures 

and an “Activity Book and Field Guide” have been produced. Processed water from the 

Treatment Plant flowed out into the riverbed close to this location, and the perennial 

lower Santa Cruz River began here. With some adjustments, the desired water processing, 

wildlife, and aesthetic values were achieved, and the area has become an outstanding 

popular recreation area for birders and a valuable educational resource. 

Over 300 species of birds have been documented at the site (City of Tucson 2016). 

However, a mosquito problem also developed and has required continuous expensive effort 

(Karpiscak and others 2004; Willott 2004). Until recently, water quality could not support 

fish and may have been a population sink for migrating aquatic insects. Water quality 

improvements since 2014 have led to establishment of a dense population of mosquitofish 

and many species of aquatic insects that have probably greatly reduced the mosquito 

population. The sewage processing method changed in 2014, and Sweetwater Wetlands was 

no longer needed for its original function. Today, if the wetlands were allowed to dry and be 

removed, there would probably be a significant public outcry over the loss of recreational 

and educational benefits. Maintaining the wetlands as a park has become separate from 

the reason for their original creation because water treatment technology has changed. 

Continuous active maintenance and management are necessary to prune trees, keep 

pathways and structures in good condition, thin cattails, and prevent mosquitoes. 

Urban Runoff—A Source of Water

In the urban environment, stormwater is often directed through a series of drains, 

sewers, canals, and culverts to run off eventually to a designated depository where it 

can enter its natural pathway or be used for aquifer recharge or other uses. Some of the 

structures created to handle urban stormwater can become important wildlife habitat 

if they are allowed or created to become so. Retention basins are designed to manage 

stormwater runoff to prevent flooding and downstream erosion and to improve water 

quality downstream. Retention basins are, at best, artificial lakes with vegetation around 

the perimeter, designed to include a permanent pool of water. Detention basins are 

designed to store water temporarily after a storm, but eventually to empty at a controlled 

rate to a downstream water body. Infiltration or recharge basins are designed to direct 

stormwater (or other water) to groundwater through permeable soils. 

Vegetation may be intentionally planted or volunteer around the margins of the ponds 

or appear as emergent or aquatic plant communities. These utilitarian stormwater drainage 

features can be managed to improve the neighborhood and may be seen as valuable 

amenities, or they may be mismanaged to become major nuisances as mosquito breeding 

sites or other pest habitats. 

The U.S. Fish and Wildlife Service, in cooperation with the Arizona Game and Fish 

Department and local government agencies, has developed a Safe Harbor Agreement 

for two species of endangered fish, Gila topminnow and desert pupfish (Cyprinodon 

macularius) that permits placing these fish in retention basins and other artificial bodies 

of water (Duncan and Voeltz 2005). As of early 2016, this agreement has not been used 
USDA Forest Service RMRS-GTR-411. 2020. 145

for placing fish in urban retention and detention basins because of the concern with fish 

moving downstream to another property (D. Duncan, U.S. Fish and Wildlife Service, 

personal communication). Fish have been utilized to control mosquitoes and other 

nuisance insects in artificial ponds and have been successfully established under this 

agreement in a few other artificial habitats. Also, their use in compliance with the terms 

of the Safe Harbor Agreement has expanded the range of remaining populations of these 

endangered fish and greatly increased the number of individuals of each species, which 

benefits the species especially if anything impacts the few remaining natural habitats.

Example: Kino Environmental Restoration Project (KERP)

An example of a detention basin developed for maximizing habitat values and 

minimizing nuisances is the Kino Environmental Restoration Project in Tucson, Arizona 

(Pima County 2015). Urban stormwater runoff from approximately 18 square miles 

around the Davis Monthan Air Force Base and the City of Tucson runs through a series 

of drainages, mostly concrete-lined box culverts or ditches, to collect in the Tucson (Ajo 

Road) Detention Basin. Prior to the inception of this project, the basin of approximately 

50 acres was surrounded by concrete with a mostly mud substrate. Often the basin was 

dry, or filled with stagnant water and debris, and provided excellent mosquito larval 

habitat and dense weeds. In the mid-1990s, the U.S. Army Corps of Engineers, Pima 

County, and the Pima County Flood Control District collaborated in a cooperative 

agreement to create the Ed Pastor Kino Environmental Restoration Project (KERP). 

The goal was to create a facility that would meet three primary purposes: create native 

ecosystems, harvest urban stormwater, and control flooding. Elimination of mosquitoes 

was also an important consideration, as was elimination of trash dumping. 

The project today is managed by the Kino Sports Complex/Stadium District in 

cooperation with Pima County departments, such as the Health Department, Regional Flood 

Control District, Department of Transportation, and the Regional Wastewater Reclamation 

Department, in addition to a number of State agencies to ensure KERP meets the State and 

Federal guidelines set forth for an environmental habitat. The final footprint of the KERP 

covers 141 acres and contains 28 acres of riparian and open water including: a 5.6-acre, 

50-foot deep pond; 21 acres of grassland, mesquite bosque, marsh and upland vegetation; 

three short stream courses; and 92 acres that include flood control structures, a basin earthen 

berm, and a 2.2-mile recreational path that surrounds the basin (figs. 23 and 24). 

Native plants associated with wetlands and riparian areas, as well as upland habitats, 

have been planted and are irrigated by a combination of floodwaters gathered by the 

basins and reclaimed sewage effluent. Some invasive plant species have come into the 

area but are managed. A pumping and valve system circulates and mixes reclaimed and 

stormwater within the basin. Stormwater coming into the project first passes a debris 

basin equipped with a trash rack to retain sediment and trap floating objects. It then passes 

into lined storage ponds that retain the runoff. Two small weirs and a weir gate regulate 

the flow of water into the deep pond. Water is circulated along the three stream courses 

that drain into the ponds, and cooler water is pumped from the bottom of the 50-foot pond 

to the top of the stream courses to improve water quality and aid with vector control. 

Water levels fluctuate depending on the availability of stormwater, and much of the 

area is designed to hold large volumes of rapidly arriving water during wet periods and 

survive as upland in dry periods. During periods of extended dryness, treated effluent (aka 
146 USDA Forest Service RMRS-GTR-411. 2020.

Figure 23—Riparian habitat: Kino Environmental Restoration Project, deep pond and vegetation, February 2014. The pond was 

created primarily a receptacle for stormwater runoff from the upstream urban area, and is managed, in part, to provide 

wildlife habitat and as a recreational amenity. Its design and management allow for rapid filling following storm events and 

minimization of mosquito habitat. Photo by author.

Figure 24—Riparian habitat: Kino Environmental Restoration Area, shallow pond and vegetation, February 2014. This pond is off 

the stormwater channel, so is not subject to rapid inundation, but is filled by water intentionally diverted from the large pond 

for the purposes of aquifer recharge, water storage, and wildlife habitat. Photo by Amy Gaiennie.
USDA Forest Service RMRS-GTR-411. 2020. 147

reclaimed water) is purchased from Tucson Water to maintain pond levels and protect 

habitats. A 16-inch reclaimed water pipeline feeds water into the pond. In addition, the 

open water areas and marsh can be fed by three additional pipelines at KERP.

Harvested stormwater, when available, can be a low-cost alternative to purchasing 

and using groundwater and reclaimed water. This project has yielded approximately 19 

million gallons per year of usable irrigation water (10-year average, range 0 to 39 million 

gallons). The water has been used to irrigate the basin’s vegetation, the adjacent Kino 

Sports Complex ballpark and athletic fields, road median landscaping and easements, and 

landscaping for nearby public buildings. The project has become a popular recreational 

site and birding area, although access to the most interesting habitats is limited, and most 

viewing must be done from a distance. A casual visitor may see an assortment of birds 

representative of the variety of ecological communities present in the project, including 

migratory waterfowl, shorebirds, and wading birds. Some informative signs have been 

posted, and most have survived vandalism. Discussions are ongoing between agencies 

regarding placing native fish under the Safe Harbor Agreement. 

Areas With Water Supplied by Agricultural Irrigation

Agriculture has historically been the largest destroyer of riparian habitat, directly by 

conversion of riparian forests and woodlands to croplands, and indirectly by damming 

streams and rivers to create reservoirs, reduce flooding, and manage water availability. 

In arid lands, riparian areas are clearly the most suitable for agriculture because the land 

is relatively level, close to water, and has suitable soils. In Arizona and much of the arid 

Southwest, the primary crops have been cotton, alfalfa, vegetables, or grass. These have 

replaced native woodlands with monoculture shrublands, fields of forbs, or grassland. 

Tree crops, primarily citrus and pecans, have created nonnative woodlands. Whereas 

a great deal of research has gone into studying how to grow these crops, very little has 

looked at the ecological effects of the conversion of native riparian areas to croplands and 

the novel communities of wildlife and native plants that develop. 

Ohmart and others (1985) found that agriculture created new habitats for wading, 

water, and shorebirds, and that the type of agricultural development and crops grown were 

important in determining species composition and population sizes. Conine and others 

(1978) found that some riparian bird species used agricultural land as foraging habitat, 

whereas others did not. Wells and others (1979) examined avian use of citrus orchards 

and found that citrus was an important habitat for doves but generally lacked flycatchers, 

small insectivores, and woodpeckers, which were abundant in native riparian habitat. 

Obviously, birds that require trees do not do well in cotton fields and many other 

animals once common in native riparian communities cannot survive the conversion 

to agriculture. Other creatures, however, do manage to adapt in various ways, perhaps 

losing nest or den sites but gaining feeding areas. If nearby sites can provide the missing 

components (e.g., nest sites near grain fields, ditch banks, and culverts near alfalfa fields), 

some animals may survive and even thrive. It appears unlikely, if not impossible, that 

intact communities with all previously present species in healthy numbers can survive, but 

some agricultural crops can provide suitable habitats for many of the species native to our 

riparian communities. 
148 USDA Forest Service RMRS-GTR-411. 2020.

Example: Pecan Orchards as Riparian Habitats

Kingsley (1989, 1985a,b) studied wildlife uses of a large irrigated pecan orchard 

along the Santa Cruz River south of Tucson (fig. 25). At the time, this was said to be the 

world’s largest irrigated pecan orchard and consisted of approximately 2,200 hectares 

with tree densities varying from 59 to 119 trees per hectare. Fields were flood irrigated 

approximately every 2 weeks during the growing season (April-October typically). Trees 

were even-aged, tall, with the understory cleared of most vegetation. Stumps of removed 

trees sometimes sprouted and created a shrub story and a mixture of forbs and grasses 

were the understory. Dense tall weeds, including several species of seed-providing 

forbs and grasses, occurred in places along irrigation ditches and tailwaters. Slash 

piles consisting of pruned branches surrounded many fields and provided a structural 

equivalent of a dense shrub story that was utilized by some wildlife species. 

Insects were abundant, most of them considered beneficial or neutral, although 

aphids that feed on pecans and mosquitoes that bite people were present, and outbreak 

populations had caused problems (Kingsley 1985a). Many aquatic insect species migrated 

into the flooded fields and irrigation ditches but probably could not survive and reproduce 

successfully since these were temporary waters. Amphibians were present and abundant, 

but the only species documented as successfully reproducing was Couch’s spadefoot 

(Scaphiopus couchi), which can develop in transitory water lasting less than 2 weeks. 

Reptiles were relatively rare, probably because they cannot survive and reproduce under 

conditions of frequent inundation, absence of topographic diversity, or lack of some 

microhabitat components (Jones and Glinski 1985). 

Kingsley (1989) listed 83 species of birds found in the orchard with populations 

ranging from rare and transitory to greater than those found in native riparian woodlands. 

Flycatchers and small insectivores were present, but woodpeckers were absent as nesting 

birds although frequently seen foraging in the orchard. Birds characteristically found in 

dense riparian thickets were not found. No dense midstory was present. Hole-nesting 

birds were not found breeding, but several were observed foraging. Tree holes were not 

allowed to develop. 

Predatory mammals such as coyote (Canis latrans) and raccoon (Procyon lotor) 

and bats of several species were abundant as foragers, but burrowing mammals were 

generally absent, probably because they could not survive frequent inundation. No survey 

of mammals was conducted, so the number of small mammals is unknown. Bats foraging 

in pecan orchards in Texas were studied by Braun de Torrez and others (2011) and Braun 

de Torrez and Kunz (2009), who found bats present in pecan orchards, especially organic 

orchards, at higher numbers and diversity than in native juniper-mesquite woodlands and 

the adjacent semiarid landscape. 

The orchard has been a popular place for local birders, especially those on their way 

to Madera Canyon, an especially popular area that is reached by a road passing through 

the orchard. The orchard and processing plant supported approximately 250 workers 

and was one of the largest industries in the area at the time. Much of the historic orchard 

has been converted to residential development as the communities of Sahuarita and 

Green Valley have grown, and the long-term plan is to phase out the orchard except in 

floodplain-zoned sites where agriculture may continue indefinitely. 
USDA Forest Service RMRS-GTR-411. 2020. 149

Figure 26—Century-old irrigation canal, still in use, with cottonwoods. Originally put in to irrigate an agricultural field, the canal 

is now part of Zion National Park. Cottonwoods apparently invaded after the canal was created and are now well established. 

Photo by author.

Figure 25—Flooded pecan orchard, Green Valley Pecans, Arizona. Photo by author.
150 USDA Forest Service RMRS-GTR-411. 2020.

Example: Unlined Irrigation Canals as Riparian Habitat

That irrigation canals support a new riparian resource has long been considered (fig. 

26). Breninger (1898 p. 128), referring to the cactus ferruginous pygmy-owl, stated: 

“And since trees planted by man have become large enough to afford nesting sites for 

woodpeckers, this Owl has gradually worked its way from the natural growth of timber 

bordering the rivers to that bordering the banks of irrigating canals until now it can be 

found in places ten miles from the rivers.”

Recently, Sueltenfuss and others (2013) examined such wetlands in northern Colorado 

using quantitative and qualitative measures. The wetlands they examined were marsh, wet 

meadows, and salt flats. Eighty-nine percent of the canals examined were unlined earth. 

They found wetlands associated with irrigation canal leakage, pond and reservoir dam 

seepage, wetlands located on the fringes of reservoirs, tailwaters, ephemeral channels 

adjacent to canals, and a few intentionally created wetlands. They found large fluctuations 

in water level in wetlands to be associated with the amount of water flowing in canals, 

sometimes with delays of up to 50 days, and not associated with precipitation events. 

They found one instance in which irrigation of an adjacent field was associated with water 

table levels in a wetland. 

The same types of wetlands have been documented in southern Arizona (K.J. 

Kingsley, personal observations, and Kingsley 1985a), and some have cottonwood trees 

or other trees and shrubs growing along and watered by them. 

Johnson (1972) presented a review of changes occurring along irrigation ditches 

in the Phoenix area, where ditches were being converted to concrete-lined conduits or 

underground pipes that would not support vegetation. Water conservation by lining or 

burying conduits has resulted in loss of these riparian resources and is likely to continue 

and expand as demand for water increases and availability decreases. Many of the 

irrigated farmlands subsequently are converted to urban and suburban communities that 

may be very different ecosystems.

Example: Tailwaters

Tailwaters are wetlands created by excess irrigation and are wet areas at the low 

points of irrigated fields or immediately downstream from the fields. Along the Gila 

River in southern Arizona, some tailwater wetlands have dense thickets of tamarisk 

(Tamarix spp.) that are known to be used by nesting southwestern willow flycatchers 

(Empidonax trailii extimus) (K.J. Kingsley, personal observations). Tailwaters appear to 

be better habitat for this species than areas along the nearby river. Water moves slowly 

in these areas and provides abundant habitat for mosquitoes, which are fed upon by this 

endangered species, whereas the river moves too quickly to provide good mosquito 

habitat. In addition, tailwaters are fairly stable environments, compared to the unstable 

river flows, and are not as susceptible to flooding with concomitant loss of vegetation 

structure. Water is provided even during the worst droughts when water might not be 

available in the river. Several other species of riparian and wetland birds may also be 

observed in this environment, but the typical tailwater is an unstable habitat with a dense 

thicket of weeds or weed trees that is kept in an early stage of succession and affords little 

habitat for species that prefer large blocks of habitat or more open spacing. 
USDA Forest Service RMRS-GTR-411. 2020. 151

Example: Abandoned Fields Persisting as Riparian Habitats

Agriculture is not always successful or permanent, and when it fails, it leaves behind 

land that will transform from the regular rows of monoculture crops to a hodge-podge of 

early succession if left to itself. More than 96,356,000 ha of farmland were abandoned 

in the United States in the 20th century (U.S. Department of Agriculture 1998 cited in 

Richter and Stutz 2002). Especially in riparian areas, however, abandoned agricultural 

fields can be returned to some semblance of the former natural community. Sometimes, 

if the area is subject to a more-or-less natural flow regime and flood timing is right, 

and especially if the water table has not dropped to a depth inaccessible to roots, it can 

grow back to cottonwood-willow forest or woodland (e.g., several historic fields along 

the Verde River in the Cottonwood-Camp Verde area, now included in the Verde River 

Greenway State Natural Area, K.J. Kingsley, personal observations). Some efforts have 

been made to encourage or restore riparian communities to abandoned farmland. 

Of some concern in abandoned agricultural lands is contamination by chemicals, 

including undocumented historic chemical dumps. Other concerns are reduced water 

input because of increased efficiency by active agriculture, unnatural flow regime, longterm unpredictability, and possible attraction or harboring of agricultural pests. To my 

knowledge, none of these has been documented.

Conclusion

Little scientific attention has been directed toward the study of anthropogenic riparian 

areas, especially those that have been inadvertently created. Many more examples could 

have been selected from a wide range of sites. The number, diversity, and quality of 

anthropogenic riparian areas are likely to change over time as human demand for water 

resources increases with increasing population and aspirations for amenities. These 

environments warrant further study and some measures of protective management to 

enhance their suitability as wildlife habitat and recreational resources. 

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restoration of abandoned agricultural fields. Restoration Ecology. 10 (4): 607-616.

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semiarid Southwest. In: Batzer, D.P.; Baldwin, A., eds. 2012. Wetland habitats of North America: 

Ecology and conservation concerns. Berkeley, CA: University of California Press: 343-356.

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U.S. Department of Agriculture, Office of Communications. Cited in Richter and Stutz (2002). 

269 p. 

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environmental history of an Arizona river. Tucson, AZ: University of Arizona Press. 296 pp.

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orchard and riparian communities. Journal of the Arizona-Nevada Academy of Science. 14: 53-58.

Willott, E. 2004. Restoring nature, without mosquitoes? Restoration Ecology. 12: 147-153. 
154 USDA Forest Service RMRS-GTR-411. 2020.

Chapter 7. The Development of Riparian Ecosystem 

Restoration in California 

John Stanley, F. Thomas Griggs, and John Rieger

Introduction

The evolution of our modern-day understanding of riparian ecology and the 

development of the field of riparian ecosystem/habitat restoration underwent significant 

advances from 1970 to 2000. The four parts of this chapter include: (1) a chronological 

overview of the major activities, events and publications that documented, and 

punctuated, the progress of our understanding of riparian ecology and riparian ecosystem 

restoration in California; (2) a discussion of the various types of riparian restoration 

projects with the projects categorized by the primary reason they were undertaken; (3) 

examples of the various types of research, experimentation, field investigations, and 

monitoring programs associated with the early riparian habitat restoration projects; and 

(4) concluding remarks.

The Society for Ecological Restoration (SER), incorporated in 1988, defines 

“ecological restoration” as “the process of assisting the recovery of an ecosystem that 

has been degraded, damaged, or destroyed” (SER 2004). We use the term “restoration” 

in this chapter in a broad sense. We use the same terminology as used by the early 

“restorationists” to describe their projects even though many of the early projects do not 

meet the exact intent of SER’s definition. For example, many of the early projects were 

described as rehabilitation, enhancement, creation/fabrication, reclamation, ecological 

engineering, etc., rather than “ecological restoration” and/or “habitat restoration.” 

In the broadest sense, restoration activities have been occurring within California’s 

rivers, creeks, and riparian ecosystems for centuries. The cultural stewardship practices of 

California Indians directed at promoting ecological services also promoted continuance 

of the ecosystems (rivers, creeks, riparian areas) upon which they relied. Anderson 

(2005) documents a number of traditional cultural practices and horticultural techniques 

(traditional resource management) used by California Indians within the riparian 

corridor. For example, Stevens (2003) studied the relationship between California Indian 

groups that tended white root—an herbaceous perennial understory plant in valley oak 

riparian woodlands with long roots used for basket weaving—and how they affected the 

distribution and ecology of the plant. Unfortunately, the traditional ecological knowledge 

of California’s Indian Tribes was seldom, if ever, incorporated into riparian area 

restoration projects implemented in the 20th century1

.

8

We recognize that prior generations of landowners performed remedial work on their 

properties (e.g. stabilization of eroding streambanks using woody live and dead plant 

material) within the riparian corridor. In addition, workers employed under the Works 

Progress Administration between 1935 and 1942 and Civilian Conservation Corps between 

1933 and 1940 performed various tasks along rivers and streams in California (Riley 1998).

1 An example of how this might occur in the 21st century is the partnership developed in 1997 between the U.S. Forest Service Lake Tahoe 

Basin Management Unit (LTBMU) and the Washoe Tribe of Nevada and California to conduct the Meeks Meadow Washoe Restoration 

Project to restore the ecological and cultural function of the Meeks Meadow.
USDA Forest Service RMRS-GTR-411. 2020. 155

The field of riparian ecosystem/habitat restoration did not evolve solely on its own, 

but rather, in combination with the development of the fields of: (1) stream restoration; 

(2) erosion control; 3) (water quality/nonpoint source pollution control; (4) biotechnical/

soil bioengineering; (5) watershed management; (6) landscape ecology; and (7) greenway 

design and management. Practitioners attempting to restore riparian areas drew from the 

experience of professionals involved in these related activities and interacted with many 

of these professionals during the course of restoration project planning and design.

There is an extensive body of gray literature associated with the field of riparian 

restoration: project proposals, conceptual plans and detailed planning documents, project 

plans and specifications, as-built plans, monitoring programs, and monitoring reports. 

We avoided referencing these documents since they are difficult to locate. In almost all 

instances the literature cited herein is from books, conference proceedings, and other 

available documents.

Over the years, a number of conferences were convened in California to address 

issues related to the conservation and restoration of riparian ecosystems. Conferences and 

symposia pertaining to riparian ecosystems, riparian ecology, riparian habitat restoration, 

and riparian area conservation throughout the western United States including California 

are presented in Volume 1 Appendix B. 

Part 1—Publications, Activities, and Events: 1960s to 2000s

The 1960s

Smith (1977) stated: “Prior to 1960, few people showed any concern for the demise 

of California’s Riparian Forest communities.” Warner and Hendrix (1984) asserted 

that “descriptions of historical extent and character of the Sacramento Valley’s riparian 

systems by Kenneth Thompson (1961) were among the first writings to demonstrate their 

importance.”

The California Department of Water Resources (CWDR) recounted, “During 1961, a 

flood of protests against the denuding of levees in the Sacramento-San Joaquin Delta was 

launched by sportsman’s organizations, wildlife conservation groups, and the public in 

general.”( At that time, “Levee maintenance regulations dictated that virtually all shrubs 

and trees be cleared from levees to insure (sic) that the flood control protection provided 

by the levees would not be impaired” (CDWR 1967). 

The CDWR (1966) reported: “In 1961, the California State Legislature authorized 

the Sacramento River and Delta Recreation Study. One of the recommendations was 

that ‘a program of pilot studies on selected reaches of levee be initiated to test various 

types of vegetation, determine control measures necessary, study methods for these 

controls, and determine the costs of this type of maintenance’. In response to that 

legislative recommendation, the Department of Water Resources initiated the Pilot Levee 

Maintenance Study in 1962.” 

Thus, “The Pilot Levee Maintenance Study was begun to conceive and test alternative 

methods of levee maintenance that would provide for multiple use of levees. A number of 

tests29were conducted to determine if vegetative growth could be allowed on levees and 

2

 Refer to part 2 below (Riparian Restoration Projects–Bank Stabilization/Erosion Control Projects) for a brief discussion of the types of 

tests conducted and the types of plant materials used in these tests.
156 USDA Forest Service RMRS-GTR-411. 2020.

berms to preserve and enhance esthetic, recreational and wildlife habitat values without 

impairing the flood control capacity of the levees” (CDWR 1967). 

Two annual progress reports were prepared in 1963 and 1964 and a Preview to 

Bulletin 167 was published in 1966 (CDWR 1966). Findings during the last year of 

study and the conclusions and final recommendations of the Pilot Levee Maintenance 

Study were presented in CDWR Bulletin 167 (CDWR 1967). The study “concluded that 

with proper vegetative management programs, certain delta levees can be adapted and 

maintained to serve the needs of esthetics, recreation and wildlife as well as the primary 

purpose of flood control.”

One of the earliest acknowledgments of the ecological values and severely limited 

extent of riparian systems appeared in Volume III of the 1965 California Fish and Wildlife 

Plan (CDFG 1965 per Warner and Hendrix 1984. Smith (1977) cited this document when 

he stated that “of the 29 habitat types listed in the ‘Inventory of Wildlife Resources, 

California Fish and Wildlife Plan’ (Vol. III), riparian habitat provides living conditions for 

a greater variety of wildlife than any other habitat type found in California.”

The 1970s

The 1970s were the period of heightened awareness of riparian habitat loss and the 

importance of riparian habitat for many of the wildlife species in the State of California. 

In An Island Called California, Elna Bakker pointed out that “no natural landscape 

of California has been so altered by man as its bottomlands. The grass-rich stretches 

of the great Central Valley are, for the most part, lost to orchards and vineyards, cotton 

and alfalfa fields. Many miles of curving green ribbon along its watercourses have been 

eradicated, replaced by the sterile concrete of flood control and navigation channels. Most 

of the tule marshes of the Delta country are now neatly diked rice paddies. On the freeway 

between San Francisco Bay and Sacramento one forgets that this was once wild land with 

golden beavers going about their industrious ways and great blue herons on guard with 

that watchful immobility so peculiarly their own. To recreate this world of slough, bank, 

and riverway takes more than the simple listing of what can be recalled, or guessed, was 

there. It needs imagination coupled with a persistent searching for the last few remnants 

of the original river country” (Bakker 1971).

Concerns arose over the loss of riparian habitat and the degradation of remaining 

riparian areas due to land clearance for orchards and field crops, logging for wood 

products, grazing by livestock, streambank stabilization, channelization and other flood 

control activities, and altered water flow due to dams and irrigation throughout California 

(especially California’s Central Valley). Consequently, scientists, conservationists, and 

agency personnel came together to share their observations, express their concerns, and 

seek solutions in the early and mid-1970s. 

 Warner and Hendrix (1984) stated: “In the early 1970s, studies by the Nongame 

Wildlife Investigations Section of the Department of Fish and Game (e.g., Gaines 

1974) began to bring into focus the impact of riparian system loss to the State’s wildlife 

populations. At about the same time, disturbing figures of riparian vegetation loss along 

the upper Sacramento River were reported (e.g., McGill 1975).” Burns (1979) reported, 

“Based on historical accounts, there were nearly 775,000 acres of riparian forests along 

the Sacramento River and its tributary streams in 1850.” Others (Katibah 1984; Roberts et 

al. 1977) cite a figure of 800,000 acres of riparian forest remaining after 1848. By 1952, 
USDA Forest Service RMRS-GTR-411. 2020. 157

only about 20,000 acres of riparian forest remained (Smith 1978). During the 20-year 

period between 1952 and 1972, 53 percent of the mature riparian forests that existed 

in 1952 had been removed and the land converted to agricultural uses (Burns 1979). 

By 1972, only about 12,000 acres of riparian forest remained in the Sacramento Valley 

(Gaines 1976; Roberts et al. 1977). Those interested in greater detail on the decline of 

California’s Central Valley riparian forests should refer to Katibah (1984).

In the fall of 1975, the California Secretary of Resources established the Upper 

Sacramento River Task Force “to solve the acute resource problems centered primarily 

along the 170-mile section of the river below Shasta Dam” (Burns 1979). The initial 

task force was made up of State agencies310within the Resources Agency. Soon after its 

formation, the task force was enlarged to include Federal and local agencies and later 

enlarged again to include special interest groups. The objectives of the task force were 

“to coordinate intergovernmental activities and to take actions to ensure the protection 

of the fish, wildlife, recreation, and aesthetic values of the river while considering the 

other beneficial used of the river and adjacent lands, for such uses as water conveyance 

and agriculture,” (Burns 1979). Encouraged by Napa County’s 1974 adoption of an 

ordinance allowing for the protection of riparian vegetation along water courses (Burns 

1977; Dunlap 1977; Gaines 1976), the task force “drafted a model, county general plan 

element, and ordinance that would bring the removal of riparian vegetation under a permit 

process,” (Burns 1979). The task force gave these models to the boards of supervisors of 

five counties along the Sacramento River and asked them to adopt similar regulations. 

Only the northernmost county adopted a riparian protection ordinance (Burns 1979).

The first riparian forest conference in California titled “Conference on the Riparian 

Forests of the Sacramento Valley” was held in Chico in May of 1976. Cosponsored 

by the Davis and Altacal Audubon Societies, the conference was organized by David 

Gaines. There were about 70 participants (Abell 1989). There were no published 

conference proceedings; however, David Gaines prepared abstracts of the presentations 

(Gaines 1976). During this one-day conference, speakers gave a historical perspective 

on Sacramento Valley’s riparian forests, described riparian vegetation, and discussed the 

animals that rely on riparian habitat. Staff from the CDWR described land use changes in 

the Sacramento River riparian zone and discussed the activities of the Sacramento River 

Task Force. A representative from the Army Corps of Engineers described Corps project 

work for flood control and bank protection on the Sacramento River. The President of 

the Sacramento Valley Landowner’s Association presented a landowner’s perspective 

of the bank erosion and loss of agricultural land due to the management of the river for 

water transport. CDWR staff described the recreational values of the Sacramento River 

and State Department of Parks and Recreation staff addressed boating on the river. A 

land agent with the California Wildlife Conservation Board discussed riparian forest 

acquisition along the Sacramento River, a supervisor from Tehama County discussed 

the county’s approach to riparian forest habitat, and a U.S. Fish and Wildlife Service 

field supervisor addressed the need to mitigate habitat loss due to bank protection. 

Summing up, an attorney encouraged everyone to fight for their vision of the future of the 

Sacramento River and its riparian forests.

3

 Department of Fish and Game, Department of Water Resources, Department of Parks and Recreation, Department of Navigation and 

Ocean Development, Wildlife Conservation Board, Reclamation Board, Water Resources Control Board, and State Lands Division.
158 USDA Forest Service RMRS-GTR-411. 2020.

The second riparian forest conference in California titled “Riparian Forests in 

California: Their Ecology and Conservation” (Sands 1977) was held at the Davis Campus 

of the University of California in May 1977. Cosponsored by the Institute of Ecology, 

University of California, Davis, and the Davis Audubon Society the “symposium” was 

coordinated by Anne Sands. This was the first riparian conference for the United States 

with published proceedings. There were approximately 128 people in attendance (Abell 

1989). Symposium topics were divided into two sessions; the morning session dealt 

with historical and ecological subjects followed by an afternoon session dealing with 

management and preservation. Presentations in the first part of the morning addressed the 

historical extent of riparian habitat loss (Roberts et al. 1977; Smith 1977), the diverse flora 

and fauna dependent on riparian areas (Roberts et al. 1977), accounts of the historical 

condition of Sacramento Valley riparian lands and remnants (Thompson 1977), and the 

concepts necessary for understanding the fluvial system (Keller 1977). 

Following the morning break, experts went into greater detail describing the 

vegetation/flora of the Sacramento Valley (Conard et al. 1977), the importance of valley 

riparian forests to bird populations (Gaines 1977), and the habitats of native fishes in the 

Sacramento River Basin (Alley et al. 1977). Afternoon speakers addressed the need for 

legislation to study and protect California’s riparian forests (Dunlap 1977), the activities 

of the U.S. Army Corps of Engineers (CE) protecting the Sacramento Valley levee system 

with rock bank protection (Kindel 1977), and the planting of trees and shrubs at selected 

sites along the Sacramento River (Kindel 1977). Speakers also described the progress 

made by the recently established Upper Sacramento River Task Force in coordinating the 

activities of the many governmental agencies that have jurisdiction over developments 

along the Sacramento River so as to ensure the protection of the fish and wildlife and 

recreational aspects of the river (Burns 1977). The symposium closed with a brief 

presentation on the value of riparian forests in today’s society (Frost 1977) followed by a 

panel discussion.

It is noteworthy that no riparian restoration projects were mentioned, nor were the 

words “restoration” or “revegetation” used in any of the papers presented at the 1977 

Symposium. The only mention of planting is by the Corps of Engineers—planting of trees 

and shrubs in 1967 along 3 miles of riverbank where the levee had been set back and a 

new berm had been created and protected by rock.

Anne Sands and Greg Howe presented a paper titled “An Overview of Riparian 

Forests in California: Their Ecology and Conservation” (Sands and Howe 1977) at 

the Symposium on the Importance, Preservation and Management of Riparian Habitat 

(Johnson and Jones 1977) held in Tucson, Arizona, in July 1977. This paper was 

comprised of abstracts from Sands (1977).

In 1978, Anne Sands (1979) presented a paper titled “Public Involvement in Riparian 

Habitat Protection: A California Case History” at a floodplain wetlands and riparian 

ecosystem symposium in Georgia. Anne Sands addressed the restoration411of damaged 

riparian areas as “probably the most difficult protection strategy, but certainly it is one of 

the most rewarding” (Sands 1979). In this paper, Anne Sands (1979) provided an example 

of how public concern over riparian habitat loss due to flood control project works had 

resulted in the State Reclamation Board commissioning a civil engineering study (Murray 

4

 This is possibly the first mention of “riparian restoration” in the literature.
USDA Forest Service RMRS-GTR-411. 2020. 159

et al. 1978) on the retention of riparian vegetation as a means for controlling bank erosion 

along the upper Sacramento River. 

Warner and Hendrix (1984) reported: “In 1978, a coalition of conservation 

organizations, led by the Riverlands Council chaired by Anne Sands, sponsored State 

legislation to protect riparian resources. The resulting legislation, AB 3147 (Fazio), 

appropriated $150,000 to the Department of Fish and Game (CDFG) for a study of 

riparian resources of California’s Central Valley and desert.” The stated goal of the CDFG 

in mounting its California Riparian Study Program (CRSP) was to “protect, improve, 

and restore the riparian resources of the State” (Warner 1984). There were multiple 

elements to the CRSP including: development of background information on California’s 

riparian resources; mapping of riparian vegetation in the Sacramento and San Joaquin 

valleys; determination of the areal and linear extent of Central Valley riparian vegetation; 

conducting a remote-sensing survey and a ground inventory of riparian vegetation in the 

Central Valley; field surveys of California’s desert riparian systems; reporting on State, 

Federal, and local programs affecting riparian systems; analysis of riparian conservation 

needs; and the development of a riparian conservation program. 

The 1980s

In 1981, the California Reclamation Board, a governor-appointed body with statutory 

responsibility for maintaining Central Valley floodways, adopted a Riparian Vegetation 

Management Policy (King 1985). This policy recognized the benefit of riparian vegetation 

in maintaining the integrity of floodways and established a permit system for its removal. 

The Reclamation Board initiated a program, managed by the CDWR to ensure the 

retention of riparian vegetation at selected sites along the Sacramento River. Beginning 

in 1981, CDWR began installing native trees and shrubs in rock reinforced levees 

downstream of Sacramento (King 1985).

The third riparian forest conference in California titled “California Riparian 

Systems12: Ecology, Conservation and Productive Management” (Warner and Hendrix 

1984) was held at U.C. Davis in September 1981. This conference drew 711 participants 

(Abell 1989). The fact that approximately 150 technical papers were presented at the 

conference of which 128 were included in the 1,035 pages of conference proceedings 

is evidence of the exponential growth in the fields of riparian ecology, conservation, 

management, and restoration. Although only seven papers were presented under the 

category of riparian restoration, many of the other papers reported on critical background 

research pertaining to the structure and function of riparian systems, especially hydrologic 

and hydraulic considerations. Also at this conference, Richard Warner (1981) reported 

on the structural, floristic, and condition inventory of Central Valley riparian systems 

conducted as part of the CRSP mentioned above (Warner 1984).

In the early 1980s, Randal L. Gray and Ron Schultze of the USDA Soil Conservation 

Service (now the Natural Resources Conservation Service) in Davis, California, 

organized the Riparian Revegetation Study Group (RRSG) to bring together individuals 

working on, or interested in, the reestablishment of riparian vegetation along degraded 

streams in California. Beginning as an interagency group, RRSG quickly expanded to 

5

 This was the first conference convened at the University of California, Davis, under the “California Riparian Systems” title (Abell 1989).
160 USDA Forest Service RMRS-GTR-411. 2020.

include Federal, State, and local agency personnel, staff of non-profit organizations, and 

consultants. Semiannual meetings were hosted by members in San Jose, Hayward, Marin 

County, and Sacramento. The agenda for most meetings included short presentations of 

restoration projects and discussions of difficult issues associated with the restoration of 

riparian areas (for example, vandalism). Some meetings included field tours of recently 

completed revegetation projects. A subgroup of RRSG was organized in southern 

California because travel time and distance to meetings held in the northern part of 

the State precluded the involvement of most southern Californians. RRSG became the 

Riparian Guild of the California Society for Ecological Restoration (SERCAL) after 

SERCAL’s formation in 1991.

Several of the papers presented at the Native Plant Revegetation Symposium (Rieger 

and Steele 1985) held in San Diego in November 1984 dealt with the restoration of 

riparian areas (Barry 1985; Capelli 1985; King 1985; Riley and Sands 1985). Two 

hundred and two people were in attendance. This meeting was the first that addressed 

restoration as the theme of the symposium. It was an effort to survey and promote 

communication between restoration practitioners in the State.

In the late 1970s and early 1980s, it became apparent that the protection and 

restoration of riparian areas was in direct conflict with traditional floodplain management 

(Riley and Sands 1985). Conservationists encouraged Federal and State flood 

management agencies and local flood control districts to find ways in which the retention 

and/or planting of riparian vegetation within channels and on floodplains could be 

compatible with flood management. At the 1984 “Native Plant Revegetation Symposium,” 

Ann Riley and Anne Sands (1985) suggested that flood control planners and engineers 

should adopt the concept that “restoring streams is environmentally less damaging and 

less expensive than traditional channelization,” an approach that was being promoted 

by Nelson Nunnally and Ed Keller (1979). Nunnally and Keller’s central theme was “to 

preserve natural stream morphology wherever possible, and when change is essential, to 

design channels with morphological characteristics similar to those of natural channels.” 

Riley and Sands (1985) argued for restoring the “natural balance between the river, its 

floodplain, and the riparian forest” and that to restore and maintain the riparian forest/

fluvial system balance, we must recognize that floodplains must be allowed to function as 

they were intended. Riley and Sands (1985) also touted the combined use of vegetation 

and structural means for bank stabilization as presented in Gray and Leiser (1982) 

and Scheichtl (1980). Williams and Swanson (1989) proposed that flood management 

agencies should take a new approach to planning channel modifications for flood damage 

reduction. They pointed out that there were significant problems with conventional flood 

control project design including: (1) underestimation of the roughness of lined channels; 

(2) failure to account for channel bed erosion and deposition; (3) failure to account for 

debris; and (4) underestimation of maintenance requirements. Williams and Swanson 

argued that flood channel design should be multi-objective and should incorporate proper 

consideration of hydrologic, geomorphic, and biological factors that influence stream 

hydraulics. They pointed to the Wildcat Creek Flood Control Project as an opportunity to 

use the integrated design approach they discussed in their paper.

During the early- to mid-80s, some flood control agencies sponsored the preparation 

of design-it-yourself manuals so that flood control planners, engineers, and landscape 

architects could make decisions on the species to be planted along reconfigured channels 
USDA Forest Service RMRS-GTR-411. 2020. 161

and their appropriate planting locations without the input of biologists on each project 

site. In 1983, Harvey and Stanley Associates, Inc., prepared a “Revegetation Manual 

for the Alameda County Flood Control and Water Conservation District Revegetation 

Program” (Stanley and Stiles 1983). The manual contained detailed information 

on the characteristics of 97 species of riparian plants including their appearance, 

ecological relationships, wildlife habitat value, planting location, planting options, plant 

requirements, and maintenance requirements. A map and table allowed project planners 

to select plant species suitable to each of the planting zones within Alameda County. This 

was to be used in conjunction with a representative cross section of a typical channel 

and an accompanying table indicating the appropriate planting zone (streambed, toe 

of channel, lower, middle, and upper slope, top of bank, outside levee slope) for each 

species in order to select the appropriate native plants for each planting location. It was 

also possible to sort for native plant species to be used, or avoided, in special situations 

(for example, plants that are invasive, fire resistant, less than a certain height, colorful, 

a barrier to access, evergreen, good for erosion control, high in wildlife food and shelter 

value, and providing screening).

In 1984, the Urban Creeks Council sponsored the Urban Creek Restoration and 

Flood Control Act which was signed by California’s Governor in September 1984 

(Riley and Sands 1985). The Urban Streams Restoration Program began in 1985 and 

was administered by the CDWR. “The purpose of the program is to provide grants 

and technical assistance to those local governments and community groups that want 

to implement less costly and more environmentally sensitive responses to erosion and 

flooding problems,” said Riley (1998). Over the next 10-year period, “the program funded 

160 alternative restoration projects, including: innovative bank stabilization projects using 

live and dead plant materials; innovative channel design to increase flood capacities; 

culvert removal and stream daylighting (sic) to correct storm-water management 

problems; and land acquisition solutions to reduce flood damages.” 

In May 1985, the CDFG released a Final Draft of “Riparian Resources of the Central 

Valley and California Desert” (Warner and Hendrix 1984). This report was the final 

product of Phase II of the CRSP that began in 1978 (see above). The report: (1) examined 

the structure and dynamics of riparian systems; (2) summarized the attributes, values, and 

vulnerabilities of riparian systems; (3) quantified their historical and present distribution 

in the Central Valley; (4) presented the major findings of field studies on riparian system 

distribution, structure, and condition in the Central Valley; (5) examined the nature 

and problems of desert riparian systems; (6) reviewed riparian resource conservation 

mechanisms available through Federal, State, and local laws, regulations, and programs; 

and (7) proposed a series of actions to reverse the chronic, long-term trends of riparian 

resource decimation and to restore some of these systems to their former status as 

productive major ecological elements in the California landscape.

In 1986, Aqua Resources Incorporated and Holton Associates (1986) prepared 

a “Riparian Planting Design Manual for the Sacramento River: Chico Landing to 

Collinsville” for the U.S. Army Corps of Engineers, Sacramento District. The purpose of 

the manual was to guide Corps personnel in designing plantings as mitigation for bank 

protection projects, mainly riprap, along the Sacramento River. The manual provided 

a list of recommended plant species for each of four plant communities (willow scrub, 

cottonwood riparian forest, mixed riparian forest, and valley oak riparian forest), 
162 USDA Forest Service RMRS-GTR-411. 2020.

which naturally occur in each of four hydrologic zones (river channel, low terrace, 

high terrace, upper high terrace) along the Sacramento River. The manual defined the 

planting zone for each plant community based on the average annual flood duration 

for each hydrologic zone. The manual also provided information on recommended 

planting densities and spacing patterns and guidelines for the preparation of revegetation 

plans and specifications. An appendix contained plant data sheets for 26 woody species 

recommended for riparian plantings (Granholm et al. 1988)

The “Second Native Plant Revegetation Symposium” was held in San Diego in April 

1987 (Rieger and Williams 1988). In attendance were 257 people. Six papers dealing 

with riparian habitat restoration in California were presented at the conference. Results 

of several riparian revegetation projects implemented by the California Department of 

Transportation were reported as well as several private development mitigations.

The fourth riparian forest conference in California titled “California Riparian Systems 

Conference6

:

13Protection, Management, and Restoration for the 1990s” (Abell 1989) 

was held at the University of California, Davis, in September 1988. This conference 

drew nearly 900 participants. Thirteen papers were presented in the session titled 

“Implementing Revegetation Projects” and six papers were presented in the session titled 

“Urban Streams.” Other sessions dealing with channel dynamics, rangeland and desert 

riparian systems, and coastal streams also addressed the restoration of riparian areas.

In 1988, Faber and Holland (1988) published Common Riparian Plants of 

California: A Field Guide for the Layman. This book of photocopies of riparian plants 

and information on their characteristics and where they typically occur (distribution and 

elevation) was used by conservationists and community-based organizations seeking to 

restore riparian areas.

In January 1989, the Society for Ecological Restoration (SER) held its first annual 

meeting in Oakland, California. The conference proceedings titled “Restoration ’89: The 

New Management Challenge” (Hughes and Bonnicksen 1989) contained a section on 

the restoration of riparian areas and a section on stream restoration. Other professional 

societies and conservation organizations also convened conferences in the latter part of 

the 1980s that included papers pertaining to riparian corridor and riparian ecosystem 

restoration. Some of these groups that recognized the relationship between restoring 

riparian areas and achieving the mission of their organization included: Association of 

State Wetland Managers, Inc.; California-Nevada Chapter of the American Fisheries 

Society; Salmonid Restoration Federation; Society for Range Management; and the Urban 

Creeks Council.

In 1989, the U.S. Fish and Wildlife Service published The Ecology of Riparian 

Habitats of the Southern California Coastal Region: A Community Profile (Faber et 

al. 1989). The purpose of this publication was to describe the structure and function of 

riparian habitat in southern California. This biological report: (1) described the physical 

setting and geofluvial processes in riverine systems; (2) outlined the effect of water 

regime on the establishment and succession of riparian plant communities; (3) described 

the most common species of riparian plants; (4) detailed the fauna dependent upon and 

that uses the riparian habitat; (5) summarized riparian ecosystem processes and values; 

and (6) spelled out the myriad of governmental jurisdictions and relationships that 

6

 This was the second “California Riparian Systems Conference” convened by U.C. Davis Extension (Abell 1989).
USDA Forest Service RMRS-GTR-411. 2020. 163

affect the use of, and the ability to conserve, riparian habitat. The publication concluded 

with a section on riparian ecosystem restoration including a number of case studies. 

Unfortunately, no similar publication was ever produced for the riparian ecosystems of 

northern California.

Also in 1989, the United States Environmental Protection Agency published Wetland 

Creation and Restoration: The Status of the Science. Volume I (Regional Reviews) 

contained a paper on riparian wetland creation and restoration in California (Stanley 

1989). An appendix to this paper contained profiles for 37 riparian restoration projects in 

California.

The 1990s

By the 1990s, some biologists, ecologists, and hydrologists had gained enough 

experience conducting riparian restoration projects that they began offering training in 

riparian habitat restoration. The Wetland Training Institute, Inc., offered workshops on 

Riparian Habitat Restoration at various locations in California throughout the 1990s. 

SERCAL and SER offered Riparian Habitat Restoration workshops in connection with 

conferences held in California during the latter part of the 1990s and 2000.

In 1991, State legislation created the California Riparian Habitat Conservation 

Program (CRHCP) within the Wildlife Conservation Board (WCB). The mission of the 

program is to develop coordinated conservation efforts aimed at protecting and restoring 

the State’s riparian ecosystems. The goals of the CRHCP, as noted in its enabling 

legislation, are to protect, preserve, restore, and enhance riparian habitat throughout 

California. To accomplish the program’s objectives, the WCB was authorized to award 

grants for riparian conservation purposes (acquisition and restoration) to non-profit 

organizations, local government agencies, State departments, and Federal agencies.

Also in 1991, the California State Lands Commission published Delta-Estuary—

California’s Inland Coast: A Public Trust Report (Argent 1991). This report described the 

Delta’s geologic, hydrologic, biologic, and cultural history and the public trust uses that 

are dependent on these resources. Citing Madrone Associates (1980), the report stated 

that “riparian habitat is used by more vertebrate wildlife, 107 species, than any other 

Delta habitat type.” The report indicated that the clearing of levees for maintenance or 

placement of rock revetment has resulted in a severe loss of riparian habitat and shaded 

riverine aquatic habitat. The report concluded with a survey of the institutions and entities 

that manage the delta’s resources.

In 1992, the National Research Council’s Committee on Restoration of Aquatic 

Ecosystems concluded: “Given that healthy, vegetated riparian habitat and bottomlands 

are essential to the natural ecological functioning of associated streams and rivers—and 

are among the nation’s rarest habitats due to prior devastation—riparian habitat and 

bottomland restoration should be made a high national priority along with the restoration 

of the stream and river channel itself” (NRC 1992). The committee recommended that 

a national aquatic ecosystem restoration strategy be developed for the United States. By 

this time, California’s State agencies were already actively moving to preserve, conserve, 

and restore California’s rivers, creeks, and riparian ecosystems in cooperation with 

local agencies, academic institutions, non-governmental organizations, and professional 

consultants.
164 USDA Forest Service RMRS-GTR-411. 2020.

In 1992, representatives of 28 agencies gathered at the request of California 

Resources Secretary Douglas Wheeler. These agencies recognized that diverse programs, 

goals, missions, regulations, and geographic regions required diverse information to 

support decisions regarding the management and conservation of California’s rivers. They 

recommended a process that began with a survey of professional judgment of California’s 

river conditions. They continued with the accumulation, organization, and internet 

publication of a large and diverse body of facts and tools dedicated to the analysis and 

management of California’s rivers. The Information Center for the Environment (ICE) the 

University of California, Davis, conducted the California Rivers Assessment714(CARA) 

Project under the oversight of the WCB. ICE submitted the Final Report for the CARA 

Project titled California Riparian Habitat Inventory and Assessment (ICE 2009) to the 

WCB in April 2009.

The first phase of the CARA Project, a Professional Judgment Assessment (PJA), 

drew upon the knowledge, expertise, and opinions of resource managers, scientists, and 

other river experts to assemble a database of information about the condition of riparian 

and aquatic resources for California’s 196 largest rivers. Riparian criteria included the 

presence (or absence) of a natural flow regime, vegetation size and land coverage, the 

trend in riparian habitat distribution over the past 25 years, and the impact of human 

activities on the areas. The PJA succeeded in collecting information for 616 segments on 

145 rivers. Numerical scores for each river segment were distributed into quartiles labeled 

“Outstanding,” “Substantial,” “Moderate,” or “Limited.” A pilot comparative evaluation 

of the PJA responses was performed to illustrate how the PJA survey information could 

provide a Statewide perspective on the relative condition of riparian and aquatic river 

resources. In the final report, ICE (2009) indicated that out of California’s 172,000 miles 

of rivers, 13,631 miles were rated by the CARA PJA. Of the 13,631 miles rated by PJA, 

1,379 miles were rated “Outstanding” for riparian and 1,828 miles were rated “Limited” 

for riparian8

.

15.

The fifth riparian forest conference in California titled “California’s River Heritage9

:

16

A Conference on Conservation Issues, Policy and Implementation Strategies” (McCoy 

1992) was held in Sacramento in May 1992. Concurrent sessions focused on: (1) 

Wild, Scenic and Recreation Rivers; (2) River Corridors and Parkways; and (3) Urban 

Riverfronts, Creeks, and Streams. Several of the speakers described river corridor 

restoration programs wherein riparian habitat restoration was, or would become, a major 

component including: Putah Creek (Sanders 1992); San Lorenzo River (Hall 1992); San 

Luis Creek (Jones 1992); and San Joaquin River (Dangermond 1992). Several stream 

restoration projects that received funding from the Department of Water Resources Urban 

Stream Restoration Program were described by Earle Cummings (1992).

In 1993, the California State Lands Commission published California’s Rivers: A 

Public Trust Report (Argent 1993). The subject of this report was the condition of the 

7

 CARA is a computer-based data management system designed to give resource managers, policy-makers, landowners, scientists, and 

interested citizens rapid access to essential information and tools with which to make sound decisions about the conservation and use 

of California’s rivers. Cara contains 39 sets of mapped geographical information system (GIS) layers, 60 sets of tabular (database) and 

textual (text) data, as well as links to 510 additional maps, tables, and texts. All these data are organized by watershed and theme.

8

 These small numbers for the ranking of riparian and aquatic resources as “Outstanding” and “Limited” were the result of the limited 

number of miles of river segments evaluated. Funding was never obtained to complete the evaluation for the remainder of California river 

segments.

9

 This was the third “California Riparian Systems Conference” convened by U.C. Davis Extension.
USDA Forest Service RMRS-GTR-411. 2020. 165

rivers of California and their watersheds. The report documented the causes of their 

alteration and the nature and extent of their degradation; identified means by which 

their degradation could be avoided or reduced; and suggested measures to be taken 

for their restoration. The report addressed the need for riparian habitat restoration and 

explained the role of numerous Federal, State, and local agencies and organizations in the 

protection, management, and restoration of riparian areas. The report indicated that “in 

California, many riparian restoration projects have been implemented, but most are on a 

relatively small scale, rather than for whole systems” (Argent 1993).

Recognizing the importance of riparian habitat for landbirds in California, California 

Partners in Flight (CalPIF) initiated the Riparian Habitat Joint Venture (RHJV) Project 

in 1994. Eighteen Federal, State, and non-profit organizations signed a Cooperative 

Agreement to protect and enhance habitats for native landbirds throughout California. 

Modeled after the successful Joint Venture projects of the North American Waterfowl 

Management Plan, the RHJV reinforces other collaborative efforts that protect 

biodiversity and enhance natural resources as well as the human element they support. 

“The mission of the RHJV is to promote the protection, restoration and enhancement of 

riparian habitat sufficient to support the long-term viability and recovery of California’s 

native landbirds and other associated species” (Chrisney 2003).

The sixth riparian forest conference in California titled “California’s Riparian-River 

Ecosystems Conference IV:1017Addressing Current Land Use and Resource Conflicts” 

(Laird et al. 1995) was hosted by U.C. Davis in Sacramento in November 1995. Speakers 

provided a historical perspective of the physical and fluvial processes and riparian 

and aquatic resources of California’s rivers. There were updates on recent flooding 

and discussions of floodplain management. River management case studies addressed 

restoration of the Klamath, Trinity, Los Angeles, Russian, San Joaquin, and Sacramento 

rivers.

In October 1998, The Federal Interagency Stream Restoration Working Group, 

comprised of 17 Federal agencies, published Stream Corridor Restoration: Principles, 

Processes, and Practices (FISRWG 1998). Part II (Developing a Restoration Plan) provided 

suggested approaches for identifying problems and opportunities; developing goals, 

objectives, and alternatives; and planning the implementation, monitoring, and evaluation 

of restoration projects. Part III (Applying Restoration Principles) contained guidance 

on analysis of corridor conditions, restoration project design, and restoration project 

implementation, monitoring, and management. Although only one agency representative 

(USFWS) from California was on the Production Team for this document, much of its 

contents was in line with the approaches and practices already in use for stream and riparian 

restoration in California. This document was useful for those entering the field of stream 

and riparian restoration and managers and administrators of restoration projects.

The 2000s

The following noteworthy documents were published after 1999; however, they report 

on, or synthesize, knowledge that was generated in the latter part of the 20th century.

In August of 2000, CalPIF and RHJV released the first version of The Riparian Bird 

Conservation Plan: A Strategy for Reversing the Decline of Riparian Associated Birds 

10 This was the fourth “California Riparian Systems Conference” convened by U.C. Davis (McCoy 1995).
166 USDA Forest Service RMRS-GTR-411. 2020.

in California. “The Riparian Bird Conservation Plan seeks to synthesize and summarize 

the current state of scientific knowledge concerning the requirements of birds in riparian 

habitats. It provides recommendations for habitat protection, restoration, management, 

research, monitoring, and policy to ensure the long-term persistence of birds dependent 

on riparian ecosystems,” (RHJV 2004). “The RHJV chose to emphasize the ecological 

associations of individual species as well as those of conservation concern. In doing so, 

the RHJV included a suite of focal species whose requirements define different spatial 

attributes, habitat characteristics, and management regimes representative of a ‘healthy’ 

system”. (RHJV 2004). 

Version 2.0 of the Conservation Plan (RHJV 2004) expanded the original list of 14 

“focal” riparian bird species to 17 “focal species.” Criteria for the selection of focal species 

included: (1) use of riparian vegetation as their primary breeding habitat in most bioregions 

of California; (2) warrant special management status—endangered, threatened, or species of 

special concern on either the Federal or State level; (3) have experienced a reduction from 

their historical breeding range; (4) commonly breed throughout California’s riparian areas; 

and (5) have breeding requirements that represent the full range of successional stages of 

riparian ecosystems.

The seventh riparian forest conference in California titled “California Riparian 

Systems11:

18Processes and Floodplains Management, Ecology, and Restoration” (Faber 

2003) was convened by the Riparian Habitat Joint Venture in Sacramento in March 2001. 

The conference focused on riparian and floodplain processes, habitat management and 

restoration, monitoring, and partnerships in riparian area activities. Seventy papers were 

presented in this volume. Most of the 18 papers in Section III (Restoration) report on work 

performed in the 1980s and 1990s.

In October of 2003, the CDFG published additional elements to the third edition 

of the California Salmonid Stream Habitat Restoration Manual (Flosi et al. 1998). The 

new Part XI (Riparian Habitat Restoration) addresses measures for the conservation 

and management of riparian habitats and measures for the the restoration of native 

riparian habitats. It provides guidance on riparian revegetation project planning, provides 

information on the sources of native plant material, and discusses revegetation techniques. 

Appendix XI-A contains fact sheets for numerous central and north coast native riparian 

plants while Appendix XI-B contains fact sheets for a number of invasive nonnative plant 

species common to riparian areas.

The eighth riparian conference in California was held by the American Water 

Resources Association (AWRA) in Olympic Valley, California, in June of 2004. Titled 

“Riparian Ecosystems and Buffers: Multi-scale Structure, Function and Management” 

(Dwire and Lowrance 2006), the conference included papers on the role of riparian 

ecosystems and riparian restoration projects in protecting the water quality of the Lake 

Tahoe Basin. Key papers were published in a special issue of the Journal of the American 

Water Resources Association (JAWRA 2006).

The ninth riparian conference in California was convened in Sacramento by the 

Riparian Habitat Joint Venture in December of 2007. Titled “Riparian Habitat Conservation 

and Flood Management in California12”19(RHJV 2007), the conference proceedings 

11 This was the fifth “California Riparian Systems Conference.”

12 This was the sixth “California Riparian Systems Conference.” It was mistakenly referred to as the “Fourth Conference on California 

Riparian Systems” in the introduction to the proceedings.
USDA Forest Service RMRS-GTR-411. 2020. 167

contained nearly 80 papers and discussions focused on integrating levee management, flood 

protection, riparian conservation, and wildlife protection so as to sustain a safe, vibrant, 

and healthy environment along the streams and rivers of California. The conference was 

specifically designed to address the critical need for restoring the essential biodiversity of 

riparian areas while also ensuring good water quality and flood safety.

In 2007, RHJV partners identified a need for guidelines for planning and 

implementing riparian restoration projects on the ground. RHJV convened a group 

of restoration experts for a workshop to produce a handbook of restoration strategies, 

standards, and guidelines. The goal was to provide practitioners, regulators, land 

managers, planners, and funders with basic strategies and criteria to consider when 

planning and implementing riparian conservation projects. River Partners, a RHJV 

partner, took the lead in developing this handbook. In July 2009, River Partners published 

the second edition of the Riparian Habitat Restoration Handbook (Griggs 2009). The 

goal of the handbook is to explain the proposal/planning process for a site-specific 

riparian restoration project for wildlife habitat to the first-time, as well as the experienced, 

restoration project manager. The handbook can be used for planning projects, creating 

budgets, and assessing restoration success. Ecological, biological, and regulatory 

components of a riparian restoration project are described. The handbook emphasizes 

the ecological river processes operating on floodplains and in river channels that create 

characteristic vegetation structure that forms wildlife habitat—as the foundation for 

planning a riparian restoration project. Case studies of Statewide riparian restoration 

projects that faced site-specific conditions illustrate implementation of the principles 

presented in this handbook. 

Part 2—Riparian Restoration Projects: 1960 to 2000

The California Wildlife Habitat Relationships system of classification identifies seven 

major riparian habitats in California: montane riparian, valley foothill riparian, desert 

riparian, palm oasis, freshwater emergent wetland, wet meadow, and aspen (RHJV 2004). 

The vast majority of riparian restoration projects in the last half of the 20th century were 

conducted in the valley foothill riparian habitat type. This is because human impacts 

were the greatest in the Central Valley, in the lower foothills of the Cascades and Sierra 

Nevada, and in the Coast Ranges. Therefore, most of the riparian habitat restoration 

projects discussed below pertain to the valley foothill riparian habitat type.

In 1990, the Riparian Revegetation Study Group conducted a survey1320of riparian 

restoration projects in California. Information was collected for more than 276 riparian 

restoration projects in California, of which 226 projects had been implemented or were in 

the process of being implemented. Fifty-nine percent of the projects were less than 5 acres 

in size, 72 percent of the projects were less than 10 acres, and 83 percent of the projects 

were less than 20 acres. The primary purposes for which these riparian restoration projects 

were conducted were: streambank stabilization (40 percent); mitigation for project 

impacts (34 percent); fish/wildlife conservation (13 percent); urban stream restoration (8 

percent); and restoration of degraded public lands (5 percent). 

13 Survey conducted by John Stanley on behalf of the RRSG. Survey records and data were not published.
168 USDA Forest Service RMRS-GTR-411. 2020.

We have divided the projects discussed in this section into the following categories, 

which differ only slightly from the “purposes” mentioned above:

• Bank stabilization/erosion control

• River/stream/watershed restoration

• Urban creeks restoration

• Compensatory mitigation

• Habitat conservation/restoration

• Rangeland restoration

• Invasive plant removal/control

We recognize that many of these early projects were multi-purpose. Assignment of 

restoration projects to these categories is based on the primary reason for the restoration 

project. Refer to table 7 for additional information on the restoration projects mentioned in 

the text, including their general location (city/county), project sponsor(s), and the various 

actions involved in project implementation. Within each category in table 7, the projects 

are organized in order of the year(s) when they were implemented. Refer to table 8 for the 

scientific names of plant species mentioned in the text and in table 7. 

We have placed our emphasis on “early” restoration projects—projects that were 

implemented in the 1970s and 1980s. However, in doing so we have not done justice to 

many of the projects implemented in the 1990s—many of which are larger projects that 

provided greater benefits to wildlife and the environment. We set an arbitrary cut-off date 

of the end of Year 1999 by which a project had to have begun construction/installation to be 

included in the text and table 7. This means that many significant projects that were planned 

in the 1980s and 1990s but which were not constructed until Year 2000 or thereafter are not 

mentioned in this chapter. We recognize that many of these early projects did not achieve 

all, or even most, of the ecological attributes of restored ecosystems as outlined in the SER 

Primer (SER 2004) and expanded upon in Clewell and Aronson (2013).

Additionally, we have focused on “Horticultural Restoration” (formerly referred to as 

“revegetation,” “active revegetation,” or “active restoration”) and have given much less 

attention to “Process Restoration” (formerly referred to as “passive restoration”). Griggs 

(2009) stated: “‘Horticultural Restoration’ refers to a high level of site management and 

external human inputs that include site preparation (land-leveling, disking), planting of 

nursery-grown trees and shrubs in predesigned patterns, irrigation, and chemical [and/

or mechanical] weed-control for three or more years. Horticultural restoration is typically 

employed along rivers where the river’s physical processes have been severely modified by 

humans with dams, levees, bank stabilization, and water diversions.” 

Griggs (2009) also specified: “‘Process Restoration’ strives to reestablish river processes 

onto the site. Process Restoration is appropriate on riparian sites along a river that retains 

functioning river processes (e.g., no dams, and few levees or water diversions). Process 

Restoration attempts to restore a site by working with existing river processes.” “Process 

Restoration” has also been referred to as “Process-based Restoration.” A Process-based 

Restoration approach might simply be the removal of a perturbation (for example, livestock) 

from a riparian corridor or a change in management (for example, grazing regime) followed 

by allowing native plants to reestablish on the site through natural processes. Process-based 

restoration projects often involve a single action (for example, removing livestock, breaching 

of a levee to reconnect the river to its floodplain, conducting a controlled burn, removal of 

invasive plants) followed by sitting back and letting natural regeneration occur. 
USDA Forest Service RMRS-GTR-411. 2020. 169

Project name Date(s) Location

Project 

sponsor Actions taken to implement project

Bank stabilization/erosion control projects

Monument Bend 

Demonstration Project 

1967 Sacramento 

County

U.S. Army 

Corps of 

Engineers

Planted a variety of trees and shrubs along 3 miles of riverbank 

where the levee had been set back and the new berm 

protected by rock.

Bull Creek Bank 

Stabilization Project

1971-

974

Humboldt 

County

CA Dept. 

Parks and 

Recreation

Planted willow cuttings and alders in riprapped banks along 

Bull Creek in Humboldt Redwoods State Park.

Intertidal Zone Levee 

Experimental Planting 

(UC Davis Environmental 

Horticulture Department)

1978-

1980

San Joaquin 

County

CA Dept. 

Water 

Resources

In 1978, UC Davis researchers planted intertidal zone sites at 

the Webb Tract and Mandeville Island with buttonbush, three 

species of spikerush, Goodding's willow, and common tule. 

In 1980, planted intertidal zone on Terminous Tract with tules 

and willow cuttings and spikerush along with erosion control 

fabric. Seeded with alkalai bulrush and water grass.

Lost Canyon 

Rehabilitation

1983-

1985

Near 

Wishon 

Reservoir, 

Fresno 

County

Pacific Gas 

& Electric 

Company

Restored eroded watercourse downstream from PG&E Helms 

Pumped Storage Project. Planted montane riparian vegetation 

along 3 miles of eroded streamside (90 acres) at elevations 

between 6,300 and 7,700 ft. Planted over 50,000 seedlings 

over a 2-year period. Riparian species planted included blue 

elderberry, mountain alder, mountain ash, black cottonwood, 

quaking aspen, and willow. Compared various planting 

treatments.

McDonald Creek 

Restoration Project 

1983 & 

1986

Humboldt 

County

Redwood 

Community 

Action Agency

Planted approx. 4,000 trees (alder, willow, spruce, redwood) 

along 1 mile of McDonald Creek and 1⁄2 mile along the north 

fork (total of 9 acres).

Tryon Creek Restoration 

Project

1984-

1987

Del Norte 

County

Redwood 

Community 

Action Agency

Planted riparian trees and upland conifers on private property 

along one-mile reach of Tryon Creek, tributary to the Smith 

River. Fenced streambanks to exclude livestock.

Prairie Creek Restoration 

Project

1986 Humboldt 

County

Redwood 

Community 

Action Agency

Planted 8 acres of streamside deciduous forest (primarily 

willows and red alders) and adjacent conifer forest (Sitka 

spruce and coast redwood) along approximately 2 miles of 

eroding streambank on Prairie Creek (tributary of Redwood 

Creek). Installed fencing to exclude cattle. Installed log and 

hog wire bank stabilization structure on some of laid-back 

banks.

Red Clover Creek Erosion 

Control Demonstration 

Project

1986-

1987

Plumas 

County

Plumas 

Corporation

Revegetation of streambanks along 1 mile of Red Clover Creek 

(elevation 5,497 ft.), tributary to East Branch of the North Fork 

of the Feather River, with montane riparian vegetation. Installed 

check dams to reduce channel gradient and downcutting of 

streambed. Studied survival rates of four hardwood species 

(coyote willow, mountain alder, black cottonwood, quaking 

aspen) native to Red Clover Valley. Compared survival of 

unrooted stakes vs. rooted liners planted in fall vs. spring. 

Installed 3 miles of exclosure fencing to control cattle access.

Georgiana Slough Pilot 

Bank Protection Project

1999 Near 

Walnut 

Grove, 

Sacramento 

County

California 

Bay-Delta 

Authority

Installed biotechnical bank protection measures (brush boxes, 

brush bundles, coir biologs) along 7,000 feet of Georgiana 

Slough. Established 8,000 feet of tules along Georgiana Slough 

and North Fork Mokelumne River.

Table 7—Examples of California riparian ecosystem restoration projects: 1967-2000.
170 USDA Forest Service RMRS-GTR-411. 2020.

Project name Date(s) Location

Project 

sponsor Actions taken to implement project

River/stream/watershed restoration projects

Lagunitas Creek 

Watershed Restoration 

Program

1982-

1988

Marin 

County

Marin County 

Resource 

Conservation 

District

There were 70 individual work sites in this project. Activities 

included gully repair, installation of check dams, post and 

wire fencing, gabion and riprap placement, and willow sprig 

planting. Treated 21 miles of unpaved roads to control erosion 

(regrading and installation of culverts, fords, and waterbars).

San Simeon State Beach 

Riparian Restoration 

Program

1985-

1986

San Luis 

Obispo 

County

California 

Dept. of 

Parks and 

Recreation 

and California 

Coastal 

Commission

Removal of nonnative trees and shrubs along San Simeon 

Creek followed by installation of riparian trees and shrubs 

along both the north and south banks of the creek. Plant 

species installed included California sycamores, Fremont 

cottonwoods, California bay, coyote bush (sic), coast live oaks, 

willows, breadless (sic) wild-rye, and purple needle grass.

Walker Creek Watershed 

Restoration Program

1986-

1990

Tomales, 

Marin 

County

Marin County 

Resource 

Conservation 

District

Rangeland erosion control and gully repair. Streambank repair 

including riprapping toes of banks, constructing of log crib 

wall interplanted with willows, and erecting gradient control 

structures. Rehabilitation of dirt road system.

Uvas Creek Park Preserve 

Restoration Project

1995 City of 

Gilroy, 

Santa Clara 

County

City of Gilroy 

Dept. of 

Parks and 

Recreation

Reconstructed sinuous, meandering channel on 0.6-mile reach 

of Uvas Creek within Uvas Creek Park Preserve in November 

1995. In-stream channel improvements included rock vortex 

weirs (placement not as designed) and log and rootwad bank 

stabilization. Most improvements washed out in February 1996 

resulting in an irregular, braided sand and gravel channel and 

eroding streambanks. Riparian plantings on channel banks did 

not have time to become established, or were not installed, 

prior to bank erosion.

Cold Creek/Pioneer 

Trail Stream Habitat 

Restoration

circa 

1998

City of 

South Lake 

Tahoe, El 

Dorado 

County

California 

Tahoe 

Conservancy

Stabilization of eroding streambanks using boulders, logs and 

rootwads. Woody material and boulders were placed in the 

stream so as to improve fishery habitat. (Lake Tahoe Basin - 

Trout Creek Watershed)

Trout Creek Wildlife 

Enhancement and Stream 

Restoration Project

1999-

2001

City of 

South Lake 

Tahoe, El 

Dorado 

County

City of South 

Lake Tahoe 

and California 

Tahoe 

Conservancy 

Relocated Trout Creek to its historic position in middle of Trout 

Creek Meadow. Constructed approx. 3.5 miles of new sinuous 

stream channel and restored 107 acres of meadow. Creek was 

reengineered to reestablish hydrologic connectivity between 

the stream and its former floodplain thereby increasing flood 

frequency and duration, raising groundwater levels, and 

improving health of wetland and riparian vegetation. Used 

stacked sod revetments to stabilize streambanks along with 

planting of willow sprigs. 

Bear Creek Meadow 

Restoration Project

1999-

2000

Shasta 

County

California 

Dept. of Fish 

and Game

Restored historic floodplain and channel connection in Bear 

Creek Meadow. Constructed and revegetated 2.2 miles of 

channel to mimic pre-disturbance conditions. Filled incised 

gully to floodplain elevations. Propagated and planted 4,500 

shrubs from 20 native species. Used sod mat transplants to 

stabilize new exposed banks. Transplanted large willows, rose, 

hawthorne, chokecherry, and spirea. Planted over 109,000 

plugs of native sedge and rush species. 

Table 7—Continued.
USDA Forest Service RMRS-GTR-411. 2020. 171

Project name Date(s) Location

Project 

sponsor Actions taken to implement project

Urban creeks restoration projects

Strawberry Creek Park 1983 City of 

Berkeley, 

Alameda 

County

City of 

Berkeley 

Public Works 

Department

Daylighted 200-foot long section of Strawberry Creek by 

removing concrete culvert (20 feet below grade) in place since 

1904. Revegetated with coastal riparian plant species along 

1,200-foot long reach of reconstructed streambanks stabilized 

with concrete recycled from removal of cross-street.

Miller Creek Restoration 

Project

1985-

1986

Marin 

County

Deerfield Park/

Lucas Valley 

Site Developer

Regraded deeply incised stream channel to create a 

compound channel with a low flow channel and an overflow 

terrace. Installed vegetated riprap and vegetated spur dikes to 

protect meander bends. Stabilized toes of eroding banks with 

rocks and willow cuttings. Planted willows and a mixture of 

native trees and shrubs and hydroseeded with native perennial 

grasses.

Wildcat-San Pablo Creeks 

Flood Control Project

1986-

1988

City of 

Richmond, 

Contra 

Costa 

County

USACE and 

Contra Costa 

County Flood 

Control 

District

Restoration of a channelized stream. Constructed “naturallike,” “two stage”, flood control channel with a meandering 

bankfull channel and floodplain. Installed riparian plantings 

along both sides of low-flow channel. 

Carmel River Biotechnical 

River Restoration

1986-

1988

Monterey 

County

Monterey 

Peninsula 

Water 

Management 

District

Regraded pilot channel flanked by second-stage floodplain 

at level inundated by approx. 2-year flood. Planted 25-acre 

Schulte Road Restoration Project site with a series of willow 

rows at toe of eroding streambanks and willow groins in 

herringbone pattern on floodplain to narrow excessively wide 

sections of channel and stabilize river meanders. Also, installed 

biotechnical revegetated riprap and post and wire revetment 

with revegetation on outside bends. Installed some Fremont 

cottonwoods translocated from Central Valley whereas Carmel 

Valley had only a native black cottonwood forest.

Strawberry Creek 

Restoration Project

1987-

1988

City of 

Berkeley, 

Alameda 

County

University of 

California, 

Berkeley

Repaired old check-dams and installed new ones on section of 

Strawberry Creek on UC Berkeley Campus. Stabilized eroding 

streambank using redwood log crib wall. Crib wall was 

backfilled and planted with about 25 species of native plants.

Strawberry Creek 

Management Plan/

Program

1987+ City of 

Berkeley, 

Alameda 

County

University of 

California, 

Berkeley

UC Berkeley Campus watershed management strategies 

consisting of erosion control, non-point source pollution 

mitigation, and stormwater management. Gully control and 

repair through the use of biotechnical and soil bioengineering 

methods utilizing native vegetation and indigenous materials.

First San Diego River 

Improvement Project

1987-

1989

City of San 

Diego, 

San Diego 

County

City of San 

Diego

Revegetation of 26.8 acres of riparian woodland along a 

7,000-foot section of the San Diego River in Mission Valley. 

Planted willows, Fremont cottonwoods, sycamores, live oaks 

and other appropriate riparian corridor species.

Buena Vista Creek 

Restoration Project

1988 City of Vista, 

San Diego 

County

City of Vista 

and California 

State Coastal 

Conservancy

Installation of series of drop structures to slow water flow. 

Riparian habitat restoration along Buena Vista Creek to reduce 

sedimentation in Buena Vista Lagoon. 

Table 7—Continued.
172 USDA Forest Service RMRS-GTR-411. 2020.

Project name Date(s) Location

Project 

sponsor Actions taken to implement project

Compensatory mitigation projects

Sacramento River 

Maintenance Area 9

1981 Sacramento 

County

CA Dept. 

of Water 

Resources

Installed 1,142 native trees and shrubs on the upper banks 

of a 9.5-mile section of rock-reinforced levee located just 

downstream of Sacramento. Planted white alder, coyote brush, 

Oregon ash, western sycamore, live oak and valley oak. Plants 

were planted in clusters of 3-5 individuals, 100 feet apart in 

accordance with Reclamation Board standards. Monitored 

plantings through 1983.

Guadalupe River 

Revegetation Project

1981 Santa Clara 

County

Santa Clara 

Valley Water 

District

Planted California native plant species on bank slope along 

channelized section of Guadalupe River.

I-8/I-15 Mitigation 1982 San Diego 

County

Caltrans Lowered upland adjacent to San Diego River in Mission Valley 

10-14 feet in elevation to create floodplain planting bench. 

Planted 6 acres of willow/cottonwood riparian woodland with 

sycamores.

Caldecott Park Creek 

Revegetation

1983-

1985

Alameda 

County

Alameda 

County Flood 

Control 

District

Installed riparian plant species along Caldecott Creek within 

Caldecott Park. Planted alder, toyon, big-leaf maple, coast live 

oak, and bay laurel.

Alamitos Creek 

Revegetation Project

1984 Santa Clara 

County

Santa Clara 

Valley Water 

District

Planted riparian trees, shrubs and groundcover along three 

miles of Alamitos Creek totaling approx. 20 acres. 

Llagas Creek Watershed 

Mitigation

1984-

1987

Santa Clara 

County

USDA SCS 

and Santa 

Clara Valley 

Water District

Revegetation with riparian trees, shrubs and herbaceous plants 

along 10-mile creek corridor.

Morena Street Site 1985 & 

1988

San Diego 

County

Caltrans Planted willow, sycamore, and cottonwood groves (3.5 acres) 

and Coastal Sage Scrub and Mixed Chaparral communities on 

9-acre site along the San Diego River.

Crescent Bypass Riparian 

Revegetation Project

1985-

1988

King County Kings River 

Conservation 

District

Planted riparian vegetation along 6 miles of the Crescent 

Bypass between the south and north forks of the Kings River.

Saratoga Creek Flood 

Control Project 

Revegetation

1986 & 

1988

Santa Clara 

County

Santa Clara 

Valley Water 

District

Installed 3,000 native riparian plant species in planting pockets 

and open bottom concrete planters within 7 acres of gabionlined (stacked and mattress) flood control channel and in bare 

earth at top of bank.

Sweetwater Bridge 

Mitigation

1986-

1987

San Diego 

County

Caltrans Lowered 2-acre site alongside the Sweetwater River down 5-8 

feet. Planted willow scrub woodland to replicate the habitat of 

least Bell’s vireo.

Lower Coyote Creek Pilot 

Revegetation Project

1986-

1987

City of 

San Jose, 

Santa Clara 

County

Santa Clara 

Valley Water 

District

Planted total of 3,640 plants comprised of 15 native riparian 

corridor plant species on 4.4-acre site on floodplain adjacent 

to lower Coyote Creek. Woody plant species included 

California box elder, white alder, Oregon ash, California 

black walnut, western sycamore, Fremont cottonwood, 

coast live oak, valley oak, red willow, yellow willow, blue 

elderberry, and California bay. Multiple types of plant materials 

(i.e., propagule types) were tested for each plant species. 

Comparison of approx. equal areas irrigated by overhead 

irrigation versus flood irrigation. Tested a variety of weed 

management strategies and techniques.

Table 7—Continued.
USDA Forest Service RMRS-GTR-411. 2020. 173

Project name Date(s) Location

Project 

sponsor Actions taken to implement project

M&T Ranch Elderberry 

Mitigation

1987 Butte 

County

CA Dept. 

of Water 

Resources

Scattered plantings of blue elderberry on 167-acre parcel as 

mitigation for endangered Valley Elderberry Longhorn Beetle 

(VELB) habitat loss associated with Sacramento River Bank 

Protection Project.

Novato Creek Flood 

Control Project

1987-

1988

Marin 

County

Marin County 

Flood Control 

District

Planted riparian trees and shrubs on approx. 7 acres of riparian 

corridor along Novato Creek. Plants included California 

buckeye coyote bush (sic), black walnut, Oregon ash, red 

alder, coast live oak, valley oak, bay laurel, and elderberry.

Spring Creek Flood 

Control Project

1987-

1988

Sonoma 

County

Sonoma 

County Water 

Agency

Off-site revegetation of approx. 12 acres of riparian habitat 

along the Laguna de Santa Rosa. Primary plantings were oaks 

in the upper regions and Oregon ash in wetter areas.

Sacramento River Mile 

154.6 Right

1987-

1988

Colusa 

County

U.S. Army 

Corps of 

Engineers

Planting of blue elderberry and other high terrace riparian 

species at two sites totaling 2 acres as mitigation for loss of 

VELB habitat due to installation of rock slope protection.

San Joaquin Marsh 

Mitigation Bank

1987-

1988

Irvine, 

Orange 

County

The Irvine 

Company

Planted 8.5 acres of willows, cottonwood and sycamore 

adjacent to San Diego Creek in Irvine out of ultimate total of 

approximately 30 acres.

SR-52 Mitigation/Mission 

Trail Park

1989-

1990

San Diego 

County

Caltrans Lowered 33 acres on 48-acre parcel adjacent to the San 

Diego River 10-17 feet. Planting of riparian species followed 

prescription for endangered least Bell’s vireo habitat generated 

by Baird and Rieger (1989). Upland slopes created by the 

grading were seeded with coastal sage scrub which provided 

habitat for the threatened California gnatcatcher.

Hwy 85 Mitigation 1993 San Jose, 

Santa Clara 

County

Caltrans Lowered 24-acre off-site mitigation area adjacent to middle 

Coyote Creek 10-15 feet in order to bring the final grade 

closer to the groundwater table. Constructed side channel 

to convey stream flows through project site. Revegetation 

included streamside, floodplain and valley oak riparian forest 

associations. A total of 10,484 container plants were installed.

Habitat conservation/restoration projects

Colorado River Dredge 

Spoil Revegetation

1979-

1980

Near Palo 

Verde, 

Imperial 

County

USDI 

Bureau of 

Reclamation

Planted cottonwoods, willows, and quail bush on three sites 

located along the lower Colorado River including one 49-acre 

site on the Cibola National Wildlife Refuge. Studied effects 

of deep tillage and irrigation on plant growth and survival of 

rooted cuttings of Fremont cottonwoods on total of 79 acres of 

dredge-spoil. 

Sepulveda Wildlife 

Reserve Revegetation 

Project

1984 & 

1986

City of Van 

Nuys, Los 

Angeles 

County

US Army 

Corps of 

Engineers

Planted 17 acres of riparian woodland within the Sepulveda 

Flood Control Basin adjacent to the Los Angeles River. 

Installed 1,544 plants in 1984. Installed plants and cuttings in 

three 1-acre test plots in 1986 to compare irrigation methods 

(overhead vs. hand watering) with no irrigation. Plantings 

included box-elder, white alder, velvet ash, western sycamore, 

Fremont cottonwood, black cottonwood, coast live oak, 

Engelmann oak, valley oak, arroyo willow, and California bay.

Kern River Preserve 

Yellow-billed Cuckoo 

Habitat Enhancement

1986-

1989

Near 

Weldon, 

Kern County

The Nature 

Conservancy

In 1986, Bertin Anderson planted 25-acre pilot project at TNC’s 

Kern River Preserve situated along the south fork of the Kern 

River. Between 1986 and 1989, Bertin Anderson supervised 

the planting of 142 acres of cottonwoods and willows (four 

species) at the Kern River Preserve. 

Table 7—Continued.
174 USDA Forest Service RMRS-GTR-411. 2020.

Project name Date(s) Location

Project 

sponsor Actions taken to implement project

Cosumnes River Preserve 

Riparian Restoration

1988-

2000s

Sacramento 

County

The Nature 

Conservancy 

and Partners

Planted 10 acres of valley oak riparian forest on fallow 

agricultural land adjacent to the Cosumnes River. New 

technology—collar and screen developed by Frank Chan 

of PG&E—was used to protect acorns and seedlings from 

predation by rodents and grasshoppers. First use of dripirrigation by TNC. Documented growth of seedlings with 

irrigation.

Kopta Slough Preserve 

Riparian Restoration

1989-

1995

Tehama 

County

The Nature 

Conservancy

Planted approx. 300 acres. Planted Fremont cottonwood, four 

willows (red, sandbar, arroyo, black), box-elder, Oregon ash, 

California wild rose, blackberry, coyote brush. First large-scale 

native grass plantings – blue rye, creeping rye, meadow barley, 

purple needlegrass. Soil moisture was studied as to depth to 

water-table and root growth rates into its surface, as revealed 

by backhoe pits. Irrigation movement through the soil was 

monitored by electrical moisture probes.

Stony Creek Preserve 

Riparian Restoration

1991 Glenn 

County

The Nature 

Conservancy 

and CA Dept. 

of Water 

Resources

Planted approx. 500 acres. Conducted more experiments with 

the timing of weed management and the timing of irrigation on 

plant growth. Refinement of implementation monitoring.

Sacramento River 

National Wildlife Refuge 

(Llano Seco Unit) 

Riparian Restoration

1991 Butte 

County

U.S. Fish 

and Wildlife 

Service and 

partners

Species planted included all of those mentioned for Kopta 

(above) plus additional understory species, such as mugwort, 

gumplant and evening primrose.

Rangeland restoration projects

Willow Creek Restoration 1980 Near Adin, 

Modoc 

County

USDA Soil 

Conservation 

Service - (Now 

NRCS)

Dumped large rocks into slots cut across the eroded stream 

channel and keyed into the channel banks and bottom to 

create a series of rock sills across the channel. Disturbed 

channel banks were planted with willows and other woody 

shrubs. Area was fenced to prevent indiscriminate use by 

livestock.

Clark Canyon Riparian 

Demonstration Area

1984-

1987

Mono 

County

U.S. Bureau 

of Land 

Management

Constructed multiple check dams along one mile of Clark 

Canyon Creek (tributary to Aurora Creek), East Walker River 

sub-basin, to control gully head-cutting, trap sediment, raise 

water table, and restore meadow riparian areas. Elevation 

7,000 to 7,300 ft.

Invasive plant removal/control

Thousand Palms Canyon 

Tamarisk Control Project

1986-

1992

Riverside 

County

The Nature 

Conservancy 

and Partners

Volunteers and California Conservation Corps crews removed 

tamarisk in Thousand Palms Canyon—a 25-acre, 1-mile long 

wetland in the center of Coachella Valley Preserve. Infestation 

of tamarisk threatened native riparian community of desert 

fan palms, coyote willow, Fremont cottonwoods, common 

reed, honey mesquite, and screwbean mesquite. Outplanted 

mesquite grown in on-site nursery. Spread seeds collected 

on-site from natives including palm, cottonwood, mesquite, 

saltbush, quailbush, and alkali goldenbush.

Table 7—Continued.
USDA Forest Service RMRS-GTR-411. 2020. 175

Project name Date(s) Location

Project 

sponsor Actions taken to implement project

Afton Canyon Riparian 

Restoration Project

1992-

1996

San 

Bernardino 

County

Barstow 

Resource Area 

Office, Bureau 

of Land 

Management

300-acre pilot project within the Afton Canyon Area of Critical 

Environmental Concern (ACEC) on the Mojave River. Actions 

included construction of barriers to OHV travel; installation 

of cattle allotment exclusion fencing; use of prescribed fire 

in dense saltcedar stands; herbicide application to saltcedar 

resprouts; manual saltcedar stem cutting and herbicide 

application; revegetation of saltcedar removal areas using 

natural revegetation; pole planting of cottonwood and willow 

trees (7,000+); and seeding of native shrubs and grasses. 

Russian River Watershed 

Giant Reed Eradication 

Program

1992+ Mendocino 

and 

Sonoma 

Counties

Circuit Rider 

Productions 

and Partners

Conducted research on control methods, including non-toxic 

approaches to giant reed removal. Conducted digital mapping 

of giant reed locations in riparian zones within the watershed. 

Prioritized sites for giant reed removal and follow-up habitat 

restoration. Educated landowners and community about values 

of riparian zones and problems associated with giant reed. 

Coordinated volunteer and community involvement in giant 

reed removal and habitat restoration. Conducted long-term 

monitoring.

Santa Ana River 

Watershed Arundo 

Habitat Management 

Program

1997+ Orange 

County and 

Riverside 

County

Santa Ana 

Watershed 

Project 

Authority 

(SAWPA) and 

Partners

Since 1997, SAWPA and its Partners (primarily the Santa Ana 

Watershed Association and the Riverside County Regional Park 

and Open-Space District) have removed over 3,000 acres (out 

of approximately 10,000 acres) of Arundo from the Santa Ana 

River Watershed. Cleared areas have been replaced with native 

riparian or wetland vegetation.

Table 7—Continued.

Table 8—Common and scientific names for plant species.

Common name Scientific name

Alkalai bulrush Scirpus robustus

Alkali goldenbush Haplopappus acradenius

Arroyo willow Salix lasiolepis

Arundo (see Giant reed) Arundo donax

Athel Tamarix aphylla

Beardless wildrye Leymus triticoides

Big-leaf maple Acer macrophyllum

Blackberry Rubus ursinus

Black cottonwood Populus trichocarpa

Black walnut Juglans hindsii

Black willow Salix nigra (aka S. gooddingii)

Blue elderberry Sambucus caerulea

Blue elderberry Sambucus mexicana

Blue rye Elymus glaucus

Box elder Acer negundo

Buttonbush Cephalanthus occidentalis

California bay or bay laurel Umbellularia californica

California blackberry Rubus ursinus

California buckeye Aesculus californica

California wildrose Rosa californica

Chokecherry Amelanchier pumila

Coast live oak Quercus agrifolia
176 USDA Forest Service RMRS-GTR-411. 2020.

Table 8—Continued.

Common name Scientific name

Coast redwood Sequoia sempervirens

Common reed Phragmites australis

Common tule Scirpus acutus

Coyote brush Baccharis pilularis

Coyote willow Salix exigua

Creeping rye Leymus triticoides

Engelmann oak Quercus engelmannii

Evening primrose Oenothera elata ssp. Hirsutissima

False bamboo (Arundo) Arundo donax

Fan palm Washingtonia filifera

Fremont cottonwood Populus fremontii

Giant reed Arundo donax

Goodding's willow Salix gooddingii

Gumplant Grindelia camporum

Honey mesquite Prosopis glandulosa

Hawthorne (hawthorn) Crataegus douglassi

Meadow barley Hordeum brachyantherum

Mountain alder Alnus tenuifolia

Mountain ash Sorbus scopulina

Mugwort Artemesia douglasiana

Oregon ash Fraxinus latifolia

Palo verde (paloverde) Parkinsonia L.

Purple needlegrass Stipa pulchra

Quailbush Atriplex lentiformis

Quaking aspen Populus tremuloides

Red alder Alnus rubra

Red willow Salix laevigata

Saltbush Atriplex polycarpa

Saltcedar (tamarisk) Tamarix ramosissima (aka T. pentandra)

Sandbar willow Salix interior; aka S. exigua

Screwbean mesquite Proposis pubescens

Sitka spruce Picea sitchensis

Spikerush Eleocharis spp.

Spirea Spiraea sp.

Tamarisk Tamarix ramosissima (aka T. pentandra)

Toyon Heteromeles arbutifolia

Tule Scirpus spp.

Valley oak Quercus lobata

Velvet ash Fraxinus velutina

Water grass Echinochloa crusgalli

Western sycamore Platanus racemosa

White alder Alnus rhombifolia

White root Carex barbarae

Willow Salix spp.

Yellow willow Salix lasiandra
USDA Forest Service RMRS-GTR-411. 2020. 177

Rieger et al. (2014) refer to these two different approaches to restoration as 

“Construction and Installation Strategies” (horticultural/active) versus “Management 

Strategies” (process-based/passive). Actually, many riparian restoration projects employ both 

strategies on the same site at the same or different times.

Most of the early restoration projects employed horticultural restoration because the 

project planners were dealing with: (1) land surfaces that were well above active flooding; 

(2) lowered groundwater tables; (3) altered stream hydrology due to construction and 

operation of dams in the watershed and an increase in impervious cover in urban watersheds; 

(4) the presence of certain invasive plants that colonize a site so quickly so as to preclude 

the natural reestablishment of native plant species; and (5) pressure from clients and/or 

regulators to get riparian vegetation established quickly.

Notwithstanding our emphasis on horticultural restoration, it needs to be recognized 

that some restoration projects have operated under the principle that “revegetation 

projects can sometimes be best accomplished by planting nothing,” (Riley 1998). Riley 

(1998) went on to state, “The best revegetation project from a standpoint of ecological 

diversity, and the most economical, may be the project that simply creates the conditions 

needed for native vegetation to ‘reinvade’ a site.” Unfortunately, the presence of invasive 

nonnative plants on, or near, most riparian restoration project sites makes this a risky 

proposition for most project funders and sponsors.

Many of the early riparian restoration projects were not given formal names; in 

other instances, the authors of papers did not provide the project name but rather only 

the project location. We have assigned names to these projects in order to facilitate their 

discussion. Also, we have used the same terms (for example, revegetation) used by the 

authors in the literature describing the project.

Bank Stabilization/Erosion Control Projects

In 1960, at the request of the State of California, Congress authorized the Sacramento 

River Bank Protection Project to protect the Sacramento Valley levee system (Kindel 

1977). Initially, bank protection was provided at the most critical areas (in other words, 

areas where erosion had extended well into the levees). Jannssen (1976) recalled: “In 

order to rebuild the levee and construct the rock protection (riprap), trees, and vegetation 

growing on the levee were removed. Public concern over this practice led to attempts 

by the U.S. Army Corps of Engineers (CE) and the state to establish experimental test 

planting areas ... to determine if selected vegetative species could be found that would 

not constitute a threat to the structural integrity of the levees.” These test plantings were 

considered unsuccessful (Jannssen 1976).

As mentioned in part 1 above, DWR conducted a Pilot Levee Maintenance Study 

(CDWR 1966; CDWR 1967) between 1962 and 1967. Field testing for the Pilot Levee 

Maintenance Study was conducted within the Sacramento-San Joaquin Delta primarily on 

levees bordering the Sacramento River. The five test sites selected were at Garcia Bend, 

Ryde, Steamboat Slough, Hood, and Isleton. Much of the work done at the test sites was 

conducted by other governmental agencies under contract to CDWR. Between 1963 and 

1965, tests were conducted under three categories: plant performance and maintenance; 

levee protection and repair; and revetment with vegetation. Specific types of experiments 

were conducted under each category. Most of the plant species used in these tests were 

nonnative groundcovers and trees, with the exception of tests involving the management 
178 USDA Forest Service RMRS-GTR-411. 2020.

of existing vegetation. The experiments conducted in this study led to the general 

conclusion that “alternative levee maintenance practices can be used to allow vegetation 

on levees” and that “this vegetation can be maintained for the multiple use of levees 

without jeopardizing the primary function of flood control” (CDWR 1967).

In the late 1960s, the Corps planted trees and shrubs at several selected sites along the 

Sacramento River to “demonstrate that such vegetation can be successfully grown, can be 

compatible with flood control requirements, and can offer a significant improvement to 

aesthetics and other environmental aspects of the river” (Kindel 1977). The most noteworthy 

project was the CE Monument Bend Demonstration Project in 1967 (Kindel 1977). 

Most early streambank stabilization projects relied heavily on structural measures 

for controlling bank erosion. Riparian vegetation was planted within these structures to 

soften the visibility of concrete or to take over the function of slowing water adjacent 

to the banks initially provided by the revetment. The California Department of Parks 

and Recreation conducted some of the earliest projects; for example, Bull Creek Bank 

Stabilization in 1971-1974 (Barry 1984; Barry 1985; Barry 1988).

Between 1978 and 1980, researchers from the Environmental Horticulture 

Department at the University of California, Davis (Whitlow et al. 1984) investigated the 

potential for using vegetation as an agent for erosion control in the tidal zone on levees 

in the Sacramento/San Joaquin Delta. They conducted experimental plantings for the 

CDWR to identify species and planting techniques suitable for application in the intertidal 

zone14.

21The ultimate goal was a vegetative replacement for riprap. Unfortunately, both 

research sites were riprapped before the investigations could be completed.

The USDA Soil Conservation Service (SCS, now NRCS) commonly worked with 

landowners in rural areas and sometimes in urban areas to control streambank erosion. 

Patterson et al. (1984) described some of the more common streambank stabilization 

techniques used by the SCS during the 1970s and 1980s. Many of these measures 

provided “physical protection” (for example, rock riprap, post and wire revetment, gabion 

baskets), while others employed “vegetative protection” (for example, woody cuttings, 

rooted woody plants, herbaceous plants) but often in combination with some form 

physical protection at the toe and/or on the lower bank slope. This woody vegetation grew 

to cover the physical measures and provide valuable fish and wildlife habitat. 

Severe storms in 1978, 1979, 1980, and 1982 caused considerable damage to streams 

in California. Drawing on Emergency Watershed Protection (EWP) funds, the SCS 

used structural and vegetative measures to stabilize severely eroding streambanks and 

reestablish riparian vegetation. From 1978 through 1982, a total of 371 EWP projects 

was completed. Some of the river systems on which these EWP streambank stabilization/

riparian revegetation projects were constructed were the Cuyama River in Santa Barbara 

and San Luis Obispo counties, Carmel and Salinas rivers in Monterey County, Aptos 

Creek in Santa Cruz County, and the Santa Clara River in Ventura County (Gray et al. 

1984). Although most of the streambank stabilization projects were small, SCS completed 

work on over 100 miles of streambanks. In 1983, Schultze and Wilcox (1985) evaluated 

the results of the revegetation work for 29 projects in California’s central coast area. 

“Early SCS revegetation efforts used nonnative species of plants or species not considered 

riparian,” (Gray et al. 1984). Although native species (primarily willows) were most 

14 Although research sites were within the intertidal zone, they were presumably influenced by fresh and/or brackish water based on the 

plant species used in the experiments. (See table 7.)
USDA Forest Service RMRS-GTR-411. 2020. 179

commonly used in SCS bank stabilization projects by the late 1970s, unfortunately, 

during this period, SCS sometimes incorporated invasive nonnative plant species in bank 

stabilization measures. For example, woody cuttings of saltcedar and athel1522were planted 

behind pipe and wire revetment and at the toe of reshaped banks on over 2 miles of the 

Cuyama River as part of an EWP project constructed in 1979 (Gray et al. 1984). Also, 

Arundo donax (referred to as false bamboo) was planted along with willows in EWP 

projects constructed on the Carmel and Salinas rivers in Monterey County in 1978 and 

1979 (Gray et al. 1984).

The Pacific Gas and Electric Company (PG&E) developed early revegetation 

techniques for riparian areas affected by its projects. The Lost Canyon Rehabilitation 

Project in 1983-1985 (Chan and Wong 1989), implemented as mitigation for erosion and 

habitat loss along a Sierra stream caused by a pipeline rupture, required the development 

of techniques for revegetation with Sierran montane riparian species in Fresno County. 

Frank Chan devised and tested some of the earliest innovative measures for native plant 

revegetation within the riparian corridor.

During the 1980s, citizens voiced opposition to the use of totally engineered 

structures for streambank erosion control. Revegetation of streambanks with native 

vegetation became an integral part of streambank stabilization projects. These types of 

projects often occurred in the northern part of the State and were driven by a concern to 

restore salmon and steelhead habitat, especially for listed fish species. Salmonid streams 

had been severely impacted by timber harvesting, road construction, and livestock 

grazing. The goal was to reduce nonpoint source sediment that was impacting salmonid 

spawning and rearing habitat. In some cases, the primary action was the construction of 

fencing to exclude livestock from the streambanks. Some of these projects employed the 

use of vegetative and quasi-vegetative bank/slope protection techniques such as presented 

in Schiechtl (1980) and Gray and Leiser (1982). 

Several natural resources employment training programs were active in stream and 

riparian restoration in the 1980s. The Redwood Community Action Agency conducted 

a number of streambank stabilization and riparian revegetation projects along coastal 

streams in northern California in the early and mid-1980s. Reichard (1989) reported 

on restoration projects performed along McDonald Creek, Tryon Creek, and Prairie 

Creek. The Plumas Corporation performed similar streambank restoration projects in the 

northern Sierra. The Red Clover Creek Erosion Control Demonstration Project (Lindquist 

and Bowie 1989; Lindquist and Filmer 1988) involved the cooperation of multiple State 

and Federal agencies, landowners, and organizations through the use of the Coordinated 

Resource Management and Planning (CRMP) process16.

23

More recent publications on the use of biotechnical/soil bioengineering techniques 

involving the use of riparian vegetation include Gray and Sotir (1996), Hoag and Fripp 

(2002), Schiechtl and Stern (1992), and Schiechtl and Stern (1997). An example of a 

biotechnical bank stabilization approach developed as an alternative to rock revetment 

(riprap) in the Sacramento-San Joaquin Delta is the Georgiana Slough Pilot Bank 

Protection Project conducted in 1999 (Hart and Hunter 2004).

15 According to Bossard et al. 2000, athel (T. aphylla) is not considered an invasive pest under most circumstances whereas saltcedar (T. 

ramosissima) is invasive. 16 The CRMP process is a collaborative public-private project planning and implementation process that seeks to involve all interested 

parties in management and restoration decisions and in project implementation.
180 USDA Forest Service RMRS-GTR-411. 2020.

River, Stream, and Watershed Restoration Projects

A wide variety of river, stream, and watershed restoration projects were implemented 

in the 30-year period between 1970 and 2000. Many of these projects were components of 

a larger vision expressed in long-range river and watershed management plans prepared 

by Federal, State, and local agencies. The USDA Forest Service and the USDI Bureau 

of Land Management prepared management plans for Federally-designated Wild and 

Scenic Rivers while The California Resources Agency prepared management plans for 

State-designated Wild and Scenic Rivers; however, these plans dealt mostly with the 

preservation and management of the designated river section and not with the restoration 

of riparian resources. CDFG (now DFW) coordinated with local agencies to prepare 

State Protected Waterway Management Plans for the San Lorenzo River (Ricker 1979), 

Big Sur River (County of Monterey 1983), and Little Sur River (Harvey and Stanley 

Associates, Inc. 1983). Riparian vegetation management and riparian revegetation were 

components of these plans. Riparian habitat protection and management within the coastal 

zone was addressed in Local Coastal Plans prepared by counties. There were also locallydriven enhancement plans for watersheds such as the Garcia River (Mendocino County 

RCD 1992) and Huichica Creek (Napa County RCD 1993). Sometimes, it worked the 

other way around in which the success of the early restoration projects prompted the 

preparation of a management plan to address issues in the entire watershed.

In-stream restoration projects focused primarily on restoring fish spawning and 

rearing habitat; however, planting of riparian vegetation on streambanks was generally 

included in these projects so as to provide Shaded Riverine Aquatic (SRA) Cover for 

fish. Large woody debris (LWD) was often reintroduced into the stream system in the 

form of logs and rootwads to improve fish habitat and cover. The LWD was typically 

secured to the streambanks and woody cuttings were inserted into eroding bank slopes. 

Boulders were used for grade control and channel stabilization. Rock groins constructed 

to slow and redirect streamflow away from the banks were planted with woody cuttings. 

Streambank stabilization typically employed biotechnical measures incorporating live 

vegetation as well as dead woody material resulting in riparian habitat benefiting fish, 

aquatic life, and riparian dependent wildlife species. Many projects also involved the 

removal or modification of migration barriers such as log jams and culverts. The provision 

of adequate filter strips of riparian vegetation was a concern for lands managed for timber 

production, especially within USFS Streamside Management Areas.

Watershed restoration projects/programs generally focused on the installation 

and use of best management practices for the control of erosion and the prevention of 

sedimentation of streambeds as well as proper stormwater management for the control of 

pollutants (nonpoint source pollution prevention/control). Landslides, unstable slopes, and 

eroding streambanks were often stabilized with biotechnical slope stabilization measures 

incorporating the use of live and dead riparian vegetation. Local Resource Conservation 

Districts often served as project coordinators—working with citizen’s advisory 

committees and governmental agencies, preparing and submitting grant applications, 

keeping track of expenditures, administering work contracts, and coordinating volunteer 

activities. The Lagunitas Creek (tributary to Tomales Bay) Watershed Restoration 

Program (Berger 1990; Marcus et al. 1987; Marcus 1989; Witkin 1990) is an example 
USDA Forest Service RMRS-GTR-411. 2020. 181

of a program using multiple approaches to control erosion coming from numerous 

sources throughout a watershed to improve coho salmon and steelhead habitat and 

reduce sedimentation in Tomales Bay. This program included preparation of an erosion 

control handbook (Prunuske 1987) containing gully control and streambank stabilization 

measures utilizing combined vegetative and structural solutions. Copies of the handbook 

were given to participating landowners and agencies. The Walker Creek Watershed 

Restoration Project (Berger 1990; Marcus 1989; Marcus et al. 1987 ) also involved the 

repair of many erosion sites through the use of riparian vegetation often combined with 

structural stabilization measures in order to reduce sedimentation in Tomales Bay.

During the 1970s, 1980s, and 1990s, there was a growing movement toward the 

development of Cooperative Watershed Associations often called watershed councils or 

alliances. These organizations generally sought to involve all stakeholders in a watershed 

or stream reach including private landowners, local, State, and Federal agencies, resource 

users, and citizen’s groups. These associations were actively involved in the planning and 

implementation of river, stream, and watershed restoration projects.

James Barry (1985) described ecosystem restoration underway at multiple sites with the 

California State Park System for the purposes of erosion control, alien species eradication, 

and natural ecosystem enhancement. Revegetation within the riparian corridor on State Park 

lands was being performed as far back as the mid-1970s, mostly to compensate for former 

logging practices and overzealous stream clearance programs. One such project was the San 

Simeon State Beach Riparian Restoration Program (Capelli 1985).

Through the years, hydrologists and fluvial geomorphologists played an increasingly 

important role in riparian corridor restoration. Their expertise was essential in a number 

of ways, for example: restoring incised and leveed stream systems (Haltiner and Beeman 

2003), which allowed for the creation of floodplain terraces that could be planted 

with riparian vegetation; restoring flooding to floodplain riparian systems including 

revegetated sites (Swenson et al. 2003); and reconnecting stream channels with their 

historic floodplains—aka re-hanging streams in meadows (Poore 2003). They recognized 

the importance of restoring the “physical integrity” (environmental health specific to a 

particular catchment river system) of rivers and their floodplains created by a “process 

of dynamic equilibrium punctuated by natural disturbances” (Haltiner et al. 1996), 

the “natural dynamic character” (management toward a more natural flow regime) of 

river systems (Poff et al. 1997), and the “natural stability” of stream channels (Rosgen 

1996). Moreover, their emphasis on fluvial geomorphological principles caused many 

restoration practitioners to focus on a process-based approach to riparian habitat and 

stream restoration (Tomkins and Kondolf 2003). The preferred approach to restoration 

was to remove or ameliorate the effects of human interventions in the river system and 

“allow the natural processes to recreate desirable habitat” (Haltiner et al. 1996). Using 

this approach, success was redefined as the “restoration of key ecologic processes 

(physical and biological conditions) that are both resilient and evolving” (Haltiner et al. 

1996). However, not all stream/riparian restoration projects withstood the test of time. 

Sometimes, in-stream channel improvements and bank stabilization measures did not 

withstand high flows long enough for riparian vegetation to become established as was the 

case with the Uvas Creek Park Preserve Restoration Project (Kondolf et al. 2001; Rosgen 

2006; Rosgen 2008).
182 USDA Forest Service RMRS-GTR-411. 2020.

Stream and meadow restoration projects in the Tahoe Basin were conducted to restore 

habitat, but primarily to reduce the amount of sediment reaching Lake Tahoe. Projects 

such as the Cold Creek—Pioneer Trail Stream Habitat Restoration Project (Tahoe RCD 

2015), designed with the assistance of Dave Rosgen, stabilized eroding streambanks with 

logs and rootwads, not only reducing sediment loads, but also creating valuable fisheries 

habitat. Projects such as the Trout Creek Wildlife Enhancement and Stream Restoration 

Project (CTC 2015) reconnected channelized streams with their floodplains, resulting in 

periodic overbank flows depositing sediment on the meadows rather than impacting the 

clarity of Lake Tahoe. 

Similarly, the Bear Creek Meadow Restoration Project (Poore 2003) in eastern Shasta 

County restored the historic floodplain and channel connection, preventing channel 

erosion and buffering peak flow events, thereby reducing sedimentation downstream 

in Fall River. Both the Trout Creek and Bear Creek projects utilized stacked sod mat 

transplants cut from the meadows to stabilize the banks of the realigned stream channels 

in conjunction with riparian plantings.

Urban Creeks Restoration Projects

In rapidly growing California cities and counties, creek corridors, however severely 

impacted, were generally the last remaining undeveloped natural environments within 

urban and suburban development. Multiple demands were placed on these areas for 

flood control, active and passive recreation, and trail systems. Urban creeks were used as 

dumping grounds for trash and discharge of polluted waters. Additionally, urban creeks 

were subjected to the impacts of significant amounts of impervious cover within their 

watersheds, which often resulted in channel incision.

In the 1970s and 1980s, there were nationwide and regional movements for the 

improvement of urban creeks and the remnants of riparian habitat on their banks. Urban 

planners, citizen groups, etc., recognized the value of these areas as refuge from urban life 

and as buffers between conflicting land uses. At the same time, scientists, conservationists, 

environmentalists, and others recognized the importance of these streams and riparian 

areas as habitat for fish and wildlife. Nationally, these “natural” corridors were referred 

to as “greenways,” basically meaning linear open space. In California, use of the term 

greenways was not common; rather, these linear corridors were generally referred to as 

“urban creeks” or “urban creek corridors” or sometimes “parkways.”

Due to the nature of the problems impacting urban creeks, many urban creek 

restoration projects required extensive planning and multiple funding requests for various 

stages of the restoration including trash removal, sanitary engineering to resolve water 

quality issues, removal of invasive plants, erosion control, bank stabilization, in-stream 

aquatic habitat enhancement, and the installation of native plant species. Funding for 

some urban creek restoration projects became available in conjunction with other 

planning efforts: for example, flood control planning and park planning. Much of the 

work was also accomplished through the use of volunteers.

In 1984, the Urban Creeks Council was formed to present an alternative flood control 

option to the Corps of Engineers, which was planning to place a major section of Wildcat 

Creek (Alameda County) into a conventional dirt/riprap and concrete trapezoidal floodcontrol channel. The Council advocated for a multi-stage channel design with a bankfull 

channel, riparian corridor, floodplain, and berms (levees). This “natural-like” channel 
USDA Forest Service RMRS-GTR-411. 2020. 183

design for the Wildcat-San Pablo Creeks Flood Control Project (Fishbain and Williams 

1988; Haltiner et al. 1996; Meyer 1989; Riley 1989a; Riley 1989b; Riley 1998; Riley 

2003) was approved and constructed by the Corps. 

The Urban Creeks Council was instrumental in getting numerous urban stream 

restoration projects implemented in the San Francisco Bay Area, the first project being 

Strawberry Creek Park (Berger 1990; Wolfe 1988), which involved the daylighting of a 

section of Strawberry Creek that had been underground in a culvert since 1904.

Many urban creek restoration projects were implemented in conjunction with the 

development of new residential subdivisions adjacent to deeply incised stream reaches. 

One of the earliest examples of the compound (multi-stage) channel approach was 

the Miller Creek Restoration Project (Haltiner et al. 1996; Yin and Pope-Daum 2004), 

wherein a low flow channel-floodplain system allowing some of the dynamics of a natural 

channel was constructed and then revegetated.

The CDWR Urban Stream Restoration Program began in 1985. Counties, cities, and 

non-profit organizations interested in improving the conditions of their watercourses were 

encouraged to submit grant proposals for restoration projects. Many of these projects 

involved the planting of riparian vegetation and the enhancement of existing riparian habitat 

in urban areas. An example of a CDWR-assisted (in other words, partial funding) stream 

restoration project with a significant riparian revegetation component is the Carmel River 

Biotechnical River Restoration (Cummings 1992; Haltiner et al. 1996; Matthews 1990).

The California State Coastal Conservancy was also actively involved in funding 

riparian habitat restoration in coastal areas especially when stream degradation was 

contributing to sedimentation of coastal lagoons. One such example is the Buena 

Vista Creek Restoration Project (Marcus 1987; Marcus 1988; Marcus 1989) to reduce 

sedimentation in Buena Vista Lagoon.

Many urban creek restoration projects involved work on only a limited portion 

or segment (sometimes referred to as a reach) of a stream. The Strawberry Creek 

Restoration Project (Berger 1990; Charbonneau and Resh 1992; Pollak 1990) and 

Strawberry Creek Management Plan/Program (Charbonneau and Rice 1989) conducted 

on the UC Berkeley campus and in its watershed are examples of urban creek restoration 

projects that included restoration work in the mid and upper watershed thereby reducing 

impacts on downstream reaches.

Locally-sponsored flood management projects evolved as alternatives to traditional 

flood control projects. These projects involved coming up with a greenbelt floodway 

design that provided flood protection but also provided for protection and restoration 

of riparian areas and public access to these “parkways.” The First San Diego River 

Improvement Project (Burkhart 1989; City of San Diego 2001; Faber et al. 1989), 

a combination of flood control, natural area, and parkway, was initiated by private 

developers to allow commercial and residential developments in the Mission Valley to be 

approved by the City of San Diego. 

The U.S. National Park Service River and Trail Conservation Assistance (RTCA) 

Program assisted local sponsors with the planning of urban creek restoration projects 

by bringing together diverse stakeholders and helping them find consensus solutions to 

restore degraded urban creeks. The RTCA published a book titled How Greenways Work: 

A Handbook on Ecology (Labaree 1992). The RTCA provided the impetus for the writing 

of Ecology of Greenways: Design and Function of Linear Conservation Areas (Smith and 
184 USDA Forest Service RMRS-GTR-411. 2020.

Hellmund 1993), a book that bridges the gap between design and ecology. Publications 

like these were important since it was often landscape architects who were responsible for 

leading the design team for urban stream restoration projects.

Compensatory Mitigation Projects

Restoration of riparian habitat was undertaken as compensatory mitigation for 

riparian habitat loss due to unavoidable impacts resulting from infrastructure construction 

projects (for example, highway and bridge construction, flood control channel 

modifications, and utility corridors) and land development projects. These projects tended 

to be within, or adjacent to, urban areas where the stream corridors were often the only 

open space remaining. The major drivers behind these mitigation projects were NEPA and 

CEQA (both passed in 1970), the California Porter-Cologne Water Quality Control Act 

(1970), CCWA (1972), ESA (1973), CE Section 404 permitting, the California Coastal 

Act (1976), and CDFG stream alteration agreements. Often there was also a strong 

desire on the part of the community to develop public access to these “restored” stream 

corridors, sometimes leading to conflicts between regulators wanting the restored lands to 

be set aside for wildlife and park planners wanting to develop infrastructure (for example, 

pathways, lighting) for visitor use.

Some of the earliest projects involving native plant landscaping and native plant 

revegetation were installed by the Santa Clara Valley Water District (SCVWD) in the 

mid- and late-1970s. The District adopted a resolution in 1974 that set policy for the 

landscaping of District projects. In 1975, SCVWD published A Landscaping Guide to 

Native and Naturalized Plants for Santa Clara County (Stiles 1975). This guide contained 

pertinent information on characteristics and suitability of numerous plants native to the 

region, including species common to the stream systems in Santa Clara County. Dr. 

Bernie Goldner (1984) referred to projects installed in 1976 on Randol Creek, San Tomas 

Aquino Creek, Los Gatos Creek, and the Guadalupe River as “landscape projects” even 

though mostly native trees and shrubs were installed along these structurally modified 

channels. Various measures of structural bank protection were associated with these 

projects. Subsequent plantings of native vegetation installed in 1979 along Calabazas and 

Berryessa creeks were referred to as “revegetation projects” (Goldner 1984). 

The fact that flood control managers were aware of the public’s concern regarding 

the impact of flood control projects on riparian habitat at a local level is evidenced by 

the publication of Valley Riparian Forests of California: An Overview of their Biological 

Significance and Physical/Chemical Processes (Stiles 1978) by the SCVWD. In 1979, 

SCVWD adopted a new policy emphasizing mitigation of substantial adverse impacts, in 

conformance with CEQA (Goldner 1984).

There was a tendency for the early “revegetation” projects to end up appearing more 

like landscaping of structurally modified flood control channels than the creation of 

valuable fish or wildlife habitat. This was in part because there was an overreliance on 

engineered structures (for example, riprap, wire basket gabions) to prevent bank erosion 

and a resistance on the part of flood control maintenance personnel to allow the planting 

of riparian vegetation at the toe of the bank slope adjacent to the channel bottom. This 

prevented the establishment of shaded riverine aquatic (SRA) Cover.

Another reason why these “revegetation” projects often resembled landscaping is 

because many of these early projects were designed by landscape architects. Plantings 
USDA Forest Service RMRS-GTR-411. 2020. 185

were arranged for various visual purposes (for example, screening and aesthetics) 

without regard to the habitat requirements (for example, plant associations and vegetation 

structure) of riparian wildlife. Sometimes cultivars with showy appearance were installed 

instead of the native species or because of a failure to plan ahead for contract growing 

of native plant materials. Some common riparian plant species (for example, poison oak 

and stinging nettle) were almost always omitted because of undesirable characteristics 

(for example, poisonous and thorniness). Additionally, plantings were laid out in a linear 

fashion so as to be watered using drip irrigation.

Examples of “revegetation” projects installed along structurally protected flood 

control channels in the early and mid-1980s by the SCVWD include: Guadalupe River 

Revegetation Project (Goldner 1984); Alamitos Creek Revegetation Project (Berger 1990; 

Goldner 1988); Llagas Creek Watershed Mitigation (Berger 1990); and the Saratoga 

Creek Flood Control Project Revegetation (Berger 1990; Gray et al. 1984). 

Up until the mid-1980s, almost all the SCVWD riparian revegetation projects were on 

the banks or immediate top-of-bank of flood control channels. The Lower Coyote Creek 

Pilot Revegetation Project (Berger 1990; Stanley et al. 1989), installed by the SCVWD 

in 1986-1987, was undertaken to determine the best means of revegetating historic 

floodplains that had been used for agriculture. Findings from revegetation on the 4-acre 

pilot project site were used for the selection of riparian species, propagule types, and 

planting, irrigation, and maintenance techniques on the remaining 28.5 acres of mitigation 

along lower Coyote Creek.

During the 1980s, other flood management agencies in the San Francisco Bay Region 

(for example, Alameda County Flood Control District, Marin County Flood Control 

District, and Sonoma County Water Agency) were also restoring riparian habitat as 

mitigation for flood control project impacts. Examples of these early revegetation projects 

include: Caldecott Park Creek Revegetation (Berger 1990); Novato Creek Flood Control 

Project (Berger 1990); and the Spring Creek Flood Control Project (Berger 1990).

Examples of early revegetation projects conducted as mitigations in the Central 

Valley in the 1980s include: California Department of Water Resources Sacramento River 

Maintenance Area 9 installed in 1981 (King 1985); and the Crescent Bypass Riparian 

Revegetation Project (Oldham and Valentine 1989; Oldham and Valentine 1990) installed 

between 1985 and 1988.

In southern California, Caltrans implemented a number of riparian revegetation 

projects as mitigation for impacts associated with highway construction. One of the 

earliest projects was the I-8 / I-15 Mitigation (Rieger 1988) constructed in 1982. Other 

Caltrans riparian revegetation projects in San Diego County included: Morena Street 

Site (Rieger 1988) installed in 1985; Sweetwater Bridge Mitigation site (Rieger 1988) 

installed in 1986-1987; and SR-52 Mitigation/Mission Trails Park (Rieger 1992) installed 

in 1989 and 1990.

Some mitigation projects involved significant alteration of the project site topography 

in order to create a planting bench or artificial floodplain with suitable flooding frequency 

and/or depth to groundwater to support riparian vegetation. Project designers were 

sometimes forced to take this option of converting upland areas to floodplain because of 

the no-net-loss of wetlands policy of the regulatory agencies. Caltrans projects involving 

significant lowering of the surface elevation include: I-8 / I-15 Mitigation in San Diego 

County (Rieger 1988); Sweetwater Bridge Mitigation (Rieger 1988) in San Diego 
186 USDA Forest Service RMRS-GTR-411. 2020.

County; SR-52 Mitigation/Mission Trails Park (Rieger 1992) in San Diego County; and 

Hwy 85 Mitigation in Santa Clara County (National Research Council 2001). Frequently, 

the cost of the excavation and earth removal was applied to the highway budget and not 

the restoration project since the excavated material was need for fill for nearby highway 

construction. 

Some mitigation projects created significant habitat for wildlife, especially migrant 

passerine bird species. Construction projects that created gaps in the riparian corridor 

(fragmentation) were often required to agree to mitigation acreage ratios of 2:1, 3:1, 

or greater, resulting in the planting of floodplain riparian habitat much wider than the 

remnant streamside vegetation that was impacted. For example, construction of a high 

flow bypass channel for the Lower Coyote Creek Flood Control Project necessitated the 

removal of approximately 15 percent of the existing riparian trees in the project area 

creating breaks in the riparian corridor. SCVWD was required to create 32.5 acres of new 

riparian habitat on the floodplain within the project levees.

Some revegetation projects focused mostly, or solely, on the mitigation of riparian 

habitat loss for special status species, especially Federally-listed endangered species. 

Caltrans Sweetwater Bridge Mitigation (Rieger 1988) and SR-52 Mitigation/Mission 

Trails Park sites were constructed primarily to provide habitat for the endangered least 

Bell’s vireo. The planting regime at the Mission Trails Park site followed the prescription 

for least Bell’s vireo habitat generated from extensive research conducted by Baird 

and Rieger (1989). Three pairs of least Bell’s vireos nested at the Mission Trails Park 

site within a year of planting (Rieger 1992). The M&T Ranch Elderberry Mitigation 

Project (Stanley 1989) in Butte County and the Sacramento River-Mile 154.6 Right 

Project (Chainey et al. 1989) in Colusa County were installed to provide habitat for the 

endangered valley elderberry longhorn beetle.

In the latter part of the 1980s, we saw the creation of mitigation banks used by 

development projects that were unable to achieve on-site mitigation; for example, the 

San Joaquin Marsh Mitigation Bank (Stanley 1989) developed by The Irvine Company in 

Orange County in 1987-1988.

Habitat Conservation/Restoration Projects

A number of governmental agencies and conservation organizations saw an 

opportunity to restore large swaths (both in terms of length and width) of riparian habitat 

by purchasing available low-lying agricultural land within floodplains and creating 

wetland and riparian conservation areas. Much of this land was prone to periodic flooding 

and no longer profitable for farming. Typically, the objective was to create a mix of 

habitat types for a variety of bird species (waterfowl, waterbirds, and riparian dependent 

bird species). 

Beginning in 1977, Bertin Anderson and Robert Ohmart undertook experimental 

revegetation projects for the reestablishment of cottonwood/willow forest on a number 

of sites along the Lower Colorado River, for the USDI Bureau of Reclamation (Ohmart 

et al. 1977). Their 1979-1980 Colorado River Dredge Spoil Revegetation (Anderson and 

Ohmart 1985a; Anderson and Ohmart 1985b) project sites totaled 74 acres, the largest 

being a 49-acre site on the Cibola National Wildlife Refuge. Interestingly, none of these 

plantings prospered—as of 2008—due to the modified soil and hydrology everywhere 

along this reach of the Colorado River. However, Anderson demonstrated hydrological 
USDA Forest Service RMRS-GTR-411. 2020. 187

needs and soil alkalinity levels for establishment of cottonwood and red willow on 

restoration sites.

The Army Corps of Engineers (CE) began revegetation at the Sepulveda Wildlife 

Reserve in the Los Angeles Basin in 1981. The CE planted 17 acres of riparian woodland 

in 1984 as part of the Sepulveda Wildlife Reserve Revegetation Project (Parra-Szijj 1990). 

Due to low survival rates of the riparian plantings, CE installed test plots in 1986 to 

demonstrate the need for irrigation and determine the best means of watering pole cuttings 

and seedlings, especially in light of the heavy growth of weeds. 

In the early 1980s, The California Nature Conservancy (TNC) saw habitat restoration 

as a new tool for the conservation of natural areas. Successful habitat restoration would 

add acres of habitat for target wildlife species on nature preserves. Habitat restoration 

would change how preserve design would configure nature preserves (based on 

restoration potential) and allow more opportunity for process-based conservation of 

wildlife. This logic carried weight with private financial donors that supported TNC.

TNC’s Kern River Preserve, situated along the south fork of the Kern River, was 

one of the earliest locations for experimentation for restoration technology in California. 

Cottonwood cuttings were planted as early as 1982; however, after initial poor results, 

TNC quickly realized the need for organized scientific testing of restoration methods—

irrigation needs, weed control, and soil factors (texture, alkalinity, and water table depth). 

TNC hired Dr. Bertin Anderson of the Revegetation and Wildlife Management Center to 

develop the restoration technology.

TNC’s Kern River Preserve Yellow-billed Cuckoo Habitat Enhancement Project 

(Anderson and Layman 1989; Reiner and Griggs 1989; Tollefson 2003) began in 1986 

when Bertin Anderson implemented a 25-acre pilot project testing planting and irrigation 

methods for the establishment of cottonwoods and willows at the site. Anderson’s 

quantitative approach allowed for the rapid development of methods that proved to 

be effective at establishing Fremont cottonwood and red willow woodlands required 

by the targeted species, yellow-billed cuckoo (YBC). From 1986-1989, 142 acres of 

cottonwoods and willows (four species) were planted at TNC’s Kern River Preserve. 

YBC began using these stands within 2-3 years of growth of the trees. As of 2001, a 

total of over 330 acres of native riparian trees and shrubs had been planted on the higher 

floodplain surfaces (Tollefson 2003). In addition, “over 500 acres of native riparian forest 

have recovered at the Kern River Preserve through ‘passive restoration,’ by limiting or 

excluding livestock grazing in low-lying areas that had been converted to pasture through 

clearing and intensive grazing” (Tollefson 2003).

TNC became involved in the restoration of riparian habitat in the Sacramento Valley 

beginning in the late 1980s at TNC’s Cosumnes River Preserve. TNC contracted with 

Harvey and Stanley Associates, Inc. (John Stanley and Harold Appleton) to assist TNC 

(Dr. Thomas Griggs) and Ducks Unlimited with the design and layout of the initial 

riparian revegetation at Cosumnes. The first phase of the Cosumnes River Preserve 

Riparian Restoration Project (Griggs et al. 1993) involved the planting of 10 acres of 

valley oak forest on fallow agricultural land in 1988. Tom Griggs supervised the initial 

restoration work at the Cosumnes River Preserve. TNC’s Habitat Restoration Team 

directed volunteers in conducting plantings in each successive year. As of 2001, a total of 

500 acres of oaks, willows, and other trees had been planted at the preserve by volunteers 

and school children (Swenson et al. 2003).
188 USDA Forest Service RMRS-GTR-411. 2020.

In early 1985, a levee protecting the farm field adjacent to the Cosumnes River 

Preserve failed. Cottonwoods and willows rapidly colonized about 15 acres of sediment 

deposited by the river on the farmland. Although the levee was repaired, the “accidental 

forest” was well established and through time provided habitat for a variety of wildlife 

species. This farm property was acquired by TNC in 1987. The rapidly growing 

“accidental forest” inspired TNC to explore how natural flooding processes could be 

enlisted to expand the riparian corridor (Swenson et al. 2003). In the mid-1990s, TNC 

reoriented its forest restoration program at the Cosumnes River Preserve to focus on 

areas where natural regeneration could be encouraged by reestablishing natural flooding. 

The Cosumnes River has close to a natural hydrograph since there are no major dams in 

the watershed—the entire Preserve area historically flooded in El Nino years, even with 

levees. In fall 1995, TNC intentionally breached the levee (created a 50-foot gap) and cut 

a shallow channel through the floodplain thereby reopening about 200 acres of bottomland 

to natural flooding. Natural “cuttings” of willow and cottonwood became established 

on the site (Mount et al. 2003; Swenson et al. 2003). The floods of 1997 caused many 

levee breaks along the Cosumnes River. The preserve and local farmers reached an 

agreement on an “unleveeing” project and convinced the CE to fund a nonstructural flood 

management project instead of traditional levee repairs. The project involved breaching 

and abandoning 5.5 miles of levees. Construction started in the fall of 1997 with the levee 

breaches and construction of a setback levee. This added about 100 acres to the floodway 

(Swenson et al. 2003).

After initiating the riparian habitat restoration at the Cosumnes River Preserve, Tom 

Griggs then moved on to plan the riparian plantings at Kopta Slough, Stony Creek, and 

the Sacramento River National Wildlife Refuge. “Cultivated restoration became necessary 

on the Sacramento River because Shasta Dam has altered natural hydrology, changing the 

patterns and extent of natural vegetation succession. Furthermore, mid to high floodplain 

soils are prone to support weedy herbaceous vegetation ...which competes with native 

woody vegetation in natural and cultivated succession. TNC developed agricultural-style 

techniques to restore relatively large acreages of riparian vegetation in a logistically and 

financially efficient manner” (Silveira et al. 2003). 

In 1989, TNC began planting riparian vegetation at the Kopta Slough Preserve 

(Griggs 1993; Griggs 1994). Kopta Slough Preserve was the R&D center for TNC for 

riparian restoration along the Sacramento River. The goal was to use all riparian plant 

species that are characteristic of riparian habitat and to study their survival and growth 

relative to a variety of factors: soil moisture, depth to water table, and irrigation water 

movement through the soil. Root growth rates and root architecture were studied by 

digging up selected plants. The first large-scale native grass plantings in riparian areas 

were undertaken at Kopta Slough. A total of about 300 acres were planted.

In 1991, TNC began riparian revegetation at the Stony Creek Preserve (Alpert et al. 

1999; Griggs and Petersen 1997; Reiner and Griggs 1989). Stony Creek Preserve was 

the first site purchased by TNC with the intention of transferring property ownership 

to the USFWS after it was restored, which occurred after 1995. Restoration technology 

was further refined at Stony Creek as the soils were more variable and irrigation timing 

and amounts were refined for most species that were planted. Implementation staff was 

of 6-month interns (mostly recent college graduates). Time and costs of inexperienced 
USDA Forest Service RMRS-GTR-411. 2020. 189

implementation became obvious. Hiring of a professional field manager with experienced 

laborers was decided for future projects. Approximately 500 acres were planted.

TNC began work at the Llano Seco Unit of the Sacramento River National Wildlife 

Refuge in 1991 (Griggs and Golet 2002; Silveira et al. 2003). Established by the U.S. 

Congress in 1987, the Sacramento River National Wildlife Refuge consists of 28 Units in 

Tehama, Glenn, Butte, and Colusa Counties comprising 10,353 acres within the 100-year 

floodplain (recent mixed alluvium and gravel bars/sandbars) along 81-river miles of the 

middle Sacramento River from Red Bluff to below Princeton (J. Silveira 2016, personal 

communication). The USFWS obtained these fee-title properties from landowners who 

were willing to sell existing riparian forest and flood-prone agricultural fields adjacent 

to the forests. There are roughly 4,581 acres of remnant riparian habitats on the refuge. 

TNC established a cooperative management agreement with the USFWS that allowed 

TNC to restore former farm fields. “Propagules from indigenous plants and local ecotypes 

are being used in large-scale restoration layouts of various designs associated with sitespecific hydrologic and edaphic conditions” (Silveira et al. 2003). 

As of summer 2015, 5,033 acres of former flood-prone agricultural lands (primarily 

walnut, almond, and prune orchards, but in some cases row crops) have been restored 

to various riparian and floodplain vegetation types (J. Silveira 2016, personal 

communication).

Rangeland Restoration Projects

The planting of riparian vegetation for gully control in mountain meadows has a 

long history. Kraebel and Pillsbury (1934) published a handbook for erosion control 

in mountain meadows in the Sierra Nevada for the USDA Forest Service in which 

they included specifications for the selection and planting of willow cuttings and the 

construction of willow wattles.

There were a number of projects in the 1980s to restore wet meadows and their 

associated streamside riparian buffer strips in eastern California. Project work typically 

involved the installation of erosion control devices, typically rock, gabion, or fabric 

grade-control structures (in other words, check dams), across eroded stream channels 

in high mountain meadows to trap sediment and raise the water table in the meadow. 

These projects were generally accompanied by the installation of temporary or permanent 

fencing to control livestock use, a reduced stocking level, or a revised grazing regime 

(for example, season of use). Many of these projects were planned using the Coordinated 

Resource Management and Planning (CRMP) Process because they involved work on 

both public and private land, including public land grazing allotments.

Sample rangeland riparian restoration projects include: Willow Creek Restoration 

(Clay 1984) in Modoc County and the Clark Canyon Riparian Demonstration Area (Key 

1987; Key and Gish 1989) in Mono County.

This work was similar in approach to the demonstration projects on Bear Creek 

(Elmore and Beschta 1987) and Camp Creek (Elmore and Beschta 1987; Winegar 1977), 

conducted by Wayne Elmore in the Prineville area of southeastern Oregon (Crook 

County) during the 1960s and 1970s. However, Wayne Elmore’s emphasis was less on 

the installation of structures and more on the improved management of rangeland riparian 

areas (Elmore and Beschta 1989). 
190 USDA Forest Service RMRS-GTR-411. 2020.

Invasive Plant Removal/Control

The removal or control of invasive nonnative plants was often an initial action at 

riparian restoration sites. Thus, most of the projects mentioned above and described 

in table 7 had some element of invasive plant eradication or control. Some riparian 

restoration projects only involved the removal of invasive plant species. In these cases, it 

was assumed that native riparian plants would reestablish on the site(s) after competition 

for light, nutrients, and water was eliminated. Some of these projects were followed by 

plantings of riparian plants after the invasive species had been removed entirely or were 

under control. 

Many modern-day strategies for invasive plant control were developed during the 

latter part of the 20th century. The Nature Conservancy played a significant role in sharing 

information on control techniques with the preparation of Element Stewardship Abstracts 

initiated in the late 1980s and early 1990s. Formed in 1992, the California Invasive Plant 

Council (Cal-IPC) contributed to the development and exchange of information on the 

eradication of invasive plants in riparian corridors through its symposia, publications, and 

sponsored research and trainings. Bossard et al. (2000) provides technical information on 

physical, biological, and chemical control measures for most of the invasive plants that 

occur in California’s riparian corridors.

There are many nonnative invasive plant species known to occur in California’s 

riparian corridors and these were generally dealt with on a project by project basis. Of 

these, two species in particular had virtually overtaken many stream corridors: giant 

reed and tamarisk. Below, we mention only four projects of the literally hundreds, if not 

thousands, of invasive plant removal projects that occurred between 1970 and 2000.

Giant Reed

Giant Reed (also referred to as Arundo and false bamboo) was recognized early 

on as a threat to the success of riparian restoration projects, not only because it often 

occurred on restoration project sites, which could be dealt with during site preparation 

and site maintenance, but also because flooding is the primary mechanism of dispersal 

of stems and rhizome fragments (Rieger and Kreager 1989). In the 1980s and 1990s, 

giant reed control projects were undertaken in river drainages in many parts of the State 

including the southern California coast, the central coast, the San Joaquin and Sacramento 

valleys, and even the north coast. Most restoration projects dealt with a specific reach 

of a stream and in many cases funding and/or authorization was not available to address 

upstream infestations of Arundo in the watershed. However, there have been some 

efforts at watershed-wide invasive plant removal. In northern California, in the Russian 

River Watershed (Gaffney and Gledhill 2003), community-based organizations have 

worked in collaboration with agencies, landowners, and community members since 1992 

to identify and map invaded sites, conduct experimental and demonstration projects, 

remove giant reed, restore native habitat, and conduct education and outreach programs. 

In southern California, a large-scale effort called the Santa Ana River Watershed Arundo

Habitat Management Program has been underway since 1997. Its purpose has been to to 

rid an entire watershed (largest drainage in coastal southern California) of Arundo and 

restore riparian areas (SAWPA 2016). Native riparian habitat has expanded into at least 

60 percent of the reclaimed floodplain, providing valuable habitat for birds such as the 

endangered least Bell’s vireo (Zembal and Hoffman 2007). 
USDA Forest Service RMRS-GTR-411. 2020. 191

Tamarisk

Removal or control of tamarisk or saltcedar and related species was one of the biggest 

challenges in the southeastern part of California, although it also had to be dealt with 

elsewhere in southern and central California. Saltcedar control in the Southwest had 

been attempted by various agencies and organizations since the 1950s (Rodman 1990), 

although not necessarily for the restoration of native ecosystems. TNC began removing 

tamarisk on the Coachella Valley Preserve in 1986 to restore native desert fan palm oases 

and associated riparian species through the Thousand Palms Canyon Tamarisk Control 

Project (Barrows 1993). 

Numerous saltcedar removal projects were undertaken along rivers in southern 

California in the 1980s and 1990s. Some of these projects involved seeding or planting 

with native riparian species while others relied on “natural” recovery from nearby 

seed sources or the seed bank. Bay and Sher (2008) evaluated the success of “active 

revegetation” (with no irrigation) after Tamarix removal in riparian systems in the 

Southwest including sites along the Lower Colorado River in California. An example of a 

large-scale effort to remove saltcedar and restore riparian vegetation is the Afton Canyon 

Riparian Restoration Project (BLM 2015) on the Mojave River begun by the Bureau of 

Land Management in 1992. 

Part 3—Research Associated With Restoration Projects

Various types of research, experimentation, field investigations, and monitoring 

programs were associated with early riparian habitat restoration projects. These programs 

addressed the following types of issues:

• Habitat requirements of target wildlife species

• Planting and irrigation techniques

• Plant survival and growth

• Plant, soil, and water relationships

• Competition from weeds and weed management techniques

• Wildlife usage of revegetated areas

• Proper functioning condition

Information on most of the projects used as examples below is presented in table 7.

Habitat Requirements of Target Wildlife Species

Significant amounts of data were often collected prior to the design of riparian 

revegetation projects, especially those projects intended to create habitat for special status 

species.

Lower Colorado River

In 1973, Bertin Anderson, Robert Ohmart, and John Discano began conducting 

studies of riparian vegetation-wildlife interactions on about 198,000 acres of riparian 

vegetation along the lower reaches of the Colorado River in an attempt to discover 

the vegetative characteristics to which birds were responding (Anderson and Ohmart 

1977; Anderson and Ohmart 1979; Anderson et al. 1979). They developed a model for 

revegetating riparian areas from their 5-year database on vegetative-wildlife interactions 

(Anderson et al. 1979). In 1977, after clearing saltcedar from an area along the Lower 
192 USDA Forest Service RMRS-GTR-411. 2020.

Colorado River, a revegetation design was developed for the site based on this model. 

While some palo verde volunteered on the site, willows, cottonwoods, and honey 

mesquite were planted in June-July of 1977. They monitored growth rates, root growth, 

plant mortality, and other factors for the first year and presumably longer. Anderson and 

Ohmart (1985a) later refined their modeling based on 7 years of data (1972-1979). 

Yellow-billed Cuckoo Habitat at the Kern River Preserve

Anderson and Layman (1989) based the design of yellow-billed cuckoo habitat at the 

Kern River Preserve on data of cuckoo habitat requirements along the Colorado River 

from 1976-1983 (Anderson and Ohmart 1984) and on data collected along the South Fork 

Kern River from 1985-1988.

Least Bell’s Vireo Habitat in San Diego County

In 1986, Hendricks and Rieger (1989) analyzed data of least Bell’s vireo nesting 

sites on the Sweetwater, San Diego, and San Luis Rey rivers in San Diego County 

representing approximately 10 percent of the known species population. A variety of 

parameters at each nest site were measured to characterize the nesting habitat of the least 

Bell’s vireo for the design of future restoration projects. Baird and Rieger (1989) used this 

baseline vegetation and habitat data for 30 nesting sites on the three rivers to develop a 

habitat restoration model for the creation of least Bell’s vireo nesting habitat at Caltrans 

mitigation sites in San Diego County (Baird 1989).

Note: Chapter 8 (Conservation Recommendations) of the Riparian Bird Conservation 

Plan (RHJV 2004) provides recommendations pertaining to the design of riparian 

restoration projects for riparian dependent bird species.

Planting and Irrigation Techniques

Lower Colorado River

In 1979 and 1980, Bertin Anderson studied the effects of deep tillage (augered holes to 

various depths) combined with irrigation (daily irrigation discontinued after variable lengths 

of time) on the growth and survival of rooted cuttings of Fremont cottonwoods on 74 acres 

of dredge-spoil sites along the Lower Colorado River (Anderson 1989; Anderson et al. 

1984). In 1981, Disano, Anderson, and Ohmart described the types of irrigation systems they 

used for riparian zone revegetation along the Lower Colorado River (Disano et al. 1984).

Kern River Preserve

Beginning in 1986-1987 at TNC’s Kern River Preserve, Dr. Bertin Anderson 

conducted research on ways in which the physical characteristics of the site (soil salinity, 

soil texture, depth to groundwater) and the type and height of saplings affected plant 

survival and growth. He evaluated various methods for the propagation of cottonwoods 

and willows. He also evaluated plant survival and growth with various irrigation regimes 

at various depths of tillage (augered and then backfilled holes) with respect to the water 

table. In addition, he studied the effects of competition from weeds and the effects of 

browsing by wildlife and livestock (Anderson 1989; Anderson and Layman 1989).

Lower Coyote Creek

At the 4-acre Coyote Creek Pilot Revegetation Project site installed in 1986-1987, 

John Stanley, Larry Silva, Harold Appleton, William Lapaz, and others conducted a 
USDA Forest Service RMRS-GTR-411. 2020. 193

3-year study on the effects of two different types of irrigation (overhead and flood) on 

the survival and growth of 3,640 plants comprised of 15 native plant species and multiple 

types of plant materials (propagule types) for each species. (Stanley et al. 1989)

Plant Protection

Frank Chan of PG&E developed the collar and screen plant shelter in the early 1980s. 

This device was used at the Lost Canyon Rehabilitation Project (Chan and Wong 1989). 

The collar (initially a cottage cheese container with the bottom cut out inserted into the 

ground) collected and concentrated precipitation into the root zone of the seedling and 

helped to deter gophers while the wire screen (tied to the collar) prevented damage from 

deer, rabbits, and insects. This device was used for direct seeding (for example, acorns) on 

numerous restoration project sites including at Coyote Creek and Cosumnes River. 

Plant Installation

In the mid-1980s, Jonathan Oldham and Bradley Valentine of the Kings River 

Conservation District developed the “hydrodriller” for the planting of woody cuttings 

on streambanks at the Crescent Bypass Riparian Revegetation Project. This device was 

similar to the “waterjet stinger” described by the USDA Natural Resource Conservation 

Service (Hoag et al. 2001).

Translocation of Live Vegetation

Sutter et al. (1989) studied the survival of transplanted mature elderberry shrubs. The 

elderberries were transplanted as mitigation for the loss of habitat for the valley elderberry 

longhorn beetle along the Sacramento and American rivers. 

John Rieger used a tree spade to translocate more than 2,000 cottonwood and 

willow trees that were being removed a mile upstream from the Mission Trails site for a 

previously approved development. The trees were installed at the Caltrans Mission Trails 

site and monitored for 5 years. Less than one-half of 1 percent mortality was observed 

(Rieger et al. 2014).

Application of Agricultural Methods

In 1989, the California Nature Conservancy gained management of the 700-

acre Kopta Slough Preserve, which is owned by the Controller’s Trust of the State of 

California. The Kopta property had been a productive almond orchard until the flooding 

of 1986 that killed over 60 percent of the orchard trees by drowning. The Nature 

Conservancy’s goal was to demonstrate the feasibility of implementing large-scale 

riparian restoration by planting a minimum of 100 acres in one year. The farmer leasing 

the agricultural land at Kopta was hired to provide advice on farming technology. Only 

large acreage restoration would make a difference to target wildlife populations along the 

Sacramento River (Griggs 1993). Achieving large acreage restoration goals would require 

the use of modern agricultural technology and equipment. This resulted in the “farming” 

of native trees and shrubs for the first 3 years of growth to ensure their establishment, 

once irrigation and weed control were halted. 

Plant Survival and Growth

Lower Colorado River

Bertin Anderson (Anderson et al. 1979) determined growth and mortality rates for 
194 USDA Forest Service RMRS-GTR-411. 2020.

palo verde, Goodding willow, cottonwood, and honey mesquite planted along the Lower 

Colorado River in 1977. In 1979 and 1980, Anderson (Anderson and Ohmart 1979; 

Anderson et al. 1984) measured the growth of rooted cuttings of Fremont cottonwoods 

planted in holes augered to various depths. Trees were measured (height and crown) three 

times during the growing season. Trees were irrigated daily with irrigation discontinued 

after varying lengths of time. Evaluations were made of the effects of deep tillage on tree 

growth and survival. The effects of differing periods of irrigation on growth and survival 

were also evaluated.

Lower Coyote Creek Pilot Revegetation Project

Each of the 3,640 plants installed at the 4-acre pilot revegetation site was monitored 

annually for the first three years (1987-1989). Assessments included survival, growth, 

vigor, and damage. A vegetation sampling program was employed to document 

semiannual changes in canopy height, canopy cover, foliage density, and diversity at 

different heights within the canopy and herbaceous cover. (Stanley et al. 1989) 

First San Diego River Improvement Project (FISDRIP)

Monitoring at the 26.8-acre riparian revegetation component of FISDRIP involved 

the collection of data for comparison with the milestone performance standards that 

were established for the project (Burkhart 1989). Both quadrats and transects were 

used to evaluate the development of the riparian woodland and groundcover. Photodocumentation was an important component of the monitoring program. 

TNC Sacramento Valley Project

Implementation monitoring involved daily evaluations of irrigation and weed control 

needs by the field manager. At the end of the growing season either a complete census, 

or a very focused sampling at preselected locations, was carried out at each project 

site. Survival and height growth were measured for each species. These data informed 

subsequent designs as to soil-plant placements.

Plant, Soil, and Water Relationships

Sacramento River National Wildlife Refuge

Lands purchased for the Sacramento River NWR were adjacent to the river channel 

and still flooded periodically. Channel deposits of sand and gravel were common on these 

lands, interspersed with the more productive floodplain deposits of fine sand and silt. In 

the first years, which species could grow on which soil texture was unknown. Likewise, 

the depth to water table affected the rooting behavior of each species differently. As 

irrigation was applied, careful monitoring was carried out of soil moisture in the soil 

profile by electrical moisture probes that had been placed at known depths previously. 

After 2 years of growth, backhoe pits were excavated to expose the root systems of 

selected saplings. Root system architecture and depth of development were mapped for 

each species, thereby informing future planning of design on sites with variable soils. For 

example, cottonwood “aggressively” grows roots deep to find the water table. Backhoe 

pits revealed abundant cottonwood roots down to 15 feet below the surface in less than 2 

years; arroyo willow growing nearby had rooted only to 5 feet, given identical irrigation. 

Drought adaptations of native trees and shrubs was virtually unknown. Monitoring of 
USDA Forest Service RMRS-GTR-411. 2020. 195

growth on different soil textures and depths to water table informed future design based 

upon soil characteristics.

Competition From Weeds and Weed Management Techniques

Kopta Slough Preserve

Early on at Kopta Slough Preserve, TNC discovered what every farmer knows: weeds 

are not compatible with target plant growth. Even a “few weeds” are not conducive 

to optimum growth of the target species. Thus, aggressive weed control—herbicides, 

mowing, disking—throughout the season was required to achieve restoration and 

horticultural success. Lessons learned from several years of implementation were that 

herbicides are necessary only during the first year, with frequent mowing being sufficient 

for optimum plant growth during years 2 and 3. If herbaceous understory is to be planted, 

then herbicides or disking of the aisle between woody plants must be carried out each year 

to inhibit the production of weed seeds.

Wildlife Use of Revegetated Areas

A number of investigators implemented monitoring programs to document wildlife 

(especially avian) use of revegetation sites. Many of these programs also monitored avian 

use of mature riparian forest adjacent to, or nearby, revegetation sites.

Lower Colorado River Riparian Revegetation

Bertin Anderson, William Hunter and Robert Ohmart monitored avian use at three 

revegetation sites ranging in size from 25-74 acres along the Lower Colorado River from 

1977-1984 (Anderson and Ohmart 1984; Anderson et al. 1979; Anderson et al. 1989)

Kern River Preserve Yellow-Billed Cuckoo Habitat Enhancement

William Humber, Bertin Anderson and Reed Tollefson censused birds utilizing 

naturally occurring cottonwood-willow habitats on TNC’s preserve and on one 25-acre 

revegetation site implemented in 1986 and on two 25-acre revegetation sites implemented 

in 1987 (Hunter et al. 1989). Note: Appendix B of the Riparian Bird Conservation Plan 

(RHJV 2004) provides documentation on how birds responded to riparian restoration at 

the Kern River Preserve.

Lower Coyote Creek Pilot Revegetation Project

Based on a Fish and Wildlife Coordination Act Report issued by the U.S. Fish and 

Wildlife Service, the CE 404 permit for the Lower Coyote Creek Flood Control Project 

required the Santa Clara Valley Water District to monitor wildlife use of at least one 

riparian revegetation site for a minimum of 10 consecutive years after initial planting 

and then at year 15 and every 10 years thereafter for the life of the project17.

24In 1986, the 

Coyote Creek Riparian Station expanded a pre-existing bird banding program at lower 

Coyote Creek to include the 4.4 acre Lower Coyote Creek Pilot Revegetation Project 

site, which was installed in December 1986. In addition to the bird banding program, 

variable-radius circular plots were established in each of the three habitats (existing 

riparian corridor, newly installed pilot revegetation site, and the ruderal overflow channel 

landward of the revegetation site). A breeding bird census was also conducted in each of 

17 This permit requirement was later modified.
196 USDA Forest Service RMRS-GTR-411. 2020.

the three study habitats from March through July of each year. Other monitoring included 

mammal, reptile, and amphibian sampling and vegetation sampling within each of the 13 

variable-radius circular plots. (Rigney et al. 1989) 

Mission Trails Mitigation Project

Caltrans monitored the SR-52 (Mission Trails) Mitigation Site in San Diego for 

compliance with Section 404 Permit requirements. Vegetation composition and structure 

was monitored on the mitigation site to determine conformity with the Least Bell’s Vireo 

Habitat Restoration Model (Baird and Rieger 1989). Bird populations were monitored 

to determine presence or absence of least Bell’s vireos and nesting success. Success was 

defined as either a vireo pair nesting on site or no statistically significant differences 

between parameters on the mitigation site and those in functioning the vireo’s habitat 

(Hendricks and Rieger 1989). Least Bell’s vireos successfully nested on the site within 

1 year with three territories and in subsequent years several other pairs nested in the 

remaining areas of the site. In addition, the vegetative parameters established in the 

habitat model were met for most of the monitored “cells” within the 5-year monitoring 

period.

Sacramento River National Wildlife Refuge

The Nature Conservancy hired Point Blue Conservation Science (PBCS, formerly 

PRBO) to monitor bird use of restoration plantings. Over 11 years of data were collected 

to show species use of restoration plantings at different times after planting (Golet et 

al. 2008). PBCS data also showed vegetation structural trends (tree-shrub ratios and 

densities) that affected bird diversity. These results informed future planting designs. 

Golet et al. (2003) reported on songbird use within the Sacramento River Project Area 

(over 100 river miles from Red Bluff to Colusa) including lands within the Sacramento 

River National Wildlife Refuge. At horticultural restoration sites, riparian bird diversity 

increased significantly over time as the revegetation sites matured. 

Proper Functioning Condition

Afton Canyon Riparian Restoration Project

BLM conducted project monitoring using photoplot ground/canopy cover analysis 

and cross-sectional riparian plant frequency/cover trend analysis. BLM relied heavily on 

the use of the qualitative evaluation process referred to as Proper Functioning Condition 

Assessment conducted by an interdisciplinary team of specialists (BLM 2015).

Part 4 – Concluding Remarks

Progress in our understanding of riparian ecology and how to go about restoring 

riparian ecosystems proceeded at a rapid pace between 1970 and 2000. Many scientists, 

conservationists, land and resource managers, and volunteers contributed to the 

development of this field. Unfortunately, it was not possible to recognize the contributions 

made by many of the individuals involved. Between the 1970s and 2000s, some of this 

knowledge was recorded in the literature. However, since many restoration practitioners 

had little time and funding to publish their findings, many of the details of how restoration 

projects were conducted, their successes and failures, and lessons learned remain buried 
USDA Forest Service RMRS-GTR-411. 2020. 197

in project plans, specifications, monitoring reports, and unpublished final project reports. 

We offer our apology to anyone, and for any projects, we have overlooked in this chapter.

The authors hope that this overview provides those involved in promoting riparian 

conservation in the 21st century with an informative historical perspective regarding 

the evolution of the field of riparian ecosystem restoration in California. For those of 

our peers who played a role in the protection and restoration of riparian ecosystems in 

the 20th century, we hope that this read provided a pleasant trip down memory lane and 

reminded you that your contribution was worth the effort.

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January 18, 2017]
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Chapter 8. Sacramento-San Joaquin System 

F. Thomas Griggs and Stefan Lorenzato

Introduction

The Great Central Valley of California occupies 22,500 square miles (58,000 square 

kilometers) in the interior of northern and central California. At the time of the Gold Rush 

in 1849, nearly 1 million acres (1,600 square miles, 4,000 square kilometers) of riparian 

vegetation covered the Central Valley floor along with approximately an equal area of 

wetlands. The riparian area flourished in the large river basins and along river channels 

(Katibah 1984; Thompson 1961). The Central Valley is partly defined by the Sacramento 

River in the north, the San Joaquin River in the south, and the Delta where the two rivers 

meet and turn westward toward San Francisco Bay (figs. 27a and 27b). The valley is 

made up of a series of basins connected by the rivers, which form a distributary floodway. 

Before development, heavy winter and spring runoff would flow out of the river channels 

and drain to the basins until waters were deep enough to continue their flow to the Delta. 

As flows subsided, water would sit in the basins until evaporated or it seeped into the 

ground. The wetland and riparian lands were nourished by these flows and extended 

across the low-lying areas in the valley trough and basin sinks. The land surface consisted 

of shallow undulating ridges and swales, creating complex soil-water-plant interactions 

that provided a great diversity of hydrology, vegetation, water depth and velocities, and 

timing. The result was a rich and dynamic system that dependably provided a mix of 

microhabitats and physical features (Kelley 1989; Thompson 1961).

These features that provided such a diversified ecology hindered agriculture. As 

cropping systems expanded to take advantage of the rich soils, riparian forests were cut 

down, land leveled for ease of production, and waterways rerouted. Today, the Great 

Central Valley of California grows over 200 crops and generates over 32 billion dollars 

annually in agricultural revenue (CDFA 2015), primarily due to its unique, and highly 

developed, hydrology and the complex patterns of alluvial soils that support diverse 

agricultural practices. But only about 2 percent to 4 percent of the riparian habitats remain 

(California Department of Fish and Wildlife’s VegCamp Program 2011). 

Central Valley River Basins and Forces of Change 

The Great Central Valley is comprised of a set of linked basins (table 9). The basins’ 

edges are marked by low rises in the land surface that control drainage patterns of 

overland flow. The rivers of the Central Valley flow through the basins in relatively small 

channels from the foothill watersheds to the Delta where north and south flows combine 

and head west into San Pablo and San Francisco Bays (figs. 27a and 27b).

Historically, flood flows in the valley were often flashy, flood peaks emerging from 

the upper watersheds in high volumes over short periods. The broad, flat basins of the 

valley floor readily absorbed the flows that escaped the channels. With these flashy flows 

came sediments that similarly were carried away from the channels and deposited in the 

basins. The resulting mosaic of soils and residual moisture provide large expanses of land 
212 USDA Forest Service RMRS-GTR-411. 2020.

Basin Approximate acres

Butte Basin 99,877

Colusa Basin 318,225

Yolo Basin 178,268

Sutter Basin 368,697

Sacramento Basin 95,545

American Basin 246,402

Lower San Joaquin 104,500

Middle San Joaquin 370,000

Upper San Joaquin 484,000

Tulare 540,000

Table 9—Hydrologic distributary basins of the Central Valley of California and their areas. 

BUTTE BASIN

SUTTER 

BASIN

COLUSA

BASIN

YOLO BASIN

AMERICAN BASIN

SACRAMENTO BASIN

Cache Creek

Putah Creek

American R iv er

Yuba River

Cosumnes River

Moke lu mne River

Feather River

Sac ramento River

Stony Creek

Butte Creek

Legend

* Basin delineation by Steven M. Brumbaugh

based on Figure 9 from Gilbert (1917)

Flood Bypass

Reservoirs

Stream

0 6 12 18 24 3 Miles

 ̄

Figure 27a—Sacramento Historic Basins. All figures in this 

chapter are by Kevin Coulton, P.E. and CFM , Seungjin 

Baek, Ph.D., P.E., cbec eco engineering, as part of a 

USEPA Wetlands Development Grant to the California 

Department of Water Resources (grant CD-00T83701).

Figure 27b—San Joaquin and Tulare Basins.

San Joaquin River

Merced River

Chowchilla Riv er

San Joaquin River

Tuolumne River

Stanislaus River

Bear Creek

Owens Creek

Kings Ri ve r

Fresno River

Tule River

TULARE

BASIN

SAN JOAQUIN BASIN

Legend

* Created by Steven M. Brumbaugh

based on USGS basin delineation.

Reservoirs

Stream

0 10 20 30 40 5 Miles

 ̄
USDA Forest Service RMRS-GTR-411. 2020. 213

where flood-adapted plants flourished and riparian forests established. The interaction of 

vegetation with the flood flows further influenced flow patterns and an even finer scale 

of microhabitats developed, supporting a great diversity of plants and animals (Scott and 

Marquiss 1984). Until the Gold Rush, people generally adapted to the patterns of flood 

and seasonal growth, exploiting the diversity and relying on natural cycles to produce the 

materials and food needed for their communities (Kelley 1989).

With the Gold Rush (1848-1855) came disturbance of land in the middle and 

upper watersheds and a tremendous amount—1.5 billion cubic feet (Mount 1995)—of 

soil, sediments, and rocks found their way into the streams and rivers from the mining 

operations. At the same time, agriculture was being established in the valley and farmers 

sought out the richest soils, which tended to be where the riparian forests stood. Farmers 

rapidly deforested the valley floors and carved canal systems into the land to move 

irrigation water diverted from the rivers, during summer low flows, into their fields. 

A conflict rapidly emerged between valley farmers and miners over the sediments in 

the rivers and streams. Because the mining accelerated the erosion and transport of coarse 

sediment in the form of sand, gravel, and cobbles, when floods occurred, these sediments 

were deposited on fields that previously were only comprised of premier farm soils. The 

coarse sediments devalued the land and made farming difficult or, in some cases, not 

possible (James and Singer 2008). 

The controversy was taken into court and in 1884 Judge Lorenzo Sawyer put a stop 

to the discharge of mine tailings into streams (Federal Reporter Vol 18, 1884; Mount 

1995). While this stopped the input of new coarse sediment, by this time millions of tons 

of debris were already loaded into streams and on their way down to the valley. Farmers 

seeking to stop flooding of their land in order to reap more crop production also wanted to 

stop the deposition of sediments on their fields. Initially, the solution was for individuals 

to build berms and levees, but the haphazard product did not resolve the issue (Kelley 

1989). Ultimately, in the early 1900s, a unified system of levees and floodways was 

designed to provide farmers with more predictable and manageable flood flows and to 

control and move sediment through the system (fig. 28). 

While the farm interests were well served by the resulting flood protection systems, the 

riparian habitats of the Great Central Valley were rapidly succumbing to agricultural demand. 

The diversity and abundance of wildlife and plants that once graced the valley began to 

shrink and the dynamics of the ecosystem began to change (Grinnell and Miller 1944). 

Another change force that emerged was the need for additional summer water for 

agriculture. By the mid-1900s, dams had become a common element of the landscape. 

With the large dam building era in full swing, major dams were placed on all the major 

rivers that drain into the Central Valley. The large dams created three very distinct 

alterations to the ecology of the Central Valley: (1) they restructured the pattern of 

flow, greatly limiting the medium sized peak winter and spring flows while increasing 

late spring and summer flows, but without significant peaks (Kondolf 1997); (2) they 

dramatically increased the acreage of land under irrigation; and (3) they cut off access by 

anadromous fish to their natal streams and the floodplains they used for rearing grounds 

(Sommer et al. 2003).

The plan for fish was to replace the loss of fish from lack of spawning habitat with 

hatchery bred fish. The result was a steady decline in populations of migrating fish. With 

this decline came a disruption of the nutrient cycles associated with fish returning from 
214 USDA Forest Service RMRS-GTR-411. 2020.

Figure 28—Sacramento and San Joaquin Valley Flood Control Systems.

the ocean and resulted in further reduction in riparian productivity (Merz and Moyle 

2006). With the flow patterns disrupted, plants and animals that had evolved to be in 

synchronization with the natural cycles were left to find limited room to persist at the 

edges of the remaining floodplain resource. Of the nearly 1 million acres of riparian lands 

estimated to exist in the Sacramento Valley before the Gold Rush (Katibah 1984), by 2011 

only 94,000 acres remained (VegCamp 2011). 

What We Currently Face 

The Central Valley is over 450 miles long, from its northern end at Redding to near 

its southern end at Fort Tejon (south of Bakersfield). The Sacramento River drains the 

northern portion of the Central Valley (27,500 square miles watershed), including the 

southern Cascade Mountains and northern Sierra Nevada (fig. 27a). Its primary tributary 

is the Feather River (6,000 square mile of watershed). The Sacramento empties into the 

Delta estuary between the City of Sacramento and San Francisco Bay. The southern 

portion of the Central Valley is drained by the San Joaquin River (31,800 square miles of 

watershed) and its larger tributaries—the Mokelumne, Cosumnes, Stanislaus, Tuolumne, 

and Merced rivers (fig. 27b). South of Fresno is the Tulare basin, home of historical 

Tulare Lake, once the largest freshwater lake west of the Mississippi River. The Tulare 
USDA Forest Service RMRS-GTR-411. 2020. 215

basin receives flows from the Kings, Tule, and Kern Rivers from the southern Sierra 

Nevada and overflows into the San Joaquin Basin southwest of Fresno. The San Joaquin 

River flows northward and empties into the Delta estuary from the south. The Delta is the 

single point of outflow for the entire valley.

Flow Patterns

Dams now control river flows on all these rivers. These dams serve both water supply 

and flood control purposes. The dams were conceived with the goal of providing water for 

agriculture and for urban uses. Two large water projects dominate the water conveyance 

systems: the Federal Central Valley Project (which has storage capacity of 13 million 

acre feet [maf] that delivers about 7.4 maf annually); and the California Water Project 

(30 dams, 20 reservoirs, 700 miles of canals, and storage capacity of 5.4 maf) (California 

DWR Bulletin 132-14, 2015). Together these projects serve over 3 million acres of 

irrigated land and supply over 25 million people with at least a portion of their water.

In an average year, the unimpaired runoff for the Sacramento Valley is about 18.2 

maf and the San Joaquin Valley about 6 maf (based on SRR and SJR unimpaired flow 

estimates—DWR Bulletin 132-14). For the Tulare basin, which only flows into the 

San Joaquin in the wettest years, the average unimpaired runoff is 3 maf. Diversions 

dramatically change the timing of the water flowing through the river system. And 

much of the water release from the dams does not reach the Delta to become outflow. 

The 8 River index, a measure of water released from the Sierra Nevada watersheds 

compared with Delta Outflow (as estimated by the Day Flow model) (see CDWR web 

site https://water.ca.gov/Programs/Environmental-Services/Water-Quality-MonitoringAnd-Assessment/RTDF-Summary) shows substantial losses as water moves through the 

system. For example, in 1993, the 8 River index showed about 19 maf of water entering 

the Central Valley but only about 5 maf leaving the Valley through the Delta). 

Note that these models produce coarse estimates; actual amounts are likely higher. 

These model comparisons also indicate how context-specific the flows are. In drier 

years, the gap between inflow and outflow tends to be larger than in in wetter years. In 

consecutive wet years, the difference is even smaller (presumably due to lack of demand), 

and in consecutive dry years the differences again tend to be small (mainly due to lack of 

water supply). For riparian areas this means that in all but the wettest years there exists 

fierce competition for water needed to support riparian functions.

Dam building started in the 1800s and accelerated through the 1950s. By the early 

1960s, dams had been built on the mainstems of the Sacramento and San Joaquin rivers 

and most tributaries before they entered the Central Valley from the mountains. Table 

10 shows the amount of storage in several of the larger reservoirs created by these dams 

(Kondolf and Batalla 2005). Of particular interest, note the amount of storage on each 

river as a percent of average annual runoff for that stream. For example, reservoirs on the 

Stanislaus River are able to trap 2.94 times the average annual runoff; on the Tuolumne 

and Calaveras Rivers nearly twice the average annual runoff is contained in reservoirs. 

In contrast, note that Shasta reservoir can only capture 74 percent of the Sacramento’s 

average annual runoff. Despite the flashiness of runoff in California, the large storage 

capacity in the San Joaquin reservoirs greatly diminishes the frequency of large floods 

and nearly eliminates the occurrence of modest floods on these rivers. Flooding is a more 

common occurrence on the Sacramento River. 
216 USDA Forest Service RMRS-GTR-411. 2020.

In both the Sacramento and San Joaquin Valleys on the valley floor, a system of 

floodway bypasses are maintained to convey flows greater than what can be carried within 

the river channels. In the Sacramento Valley, the mainstem rivers are used to convey water 

supplies for agriculture and urban use while also functioning as a conduit for flood waters. 

The combined purpose has resulted in floodplains isolated from their river channels, with 

the timing of flows shifting away from winter and spring flood pulses to late spring and 

summer steady medium flows. This has simplified the hydrology of the rivers. Coupled 

with the transformation of riparian lands to farms, the riparian habitats have been left with 

truncated processes and limited area to support the historic mix of plants and animals 

(Hunter et al. 1999; Mount 1995). As a result, much of the riparian habitat is now also 

simplified and shrunken to a point that species that were once common are now absent 

or rare. Vaghti and Greco (2007) describe seven contemporary plant communities of the 

Central Valley and trends in species declines.

For example, the least Bell’s vireo was the most common species in the willow 

thickets along the Sacramento River in the 1930s (Grinnell and Miller 1944; Howell 

2010). It disappeared during the 1950s, retreating to coastal rivers in southern California. 

Yellow-billed cuckoo formerly ranged as far north as the Columbia River, but today 

the species persists in very low numbers only along the Sacramento River. Likewise, 

sawgrass (Cladium maricus ssp. californicum) was an important basket-making plant 

for the Yokuts of the San Joaquin Valley (Latta 1938). Today, sawgrass is apparently 

extinct from the Valley (F.T. Griggs, personal observation). Slough thistle (Cirsium 

crassicaule) formerly grew in riparian areas in the San Joaquin Valley from Kern County 

northward into the Delta with collections as recently as the late 1930s (CalFlora). Today 

one population persists on the Kern NWR. Numerous species of native fishes have 

disappeared in the Delta due to introduced aggressive competitors (Sommer et al. 2003).

River Processes

Riparian ecology is driven by physical river processes of flooding (magnitude 

and timing) and sediment transport (geology of watershed, bank, and surface erosion; 

River

Average annual 

runoff 

M3 x 109

Total reservoir storage 

capacity 

m3

 x 109

Impoundment ratio 

(As percent of 

annual runoff)

Sacramento Valley

Sacramento at Keswick 7.278 5.384 0.74

Feather at Oroville 5.215 6.714 1.29

San Joaquin Valley

Mokelumne at Camanche 0.744 1.032 1.39

Calaveras at New Hogan 0.205 0.396 1.93

Stanislaus at Knights Ferry 0.202 3.518 2.93

Tuolumne at La Grange 1.772 3.444 1.94

Merced at Snelling 1.290 1.305 1.01

San Joaquin at Friant 2.095 1.140 0.54

Table 10—Central Valley rivers with impoundment ratios of average annual flow for each. (Excerpted from Kondolf and Batalla 

2005.) Each watershed contains more than two major reservoirs.
USDA Forest Service RMRS-GTR-411. 2020. 217

floodplain deposition; and channel meander) (Stillwater Sciences 2007). Riparian plants 

and wildlife are adapted to these river processes. River processes also determine plant 

community succession and the evolution of its physical structure. Flooding eliminates 

species that are not adapted to flooding, such as many invasive plants and animals, 

especially rodents. Native floodplain-adapted plants can tolerate flooding for weeks or 

months. Flooding recharges the local groundwater table that provides moisture to all 

plants through the growing season. The season after a flood typically sees season-long 

growth by trees and shrubs that tap into the abundance of soil moisture. 

Flood flows disperse seeds of many plants and deposit them in appropriate locations 

for germination and establishment (Stella et al. 2011). The timing of flood flows, 

especially the recession limb of the hydrograph, is critical for seedbed preparation, seed 

germination, and seedling development (Mahoney and Rood 1998). Seeds generally 

need the moisture available in the soil during the recession of flood waters, but seeds also 

require the oxygen available in those drying soils. If the seeds instead fall on dry ground, 

they cannot sustain their growth, desiccate, and die. Similarly, if they are subjected to 

repeated flooding, as often occurs in rivers managed for irrigation water, the seeds are 

starved for oxygen and die. 

Many plants have co-evolved with the flood timing of the Central Valley to release 

seed when flood waters recede, such as Fremont cottonwood and several willow species. 

With the dam-imposed flow patterns, much of the seed bed preparation and some of the 

timing queues have been removed, greatly reducing the success of floodplain-adapted 

plants establishing in recently flooded lands. This disrupts the cycles of disturbance and 

succession common to floodplains and limits the quality and extent of riparian vegetation 

communities. Birds, fish, and other animals similarly have developed life cycles that 

synchronize with the flood cycles and, with the lack of cycles imposed by dam and 

floodway operations, are showing similar declines in their populations (Strahan 1984). 

Plotting these natural cycles against river hydrographs of current and former flow regimes 

(figs. 29a and 29b) can give a good indication of the potential of disruption to riparian 

systems. 

Riparian vegetation on the valley floor arranges itself in relation to the magnitude 

and duration of flow. As riparian plant communities age and flow changes, a predictable 

sequence of events unfolds. This sequence involves a shift in community structure from 

plants adapted to high flow unstable channels to plants suited to slow water and less 

frequent inundation (Griggs 2009). High flow plants tend to be highly flexible and of 

relatively short stature that can easily lay down in the face of fast flood waters. As the 

plant community ages, the flexible stemmed plants shift the hydraulics in very localized 

spaces, which then allow other plants to take advantage of the relative calmer flows. This 

leads to even more stable plants, eventually leading to a mixed elevation canopy that 

includes ground level plants, mid-height shrubs, and various trees. 

The early stage of the full mixed canopy riparian forest is marked by a high density 

of trees. As the tree-canopy develops, light and shade become dominant characteristics of 

the forest and many of the trees succumb to the dense shade, leading to fewer and fewer 

stems per acre (Swiecki and Bernhardt 2003). The riparian gallery forest of 80-100 years 

old has relatively few large trees with various shade tolerant shrubs, vines, and ground 

covers underneath the canopy. It is one that grades for relatively short-lived flexible plants 
218 USDA Forest Service RMRS-GTR-411. 2020.

0

10000

20000

30000

40000

50000

60000

70000

October November December January February March April May June July August September

1940 Pre‐Dam “Wet”

1952 Post‐Dam “Wet”

1939 Pre‐Dam “dry”

2013 Post‐Dam 

“Dry”

SEED 

DISPERSAL

SEED 

DISPERSAL

ACTIVE GROWTH

DORMANCY PERIOD

ACTIVE GROWTH

DORMANCY PERIOD

COTTONWOOD Sacramento River @ Keswick

SANDBAR WILLOW

Fall Run Late‐fall Run Winter Run Spring Run

* Adapted by Steven M. Brumbaugh from USBR Trinity River Restoration Program Summary. McBain and Trush, Inc.

0

5000

10000

15000

20000

25000

30000

35000

40000

October November December January February March April May June July August September

1938 Pre‐Dam “Wet”

San Joaquin River @ Friant

1952 Post‐Dam “Wet”

1934 Pre‐Dam 

“Critically Dry”

2015 Post‐Dam 

“Critically Dry”

SEED 

DISPERSAL

ACTIVE GROWTH

DORMANCY PERIOD

ACTIVE GROWTH

DORMANCY PERIOD

COTTONWOOD

SANDBAR WILLOW

* Adapted by Steven M. Brumbaugh from USBR Trinity River Restoration Program Summary. McBain and Trush, Inc.

SEED 

DISPERSAL

XX

Figure 29a—Sacramento River Hydrograph.

XX

Figure 29b—San Joaquin River Hydrograph.
USDA Forest Service RMRS-GTR-411. 2020. 219

at the water’s edge to groves of oak and walnut in areas set away from the channels. 

Managing for flood suppression and water supply often disrupts this plant community 

evolution, diminishing the occurrence of gallery forests. (Griggs 2009). One of the 

important research questions is: What is the successional trajectory of the riparian plant 

communities under the modern hydrology?

Managing for irrigation supply will tend to produce an ecologically dysfunctional 

flow pattern that increases flows in spring and summer when the natural pattern would be 

declining flows, as on the Sacramento River (fig. 30). Many dammed rivers—Stanislaus, 

Tuolumne—have hydrographs that are essentially flat lines. We do not have indications 

that native species can adapt to such extreme changes. For example, because of the 

management of Shasta dam, today it is difficult to find stands of seedlings or saplings 

anywhere on the point bars or floodplain.

Sediment Transport

Dams, levees, bypasses, farming, road building, and urban encroachment into 

floodplains have all contributed to changing the patterns of sediment scour, transport, 

and deposition. Riparian assemblages take advantage of each of the variations in scour, 

transport, and deposition of sediment. The shifts in sediment dynamics have led to 

different niches and microclimates that in turn have shifted the character and quality 

of riparian areas (Rood et al. 2003; Stella et al. 2011) Most dams effectively capture 

sediment moving in the channels above the dam and leave the water released from the 

dam stripped of its sediment. The suspension and transport of sediment requires energy. 

Figure 30—Dates for seed-release by water-dispersed riparian tree species and the mean monthly discharge for the Sacramento 

River pre-Shasta dam and post-Shasta dam.
220 USDA Forest Service RMRS-GTR-411. 2020.

The energy of flowing water provides the forces needed to keep sediments suspended or 

moving downstream. When sediment is deposited in reservoirs above dams, the water 

released has no sediment load to occupy its energy. 

The water is then capable of working on the bed and banks and tends to erode these 

features until equilibrium in energy is reached where the energy of the water is balanced 

by the energy needed to suspend and transport sediment. The water released from the 

dam that contains little or no sediments has been termed “hungry water” (Kondolf 1997) 

because it is much more erosive than water that is full of sediments. Bank erosion and 

channel scour are often seen below dams. The sediment scoured away often exhibits 

particle size distributions different from what was being carried in the channel above 

the dam. Where these sediments fall out of suspension and are deposited, they can harm 

seedbeds for plants and nursery grounds for fish.

A feature of the shift in flow produced by dams is that the larger, more energetic flows 

of moderate floods are eliminated. Despite the impacts of hungry water, eliminating flood 

peaks stifles the process of channel meander, bank erosion, and bar formation that is a 

normal part of riparian areas. Rivers are dynamic systems; they continually adjust their 

bed and banks to the forces of flowing water (Knighton 1998). The result is a regular 

process of building bars and eroding them, moving banks, and rebuilding them. These 

processes depend on the more energetic, higher intensity flows to provide the energy 

needed to mobilize the bed and banks. 

Deposition occurs as the storm events pass and the flows subside (Knighton 1998). 

Riparian vegetation has evolved to take advantage of these cycles. Species are adapted 

to colonizing newly formed bars, but they give way either to even greater flood flows or 

to encroachment from other plants better suited to the stability created by the colonizers. 

The result is a succession of plants with distinctly different characteristics. This plant 

sequence is similarly colonized by animals that align with the plant communities present. 

By making flows very constant, the dynamics of the system are eliminated and the ability 

for the plant succession to occur is greatly restricted. One example of this is when mid 

channel bars are covered with unnatural vegetation that would have been removed if more 

natural flow patterns were experienced. The loss of open bars has consequences for fish as 

well, limiting their spawning habitat, and often introducing algae or other organisms that 

pose problems for fish.

Levees and Flood Bypasses

Most of the river banks in the Central Valley mainstem rivers and tributaries are lined 

with levees. Together with the bypass areas, also bounded by levees, these features further 

restrict riparian vegetation. Many of the levees of the Central Valley have supported 

riparian trees in the past, but recent policy initiatives (Central Valley Flood Protection 

Plan 2016) have set in motion a process to remove trees from levees. Removing levee 

vegetation threatens to further fragment functional riparian areas and disaggregate riparian 

corridors. Levee construction along the river channel also separates, or disconnects, the 

floodplain from the river, eliminating flooding of the floodplain, sediment deposition 

on floodplains, groundwater recharge, and access to shallow water habitat that many 

organisms need for key life stages. 

It has been argued that the flood bypasses provide floodplain functions (Sommer et 

al. 2003). But in many respects, the bypasses do not mimic floodplains. The bypasses 
USDA Forest Service RMRS-GTR-411. 2020. 221

are used for crop production. Farmers have rearranged the land surface to improve their 

farm operations. By grading and orienting flow to accommodate water inlets or drainage 

outlets, farmers have recast the surface that floods flow across, leaving uniform surfaces, 

sometimes tilted in directions inconsistent with natural flow, and without topographic 

variation. This simplified form removes potential for microhabitats and specialized niches. 

With this large-scale regrading of the land surface comes a dramatic decline of the area 

supporting riparian vegetation. The sparseness of native plant structure further fragments 

and isolates floodplain features and diminishes the processes that once supported riparian 

ecology.

Levees impose another significant ecological limitation. They isolate the floodplain 

from the river channel and thereby limit access of fish, plants, and other organisms 

to the highly productive shallow water habitat once found in vast areas of the Central 

Valley. For example, salmon have recently been shown to grow much faster on inundated 

floodplains than in the corresponding river channels (Jeffres et al. 2008). The floodplains 

(and inundated agricultural fields) produce vast quantities of food for fish that cannot be 

produced in the deeper, colder, and swifter river waters. Levees limit the ability for fish to 

gain access to this critical food supply.

Riparian Floodplain Vegetation

Simplified Native Structure

Due to these modifications in flow patterns and sediment transport, today’s floodplain 

is variously covered by stands of native woody species within a matrix of both woody 

and herbaceous invasive species. The remnant stands of riparian vegetation are dominated 

by trees and shrubs of later successional development, such as valley oaks, elderberry, 

and box elder. Early-successional species—Fremont cottonwood and several willows—

occupy a relatively small proportion of the vegetation today (Strahan 1984). Many of 

the terrestrial wildlife species that are rare or eliminated from the Central Valley riparian 

are dependent upon the early-successional vegetation (yellow-billed cuckoo, least Bell’s 

vireo, southwestern willow flycatcher, yellow-breasted chat). 

Invasive Exotics

Invasive, nonnative plants compete with native species for space and resources. 

Woody invasive species include giant reed (Arundo donax), Saltcedar (Tamarix spp.), 

Himalaya blackberry (Rubus procerus), hybrid black walnut (Juglans X hindsii), water 

primrose (Ludwigia peploides), and perennial pepperweed (Lepidium latifolium), with new 

invaders, as Sesbania punicea and Chinese pistachio (Pistacia chinensis), arriving annually. 

Numerous herbaceous agricultural weeds cover the floodplain to the exclusion of any 

native herbs. These species are well adapted to the current farm and floodway management 

practices, while native plants often struggle under these management paradigms. 

Agricultural and Urbanized Floodplains

The cities of Sacramento, Stockton, and Marysville-Yuba City have covered 

thousands of acres of floodplain with homes and businesses, with more recent 

construction still occurring on high quality agricultural land west of Stockton (into the 

Delta) and north of Sacramento (Natomas area). These portions of the floodplain are cut 

off from the river by tall levees. The rich, alluvial soil of the floodplains that are protected 
222 USDA Forest Service RMRS-GTR-411. 2020.

by levees and not under residential and business development produces very large 

economic returns for agriculture and the surrounding community.

Summary

Severely broken river processes that change the ecology for plants and animals 

include the following:

(1) Water availability: flooding—reduced magnitudes, and shifted timing, diversions, 

and groundwater pumping.

(2) Sediment: trapped in reservoirs above dams, scour and erosion below dams

(hungry water), particle size change and different deposition patterns, lack of 

floodplain deposits, channel and bank erosion.

(3) Hydrograph: inverted or flat, constant high spring summer flows, loss of flood

events/peaks.

(4) Floodplain: land leveling and homogenization of land surface, loss of niches, loss

of hydrologic connections (primarily due to levees).

(5) Native vegetation: loss of vegetation community connectivity, simplification of

species complex, significant loss of extent (95 percent gone) of flood-adapted 

vegetation.

(6) Exotic/invasive species: competition with natives, exploits farming operations.

(7) Urban encroachment.

Current Ideas and Actions for Riparian Management 

In 1987 the Sacramento River National Wildlife Refuge was created by the U.S. 

Congress. Early-on willing sellers of flood-prone agricultural land adjacent to the river 

came forward to sell their marginally productive farmland for conservation purposes 

(Reiner and Griggs 1989). The properties were not only flood prone, but the soils were 

heterogeneous with areas of channel deposits (sand and gravel) mixed with the more 

productive silt and clay loams. This resulted in soils less than ideal for agricultural 

production and consequently of low economic return. Flooding and heterogeneous soil 

patterns were ideal for restoration of structurally diverse riparian vegetation that would 

function as habitat for numerous characteristic riparian wildlife species, as well as specific 

target species (sidebar 1) (Griggs 1993, 1994). 

Riparian Obligate Species

Yellow warbler

Common yellowthroat

Yellow-breasted chat

Song sparrow

Black-headed grosbeak

Blue grosbeak

Warbling vireo

Yellow-billed cuckoo

Southwestern willow flycatcher

Riparian Associate Species

Lazuli bunting

Nuttall’s woodpecker

Ash-throated flycatcher

Bewick’s wren

Spotted towhee

Sidebar 1—Focal Bird Species That Guide Riparian Restoration Design
USDA Forest Service RMRS-GTR-411. 2020. 223

Horticultural Restoration

Horticultural restoration involved planting native trees and shrubs that are 

characteristically found growing on the alluvial soils along the channel and on the 

floodplain (Hujik and Griggs 1994a,b) (sidebar 2). These native trees and shrubs (and 

understory) would form the type of structure that target wildlife required (Griggs 2009). 

The vegetation that has grown in the restoration plantings has performed well (Alpert 

et al. 1999; Griggs and Golet 2002; Griggs and Peterson 1997; Griggs et al. 1993). Results 

of increasing wildlife use of the restoration plantings have been impressive (Borders et 

al. 2006; Golet et al. 2008, 2009; Williams 2010). Monitoring of bird populations was the 

most intensive and shows how different species colonized the sites over a decade. Valley 

elderberry longhorn beetle (listed as threatened) also moved into the planted elderberry 

shrubs. Special bird species—such as yellow-billed-cuckoo—nested in the new plantings, 

and least Bell’s vireo nested in a restoration planting for the first time in 60 years in the 

Central Valley (Howell et al. 2010). 

As of this writing, approximately 9,000 acres have been planted by The Nature 

Conservancy, River Partners, Department of Water Resources, U.S. Army Corps of 

Engineers, California State Parks, and Resource Conservation Districts since 1989 in the 

Sacramento Valley, all on public lands (Refuges and Preserves). An additional 2,500 acres 

have been planted in the San Joaquin Valley. Funding came from Federal and State grants 

and contracts. 

Levee Set-Backs

The levees that protect several cities and much farmland from large winter flooding 

were designed over 100 hundred years ago under very different flow and sediment 

Trees

Fremont cottonwood Populus fremontii

Valley oak Quercus lobata

Red willow Salix laevigata

Black willow Salix gooddingii

Large Shrubs

Arroyo willow Salix lasiolepis

Box elder Acer negundo

Oregon ash Fraxinus latifolia

Elderberry Sambucus mexicanus

Buttonbush Cephalanthus occidentalis

Flexible-stemmed Shrubs

Coyote brush Baccharis pilularis

Mulefat Baccharis salicifolia

Rose Rosa intermontana

Sandbar willow Salix exigua

Quailbush Atriplex lentiformis

Vines (Lianas)

Grape Vitus californicus

Pipevine Aristolochia californica

Clematis Clematis ligustifolia

Blackberry Rubus ursinus

Poison oak Toxicodendron diversilobum

Herbaceous Grasses

Creeping rye Leymus triticoides

Blue rye Elymus glaucus

Meadow barley Hordeum brachyantherum

Purple needlegrass Nassella puchra

Squirrel-tail Elymus elymoides

Broadleaves

Mugwort Artemisia douglasiana

Gumplant Grindelia camporum

Evening primrose Oenothera

Goldenrod Euthamia occidentalis

Sidebar 2—List of Species Used in Central Valley Riparian Restoration Projects
224 USDA Forest Service RMRS-GTR-411. 2020.

transport conditions than today. Hydraulic mining in the foothills of the northern Sierra 

Nevada in the late 1800s generated sediment that filled canyons and when in the Central 

Valley it filled channels causing flooding. Levee placement adjacent to the channel 

allowed the much deeper flows to scour and to move sediments through the system. 

Unfortunately, today the sediment has been flushed away and the artificially deep flows of 

“hungry water” from below dams are now eroding the levees at their bases.

One solution to this situation is to build new levees set back from the edge of the 

channel that would allow the channel to meander across the new floodplain that will be 

created. This would take away the intense scouring of the levees and provide acreage 

for riparian vegetation restoration. Benefits for flood maintenance would be the better 

absorption of peak flows, the reduction of maintenance costs, and the risk of levee failure 

(Greco and Larsen 2014; Larsen et al. 2006). The set-back area would become restored 

riparian vegetation to function as high quality wildlife habitat and an important means of 

replenishing local groundwater. See (www.multibenefitprojects.org), NGO web site.

Restoring River Processes

Ecological Flows

The ecological flows model (The Nature Conservancy et al. 2008) identifies the 

timing for dam releases to create flows large enough to affect sediment/scour processes 

and encourage recruitment of seedlings of Fremont cottonwoods and support the life 

histories of Chinook salmon, steelhead, green sturgeon, bank swallow, and western pond 

turtle.

Applying this hydrologic model to releases from a reservoir should mimic the natural 

hydrograph and encourage seedling establishment of cottonwood and willows. This has 

been carried out with success on the Truckee River through and downstream of Reno, 

Nevada (Rood et al. 2003). One issue that restricts this approach in the Central Valley 

has been dollar cost, or lost opportunity, from releasing water from dams that could have 

been used to grow crops.

Williams et al. (2009) have recently developed the Floodplain Activation Flow (FAF) 

calculation for regulated rivers to quantify the area of floodplain covered by the smallest 

flow (2-year recurrence) that will initiate biological activity. In this example, the 2-year 

recurrence flooding-interval was selected as necessary for juvenile salmon to move onto 

the floodplain. The flood must occur during March-May with a minimum duration of 7 

days. The FAF calculation estimates area of floodplain that will be inundated at specific 

flows for a specific reach of a river. The manager can define quantified ranges for the 

criteria, allowing for better planning of the hydrological/biological objectives for any 

project on the floodplain. 

Managed Hydrology in Farm Fields to Benefit Juvenile Rearing of Salmonids

Salmon and steelhead in their juvenile stage forage on the floodplain for invertebrates 

during spring floods—the same spring floods that prepare the seedbeds for the 

cottonwoods and willows. Due to dams and the altered flows, floodplains rarely flood 

today. Fish hatcheries raise juvenile salmon in concrete runways and feed them artificial 

food, releasing them into the river in April or truck them to the SF estuary and release 

them to the delight of invasive striped bass.
USDA Forest Service RMRS-GTR-411. 2020. 225

An alternative method for raising the juvenile salmon is to manage the hydrograph 

on an artificially flooded farm field (that functions as a pond) in the spring (February and 

March). The fish forage on native invertebrates in the pond water and rapidly increase in 

size and weight (Sommer et al. 2003). After 4 to 6 weeks in the pond, the fish are released 

into the river in March or April (Jeffres et al. 2008). This is a relatively new concept with 

only four seasons of demonstration projects of varying acreage. Results appear promising 

with significant growth (3 to 5 times weight gain) during their 4 to 6 weeks of residence 

on the flooded fields.

Managing Riparian Vegetation for Flood Flow Conveyance and Ecological Benefits (Flood Management)

Riparian vegetation in Central Valley today occurs between flood management levees. 

The California Department of Water Resources (CDWR) is legally charged with the 

management of the majority of levees and the floodway for the purpose of Public Safety 

(CDWR https://water.ca.gov/Programs/Flood-Management) and Greco and Larsen (2014) 

(fig. 28). Farmers working the lands between and outside the levees continue to try to 

optimize conditions for commercial production with little or no attention to river process or 

ecological quality. In many cases riparian vegetation is seen as a threat or nuisance. Thus, 

flood management and farming goals supersede the ecological goals of riparian ecology 

everywhere outside of the Federal refuges. 

The challenge today is how to manage the landscape such that public safety, farming, 

and ecological riparian benefits are all realized for people and wildlife. The discussion 

below is an attempt to describe how ecological management of riparian vegetation can fit 

into the logic of flood managers that think in terms of risk and system reliability.

All of the flood-prone lands, floodways, and flood bypasses in California readily grow 

plants. These sites possess some of the most fertile soils on the planet, and when coupled 

with even a little moisture, allow vast number of species to grow. Whether those plants 

are useful for us and the ecosystem depends on how and what is managed. Native floodadapted plants offer many advantages over exotic plants. (Micheli et al. 2004). They better 

support the historic ecology, are better adapted to the timing and processes needed to 

restore ecological functions, persist better under flood and drought and are therefore more 

predictable and reliable, and support native animals better than exotics. 

Floodway Design for Flow Conveyance

Flood managers use two-dimensional hydraulic computer modeling for designing 

flows through the floodway. Recent work using two-dimensional computer models for 

flood flows shows that plants can be arranged to assist with managing flows to reduce 

flood risk. Until recently, flood modelers typically looked at vegetation as a problem or, at 

best, a necessary evil in managing floodways. Understanding of how plants can be used 

to help direct flood flows, reduce scour, and control erosion and deposition is now taking 

hold among flood modelers (Anderson et al. 2006; Stone et al. 2013). With this awareness 

comes the prospect of designing floodways that integrate ecological goals with risk 

reduction and economic goals and constraints (Fischenich 2006). Promoting native floodadapted plants has the added advantage of assisting agriculture by reducing the seed banks 

of farm weeds and limiting the way weed seeds can be deposited on farm fields.

Managing Flood Risk With Riparian Vegetation

In flood management, the risk of property damage or injury of people has been 
226 USDA Forest Service RMRS-GTR-411. 2020.

paramount. Recently the idea that risk extends to the ecosystem has been brought into 

flood management discussions. In the Central Valley, the large number of threatened and 

endangered species that occupy floodplains and riparian lands means that flood managers 

must now consider the implications of their flood designs not only on traditional risk 

factors but also on features of the landscape that support the listed species. Assessment of 

risk has largely been accomplished using computer models. In these models, vegetation is 

associated with a roughness factor and contributes to defining the maximum water surface 

elevation during a flood event (Aberle and Jarvela 2013; Freeman et al. 2000). Modelers 

have typically adjusted the roughness factor to get model outputs to align with calibration 

points (water surface elevations known from previous flow events). This practice has led to 

vegetation removal justified by a need for smoother surfaces as depicted by the models. With 

the widespread adoption of two- and three-dimensional flood models (Wilson et al. 2006) 

we now have the ability to look beyond the simple application of roughness as a calibration 

parameter and consider roughness as a design feature (McKay and Fischenich 2011).

Emerging work is showing that native flood adapted plants can be equal to, or less 

rough than, bare soils (Kavvas et al. 2009). Two-dimensional modeling is also showing 

that a combination of roughness values can be used to our advantage in lowering flood 

risk (Soong and Hoffman 2002). Older models averaged the measurements of many 

parameters. Two-dimensional and three-dimensional modeling allows those parameters to 

be treated more independently with the result that we can now design flow paths within 

floodways that provide for ecosystem processes to be expressed while still managing 

flood risk at low levels.

With these new modeling capabilities, we can now discuss ecological risks, the role 

and placement of specific plant species, mixed age classes, mixed canopy structure, 

and other ecologically important attributes in the context of reducing risk to people and 

property (Fischenich and Copeland 2001). Using these tools we can visualize how complex 

roughness can be applied to protect sensitive areas, be they levees and homes or delicate 

habitats. Achieving the desired mix of roughness can be accomplished by placement of 

plants with appropriate physical structure. This dramatically changes the perspective of 

vegetation in flood management. Instead of being an impediment that increases flood risk, 

vegetation now becomes a valuable tool used to reduce and optimize risk exposure. 

Improving Flood Management Reliability With Riparian Vegetation

In considering reliability we can again borrow from the world for flood management. 

Reliability is the potential for a particular feature or facility to provide a repeated 

outcome. Flood managers want levees and floodways to be reliable at conveying flows 

and keeping flood waters away from valued property and people. Floodways and levees 

undergo constant change that in turn shifts their reliability. Much of the change arises 

because of the ability for plants to colonize and grow in flood systems. Maintenance of 

levees and floodways is a constant challenge where floodway operators remove vegetation 

to maintain flow capacity (reduce roughness). Denuding these features provides 

temporary improvements in capacity but also primes the sites for recolonization and 

starts the cycle over. The maintenance needs and costs are considerable and maintenance 

programs are regularly facing shifting budgets and damage repair. 

As noted above, succession in floodplain and riparian plant communities is marked by 

predictable changes in the density of plants, the number of stems per area, and height of the 

leaf canopy above the ground (Griggs 2009). The typical change in roughness peaks after 

colonization when vegetation communities are transitioning to longer lived, larger species 
USDA Forest Service RMRS-GTR-411. 2020. 227

and then declines as increasingly shade limits understory growth. Floodway maintenance 

schedules tend to remove vegetation at peak roughness, creating a disturbance that is 

somewhat unpredictable, leading to lower reliability in system performance. 

Sidebar 3—Using Stand Design for Flood Risk Reduction

An alternative approach to periodic wholesale removal of high roughness vegetation is to deliberately 

manage to transition the plant community to mixed age class structures that include substantial 

percentages of older, lower roughness stands. This can be done by implementing a rotational thinning 

process. Two-dimensional flood modeling of Star Bend on the Feather River has illustrated the point. 

At this site, high roughness on the floodplain creates a threat of increasing channel scour along the 

left bank of the river channel where a water diversion facility is located. A rotational thinning option 

for the floodplain was modeled that removed a swath of plants longitudinally through the center of 

the floodplain at time zero. At 10-year intervals, additional thinning occurred along adjacent, parallel 

swaths. After four cuttings, the site had an age class distribution of 0-, 10-, 20-, 30-, and 40-year-old 

vegetation. Each cutting was sized to retain acceptable scour potential at the water diversion (fig. 31). 

At each thinning episode, a smaller swath than the previous cut was needed to maintain acceptable 

conveyance and roughness. At the 5th interval, it was estimated that normal tree fall and plant 

community dynamics would sustain the needed roughness distribution without further thinning. 

While hypothetical, this exercise illustrates the potential for integrating plant stand management with 

risk reduction to provide a dependable, repeatable, reliable conveyance system. Rather than depend 

on regular infusions of funds for clearing plants, this option consistently reduced costs until a selfmaintaining system emerged. At this point, the system is highly reliable, being self-regulating, and 

funding can be reduced to that needed for periodic assessment to ensure the stand retains its risk 

reduction capabilities. At the same time, the mixed age class structure increases the biodiversity and 

hydraulic complexity of the site, providing for a better mix of ecological process and function.

Figure 31—Rotational thinning that targets complex age class forest. Floodplain velocities change over time with vegetation 

growth and management resulting in higher/distributed velocities across floodplain and lower velocities (reduce erosion 

potential) through channel.
228 USDA Forest Service RMRS-GTR-411. 2020.

Monitor and Map

To achieve a greater expanse of Central Valley riparian habitat, it will be necessary to 

understand the quality and extent of existing habitat. Recent work has produced a finescale vegetation map of the riparian habitat of the Central Valley (California Department 

of Fish and Wildlife’s VegCamp 2011). Conducted as part of the flood planning work 

for the Central Valley, the new vegetation maps provide data layers for GIS use, for the 

first time providing a valley-wide picture of where riparian vegetation flourishes. Now 

that a baseline has been established, changes in riparian vegetation communities can be 

tracked. Future monitoring will provide insights into how plants and animals are doing in 

the reconstituted riparian forests and whether these areas can produce the same scale of 

diversity and sustainability that once occurred.

Climate Change

The riparian landscape we experience going forward will not likely be the same as 

what we lost, or even what we have. Climate models predict a much hotter California 

in the future, especially in the San Joaquin Valley and Tulare Basin. Whether that heat 

comes with a shift in the amount of precipitation is unclear. But the models do predict 

more variation in precipitation and a loss of snowpack (State of California, Department of 

Water Resources 1993). This will likely result in more frequent flood flows of moderate 

scale and an increase in the frequency of occurrence of large damaging floods. The 

large dams and reservoirs provide a great deal of flood risk reduction for the modest 

events but are likely to operate much like a free-flowing river during large events, when 

the reservoirs are full and the dams must release everything coming at them. In these 

situations, we must anticipate where the flood flows will go and how the system can 

recover from damaging, high volume flows. One strategy is to put more flood capacity 

back on the floodplains, widen the bypasses, and to stabilize the most vulnerable flow 

paths using native floodplain adapted vegetation to maintain the floodways (Seavy et al. 

2009). This is a strategy that merges flood risk management with conservation ecology 

in a sustainability context. Doing this would allow a revitalization of the Central Valley’s 

riparian landscape, and the opportunity to take advantage of many contemporary and 

emerging tools in riparian assessment and flood modeling.

Conclusion

California’s Central Valley once supported vast arrays of riparian and wetland habitats 

that consisted of complex topography and microhabitats. Today much of that ecosystem 

has been replaced with industrial agriculture and, more recently, with expanding urban 

landscapes. The combination of climate change and evolving social needs will continue to 

change the landscape structure of the Central Valley going forward. Riparian ecosystems 

can play an important part in providing a sustainable landscape, posing acceptable 

levels of risk from flood, drought, and other disruptions. To achieve this will require the 

combined efforts of land managers, conservationists, city and regional planners, and 

business interests all working to find common strategies that provide the desired mix of 

habitat features. New and emerging tools in flood modeling and conservation can be used 

to help visualize and design these future landscapes. Clear objectives will be needed to get 

the most value from these efforts.
USDA Forest Service RMRS-GTR-411. 2020. 229

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USDA Forest Service RMRS-GTR-411. 2020. 233

Chapter 9. Recreation Habitat Versus Ecological Habitat in 

Riparian Areas: Can We Manage for Both? 

Bo Shelby and Doug Whittaker

Introduction

Yosemite Valley is a narrow, steep-walled canyon with fantastic natural areas, and its 

scenic and ecological wonders inspired the National Park concept and preservation ethic. 

But the Merced River through Yosemite Valley is one of the most heavily used non-urban 

riparian areas in the world. Over 4 million people visit each year, and 90 percent go to the 

box canyon where towering granite summits rise over 4,000 feet above the meandering 

river. About 8 miles long and less than one mile wide, the valley has a mix of riparian 

vegetation, open meadows, and forests. There are also roads, bridges, trails, acres of 

parking lots, campgrounds, four hotels or lodges, gift shops, and restaurants concentrated 

in the valley’s eastern end. With over 600 campsites and 1,000 hotel units, and 7,000 

vehicles and 20,000 people on busy days, the Valley’s iconic resources are challenged by 

overuse and development. 

The latest attempt to address these problems is the 2013 River Management Plan 

for the Merced, a designated part of the National Wild and Scenic Rivers System. 

Two previous plans were overturned by lawsuits brought by environmental groups, 

eventually leading to a settlement agreement filed with the Ninth Circuit Court that guided 

development of the third plan. The settlement identified many issues, including traffic 

and parking congestion, facility overdevelopment, attraction site crowding, and riverside 

impacts that degrade scenic, recreation, and ecological values. Because of our work on 

carrying capacities in Grand Canyon and other river corridors, we were identified in the 

settlement agreement to assist with the Park’s planning effort. 

Much of Yosemite Valley’s 150 years of development has occurred within a halfmile of the river, resulting in many riparian impacts. At the largest scale, the 1879 breach 

of the El Capitan glacial moraine to drain seasonally wet meadows lowered the river 

bed by 5 feet and increased channel incision and subsequent erosion (Milestone 1978). 

As development increased, infrastructure was protected from flooding and lateral bank 

erosion through revetments (riprap), log jam removal, in-channel sediment mining, 

and riparian plantings. By the late 1970s, about 2.5 miles of the river’s banks had been 

channelized, mostly near East Valley roads, hotels, and campgrounds. Infrastructure 

included as many as 38 bridges, some constricting the natural channel by 50 percent and 

collectively altering the river’s natural dynamism (Cardno Entrix 2012; Milestone 1978). 

Despite narrowing at bridges and revetments, the overall result was increased channel 

widths and bank instability (Madej et al. 1994).

In addition to these larger-scale changes, the Merced’s riparian areas received 

increasing levels of tourist activity. Hotels, cabins, campgrounds, and employee housing 

were often located within a hundred feet of the river or adjacent meadows, and networks 

of roads and trails provided access to riverside attractions. Infrastructure and access 

encouraged visitors to go to the river, and activities such as swimming, picnicking, and 

boating became important parts of the Yosemite Valley experience. Resultant impacts 

included spider webs of user-created trails in meadows and along the river, vegetation 
234 USDA Forest Service RMRS-GTR-411. 2020.

trampling along the banks, woody debris removal, and compaction/erosion around the 

roots of large trees (Madej et al. 1994). Coupled with larger-scale geomorphic changes 

that disconnected the river from meadows and riparian areas, these site-specific recreation 

impacts have become a major challenge for the Park Service. 

Park scientists often see these impacts through a restoration lens. Geomorphologists 

contemplate removal of riprap and bridges as obvious solutions to large-scale impacts. 

Similarly, riparian ecologists see exclusions of recreation use as the preferred way to 

address site-specific impacts and restore native vegetation and riparian function. These 

kinds of restoration plans could radically alter the amount, location, and function of 

infrastructure that currently handles the volume of use. These plans could also prevent 

visitors from enjoying many sites along the river. This applies pressure on the Park 

Service’s dual mandate to protect resources while allowing visitors to enjoy them. 

As a practical matter, it is politically and administratively unrealistic to substantially 

reduce the hotels, campgrounds, and roads that accommodate millions of visitors. Few 

experts expect Yosemite Valley to return to mid-19th century levels of development and 

use. The question then shifts to finding an appropriate balance between infrastructure, 

use, and restoration in a high profile management plan for one of the world’s best-known 

national parks. How do we address riparian impacts in such an environment? 

The answer lies in several seemingly simple questions. 

• What, exactly, are the impacts of concern?

• Where, when, and how do the impacts occur? 

• To protect riparian values, can we get people to change the way they recreate, or 

must they be excluded to protect sensitive areas?

The remainder of this chapter considers riparian impacts on the Merced, how visitors 

perceive them, and whether visitors support management actions that might be used to 

reduce impacts or restore riparian conditions. Data come from a study of river use and 

visitors (Whittaker and Shelby 2012), as well as broad indices of ecological health in the 

riparian zone and adjacent meadows.

Yosemite Valley’s Current Riparian Condition

Yosemite National Park has conducted several studies to document riparian ecological 

conditions (Cardno Entrix 2012). A comprehensive review of that work is beyond the 

scope of this chapter, but the plan focused on two primary indicators: (1) California Rapid 

Assessment Method (CRAM) scores, a multi-variable measure of riparian health; and (2) 

Largest Patch Index (LPI-5) scores for adjacent meadows, which measure fragmentation. 

Riparian Health

The CRAM protocol was developed for assessing overall health in unconfined 

riverine riparian areas. It measures 14 variables scored on a scale from “A = best 

achievable” to “D = worst commonly observed” (Collins et al. 2008). These scores, 

in turn, are combined to create four overall attribute scores (ranging from “0 = poor” 

to “1 = best achievable”), or a single overall CRAM score (by evenly weighting each 

of the attribute scores, also ranging from 0 to 1). The scores can be used to compare 

areas, changes over time, or the contributions of different attributes to overall scores. 

This provides an overall indicator of riparian health and identifies areas with higher 

impacts that need more intensive study. The 14 variables were organized into four major 
USDA Forest Service RMRS-GTR-411. 2020. 235

categories, including buffer and landscape context, hydrology, physical structure, and 

biotic structure.

CRAM scores were developed for 81 alternating-bank assessment areas along a 

16 km (9.9 mile) reach of the Merced River in Yosemite Valley (Cardno Entrix 2012). 

Overall scores ranged from 0.56 to 0.93 (median = 0.77), with generally lower scores in 

reaches adjacent to the higher density use and development in East Valley. 

Several metrics showed good condition and varied little across the study area (including 

landscape connectivity, water source, plant layers, and percent invasives), while other 

metrics influenced differences in overall scores (buffer condition, channel stability due to 

revetments, buffer width, biotic structure, and topographic complexity). Areas with the 

lowest scores tended to have bank protection (riprap) or extensive bank erosion, narrow 

or absent riparian buffers, few different types of habitat features, lower topographic 

complexity, and greater human use. Conversely, areas with higher scores had wider buffers, 

minimal bank protection, greater structural complexity, and lower human use. 

The assessment called attention to lower scores caused by channel stabilization 

around meander bends and near bridges, as well as bank erosion associated with human 

use from road turnouts, campgrounds, or hotels. Although CRAM scores are not designed 

to assess the relative influence of specific stressors (e.g., geomorphic issues from bridges 

and revetments vs. recreation impacts), the scores for Yosemite Valley suggest that 

recreation impacts in East Valley are part of the problem and need management attention. 

Meadow Fragmentation

The seasonally wet meadows adjacent to the Merced River are also connected to the 

Valley’s riparian values and threatened by use. Meadow function and health were assessed 

by a Largest Patch Index (LPI-5), which is the percent of a meadow in its five largest 

patches (adapted from McGarigal and Marks [1995] by Yosemite National Park staff). 

This measure is sensitive to sizes of intact areas and amounts of user-created trails, and it 

has been associated with impacts related to meadow hydrology, soil moisture, nonnative 

species, habitat quality, and barriers to small mammals (Forman 1995; Gaines et al. 2003; 

Holmquist 2004; Knight 2000; Leung et al. 2011a; Lindenmayer and Fischer 2006). In 

Yosemite Valley meadows, user-created trails and fragmentation impacts have been well 

documented and used to develop restoration priorities (Foin et al. 1977; Holmquist 2004; 

Holmquist and Schmidt-Gengenbach 2008; Leung et al. 2011b). 

Stoneman Meadow in the heart of the East Valley provides an interesting example. 

A 1978 photo shows an extensive web of user-created trails (see fig. 32), a total of 3,170 

meters or about 2 miles. The fragmentation index based on this photo is 40.40 percent (out 

of 100 percent possible), which met the National Park Service definition of a degraded 

meadow. The Park Service subsequently removed and restored most of these trails, then 

used barriers and walkways to funnel use to a single formal path across the meadow 

with a raised boardwalk surface. Over the next 30 years, the meadow recovered to the 

condition shown in figure 33. There was a 90 percent decrease in trails (to 327 meters), 

and the fragmentation index improved to 99 percent (nearly full restoration). Throughout 

this time, there had been no effort to limit numbers of visitors in the meadow, and from 

1978 to 2011 the annual number of people visiting the Park increased by 54 percent. The 

two sets of trails are superimposed in figure 34. Ongoing monitoring continues to assess 

meadow condition, use levels, and visitor compliance with barriers and walkways to 

better understand relationships among these variables.
236 USDA Forest Service RMRS-GTR-411. 2020.

Figure 32—1978 photo of trails 

and road through Stoneman 

Meadow in East Yosemite 

Valley (Merced River is just 

outside the top of this photo). 

National Park Service photo.

Figure 33—2011 photo of restored Stoneman Meadow (main trail with boardwalk is visible near the T-junction). National Park 

Service photo.
USDA Forest Service RMRS-GTR-411. 2020. 237

Recreation Impacts in Yosemite Valley’s Riparian Areas

For the sake of discussion, the photo in figure 35 shows examples of typical riparian 

impacts caused by recreation use. Riparian impacts are complex, but Park ecologists 

assessing this site would identify trampling, loss of vegetation, soil compaction and 

erosion, and tree root exposure. If asked to evaluate these impacts, we suspect few 

ecologists would find them acceptable. 

We were interested in how visitors perceive the same impacts. As part of a larger river 

study, we sampled 806 river users, including both floaters and shore users (Whittaker 

and Shelby 2012). The survey form included this photo (fig. 35), with the following 

instructions:

This river bank photo shows an area used by Park visitors along the Merced. 

National Park Service scientists evaluate river banks from an ecological perspective, but 

we are interested in how visitors perceive them. Please rate the acceptability of this river 

bank from your perspective.

Respondents answered on a 9-point acceptability scale; results are given in figure 36. 

Overall, 76 percent rated this river bank as acceptable, and an additional 13 percent rated 

it marginal; only 11 percent rated it unacceptable. This raises interesting questions about 

differences between scientists and recreation users, goals for preservation versus use, and 

definitions of acceptable impacts. 

Figure 34—NPS comparison of 

trail lengths and meadow 

fragmentation values (LPI-5) 

for Stoneman Meadow in 

1978 and 2011. 
238 USDA Forest Service RMRS-GTR-411. 2020.

Figure 35—Photo of impacted riparian area along the Merced River in East Yosemite Valley, evaluated in survey of river users. 

Photo by Confluence Research and Consulting.

Figure 36—River users’ ratings of riverbank conditions shown in the photo in figure 35. Data from Whittaker and Shelby 2012.
USDA Forest Service RMRS-GTR-411. 2020. 239

Where, When, and How Do Impacts Occur?

The impacts pictured in figure 35 occur along the Merced in Yosemite Valley, but 

they are not evenly distributed. They are found primarily in accessible places where trails, 

roads, and bridges approach the river, or where the river offers attractive features such as 

scenic views, beaches, or swimming holes. Reach-based riparian monitoring efforts (e.g., 

CRAM scores) help recognize the general problems of recreation impacts but are not 

specific enough to identify precise locations, types of impacts, or causes. 

Ecologists hypothesized that riparian impacts were associated with boating use, a 

popular summer activity that occurs on a 2.5-mile reach adjacent to the highest use areas 

in East Valley. A raft rental and shuttle concession provides 150 to 200 boats on peak use 

days (with two to four people per boat). Private floaters in rafts, canoes, kayaks, and tubes 

may add 50 to 100 boats per day. The Merced River Plan included proposals to limit or 

eliminate river running, so our study observed boaters and shore users to document the 

way they used riparian zones. 

Results showed that floaters mostly stayed on the water, in their boats or tubes. When 

they stopped on shore, it was primarily at open sandy beaches, where they generally 

stayed for short periods. Exceptions included the concession boating put-in (an eroded 

trail and unhardened, unvegetated bank next to a bridge) and trails leading from a sandy 

beach to a foot bridge commonly used for jumping (despite Park regulations that prohibit 

it). Floaters rarely stopped along vegetated riparian shorelines or steeper banks. 

In contrast, land-based visitors utilized networks of user-created trails into the riparian 

zone, which connect parking lots, road pull-outs, campgrounds, and Valley developments 

with river overlooks, beaches, and other open shoreline areas. Shore users often set up 

on sand beaches, but they also dispersed along upland terraces, under the shade of larger 

trees, or even on riprapped banks. These locations tend to become clearings with minimal 

understory, possibly attracting more use because they look like designed recreation sites 

(similar to those in campgrounds). 

Figure 37 illustrates riparian characteristics that are important for riverside recreation 

users. Observations from the river study indicate that river users are attracted to places like 

this, which we identified as “recreation habitat.” Characteristics of recreation habitat include 

a convenient location with easy access, a sand beach, good places to sit, views of the river 

or valley, sun (for swimmers/sun-bathers), and shade (for sun protection). The recreationrelated goals of visitors appear to be different from the ecological goals of scientists, so it 

makes sense that recreation habitat may be different from plant or animal habitat.

Are Users Willing to Change Their Impact-Related Behaviors?

Impacts depicted in the photos are seen differently by visitors versus ecologists. But 

do visitors care about ecological impacts, and are they willing to do anything differently 

to mitigate the effects of recreation use?

To measure support for management, study respondents were shown example photos 

of (1) split rail fencing and (2) a boardwalk and stairs. These are management actions the 

Park Service could use to reduce bank and meadow impacts associated with recreation 

use. Reasons and possible consequences were given, and visitors were asked to rate 

acceptability on a 9-point acceptability scale. The photos and questions are given below:
240 USDA Forest Service RMRS-GTR-411. 2020.

Figure 38—River users’ ratings of fencing and boardwalks. Data from Whittaker and Shelby 2012.

Figure 37—Visitors using riparian “recreation habitat” along the Merced River in East Yosemite Valley. Photo by Confluence 

Research and Consulting.
USDA Forest Service RMRS-GTR-411. 2020. 241

To reduce bank and meadow trampling along the river, the Park Service could close 

sensitive areas (see split rail fencing photo, fig. 40) and direct people toward areas that 

can withstand use (see boardwalk and stairs photo, fig. 41). However, these actions may 

decrease “naturalness,” prevent access to some areas, or lead to congestion in other 

areas. Please rate the acceptability of the following actions.

• Longer split rail fences (over 200 feet) to protect large areas from trampling, with 

short openings for river access.

• Shorter split rail fences (under 50 feet) to restore small sites with heavy 

trampling.

• Occasional boardwalks and stairs through meadows and sensitive areas to 

provide access to areas like beaches.

• Trail networks with many boardwalks and stairs directing use to less sensitive 

areas and discouraging off-trail use.

Figure 38 shows acceptability ratings for the four fencing and boardwalk options. 

The two lower development options (“occasional boardwalks and stairs” and “short split 

rail fencing”) were most acceptable (72 percent and 66 percent acceptable, respectively). 

Even the lowest-rated option (“trail networks with many boardwalks and stairs”) was 

acceptable for 53 percent, marginal for 20 percent, and unacceptable for only 27 percent. 

Respondents were also asked to evaluate more general approaches such as education 

and regulation, using a 5-point support-oppose scale. The three approaches were described 

as “education efforts that teach visitors to avoid sensitive areas,” “close user-created trails 

that lead into sensitive areas,” and “prohibit off-trail or off-beach use in sensitive areas.”

Figure 39 shows widespread support for such actions to protect ecological or aesthetic 

values along the river. There was 81 percent support for education, 73 percent support for 

closing user-created trails, and 62 percent support for prohibiting off-trail/off-beach use in 

sensitive areas. 

Figure 39—Support/opposition for education and regulations to address riparian impacts. Data from Whittaker and Shelby 2012.
242 USDA Forest Service RMRS-GTR-411. 2020.

Figure 40—Carefully designed recreation infrastructure can facilitate use and address riparian impacts.

Figure 41—Elevated light-penetrating boardwalks and stairs provide access and protect riparian vegetation on the Kenai River, 

Alaska; high densities of salmon anglers decimated banks before the advent of this simple but effective technology. Photo by 

Confluence Research and Consulting.
USDA Forest Service RMRS-GTR-411. 2020. 243

Management Implications Regarding Riparian Use Impacts

Data from the study indicate users do not commonly recognize riparian impacts, and 

they seek recreation habitat next to the river, where they contribute to these impacts. It 

may be difficult for the Park Service to keep visitors out of attractive recreation habitat, 

but the study shows broad public support for management actions to reduce impacts. 

These include defined trails, boardwalks, fences, and even closures, especially if applied 

judiciously. 

The challenge is to organize use in riparian areas so that visitors can access attractive 

recreation habitat that is less susceptible to ecological impacts. A hypothetical example 

in figure 40 shows how carefully developed recreation infrastructure can direct human 

use to appropriate areas in the riparian zone. Trails are designed to keep use away from 

sensitive banks and riparian vegetation, while delivering people to more resistant areas 

that provide good recreation habitat next to the river (e.g., beaches and purposely chosen 

viewpoints). Where trails unavoidably cross or arrive at wet or sensitive areas, hardened 

surfaces can constrain use to a minimal footprint and ensure impacts do not exceed 

acceptable levels. Split rail fences at the back of the beach keep users off vegetated banks 

and direct them to trail access at either end. 

Technological advances have improved the ecological benefits of hardened surfaces. 

For example, light-penetrating boardwalks and stairs (made of steel or aluminum, 

sometimes with fiberglass or composite grating) have become common on high density 

fishing streams in Alaska (see fig. 41). They allow diverse vegetation to grow underneath, 

preventing fragmentation-related impacts while providing anglers access to wading-based 

fishing areas. This is especially important on salmon streams, where month-long seasons 

can attract shoulder-to-shoulder densities commonly described as “combat fishing.” Such 

pressure would otherwise decimate riparian vegetation (King and Hansen 2001; Larson 

and McCracken 1998; Lieptz 1994; O’Toole et al. 2009). 

Data indicate users support this infrastructure and appreciate the benefits. A study of 

bank anglers on the Kenai River (Whittaker and Shelby 2010) showed that 71 percent 

support fencing and directed trail use (compared to 12 percent opposition), and 66 

percent support additional fishing platforms (11 percent opposition). There are tradeoffs 

for primitiveness and visual quality, but most of the Kenai is not “wilderness-like” and 

has extensive development (e.g., private homes with docks, along with dozens of parks 

and campgrounds with boat launches and other facilities). 

Conclusions

Differences between visitors and scientists or managers

The Merced River study echoes resource management literature that shows differences 

between visitors and scientists or managers for variables as diverse as depreciative behavior 

in developed campgrounds (Clark et al. 1971); acceptability of bare ground and fire ring 

impacts at wilderness campsites (Shelby et al. 1988); and evaluations of scenic beauty and 

recreation quality after different logging and silvicultural treatments (Brunson and Shelby 

1992). Differences in perceptions of impacts can come from different goals, different 

specialized training in science or management, or specific responsibilities for protecting or 
244 USDA Forest Service RMRS-GTR-411. 2020.

enhancing resources. Differences do not mean that either perspective is somehow wrong, 

only that we need to recognize and work with those differences. 

Recreation Habitat Is Different From Ecological Habitat

Recreation habitat is usually different from ecological habitat. This comes from the 

different goals and needs of recreation users in riparian zones, and it makes sense when 

we think of it this way. Resource managers often think carefully about providing for the 

needs of four-legged animals; the recreation habitat concept explicitly recognizes the 

need to do the same for two-legged ones. The recreation management field addresses the 

complexities of managing people in natural settings (Manfredo 2002; Manning 2011). 

While it is beyond the scope of this chapter to review this field, the conclusions listed here 

fit with literature that applies to riparian issues.

Users Endorse Ecological Goals

In spite of the visitor-scientist/manager differences listed above, visitors endorse 

ecological goals and express willingness to “be managed” to support them. It is possible 

to be cynical about this and assume that what people say is different from what they do 

(true, and well documented in the social psychology literature), or that even with majority 

support it only takes a few people behaving badly to defeat restoration efforts (also true, 

and well documented in the river management literature). But given this, how can our 

efforts be most effective?

Management Frameworks Can Help

Several decision-making frameworks have been developed by researchers to help 

planners and managers address recreation impacts or capacities, including Limits of 

Acceptable Change (LAC; Stankey et al. 1985); Carrying Capacity Assessment Process 

(C-CAP; Shelby and Heberlein 1986); and Visitor Experience and Resource Protection 

(VERP) (National Park Service 1997; Manning 2001). Each has differences in orientation, 

emphasis, terminology, and specific steps, but all were built on the same foundation of 

scientific and professional literature (Manning 2004), and they are more similar than 

different in their general approach. The concepts common to most include those discussed 

below. 

Clear Management Objectives

The resource literature has many admonitions about the need for clear management 

objectives (Heberlein 1977; Hendee et al. 1977). At the broadest level, clarifying 

objectives may show how scientists’ riparian restoration goals differ from visitors’ 

recreation goals. As objectives are further specified, we can identify areas of convergence 

and areas of divergence, possibly identifying locations where ecological protection trumps 

recreation access, or vice versa. Clear objectives will help with effective management of 

riparian areas and the visitors who want to use them.

Precision About Recreation Needs

We need to be precise about recreation needs, recognizing that we cannot provide 

for all uses in every location. Helpful information includes desired activities, recreation 

opportunities to be provided, specific groupings of necessary attributes, sites available, 
USDA Forest Service RMRS-GTR-411. 2020. 245

locations where sites are needed, sites to be created, and sites that can be eliminated. 

Demand for and supplies of recreation resources are not evenly distributed. Demand 

is often concentrated in time (e.g., seasons of the year, days of the week, hours of the 

day) and space (e.g., at places with attractions). Supply of recreation sites is also uneven 

(e.g., camping beaches that are available only at low water, steep-walled sections of a 

canyon that have fewer campsites). Successful management, from ecological or social 

perspectives, needs to recognize (and when possible take advantage of) such distributions, 

and solutions are site-by-site as well as cumulative.

Greater Challenges in Lower Development Settings

It is often easier to manage people in front-country settings; development is done with 

people in mind, and users expect infrastructure, education, and regulations to direct and 

enforce appropriate behaviors. In backcountry settings, many people value the lack of 

management (e.g., the Wilderness Act specifies “primitive and unconfined recreation”), 

even though recreation habitat is more likely to conflict with ecological goals, and small 

numbers of well-intentioned users can produce unacceptable impacts. 

Clarity About Impacts and Indicators

We need to be precise about impacts: what, where, and when. For example, on a 

cultural resources project, one stakeholder identified “impact to the entire river corridor,” 

while another specified “visitors standing on the hearth” (but did not want to say where 

that was). Without judging the validity of either concern, we can manage visitors most 

effectively if we know exactly what and where the problems are, and how they are 

connected with visitor use. Defining specific, measurable indicators is central to recreation 

management frameworks. Yosemite managers found that broader CRAM scores and 

fragmentation indices identified impact issues, but solutions required specific information 

about site-level impacts and the recreation users causing them. 

Specific Standards

Even with clear indicators, it can be challenging to specify when an impact goes from 

“acceptable” to “unacceptable.” Reporting the level of a particular impact is descriptive; 

deciding whether that level is acceptable is evaluative. This important distinction has been 

made in the recreation management (Shelby and Heberlein 1986) and riparian ecology 

(Schmidt et al. 1998) literatures. All resource management decisions have an evaluative 

component, and most resource conflicts come from disagreements about evaluative 

judgements such as goals, objectives, or standards.

Diverse Management Approaches

Recreation management literature identifies four general categories of management 

actions for visitor impacts: (1) capital developments that build things or modify the 

environment; (2) education to modify human behavior indirectly through persuasion; (3) 

regulation to modify human behavior more directly; and (4) visitor capacities that limit 

use. Folded into these actions are technological, cognitive, and structural fixes (Heberlein 

2012). For example, a light-penetrating walkway is a capital improvement using a 

technological fix that removes foot contact with the sensitive area, with a structural fix 

element that directs and concentrates use. It works only if it gets people where they want 
246 USDA Forest Service RMRS-GTR-411. 2020.

to go. Some impacts are unintentional, and education (a cognitive fix) might help. Each 

approach has complexities and tradeoffs.

How Are Impacts Related to Recreation Use?

The relationship between use and riparian impacts varies by the type and amount of 

impact: 

• Some impacts are related to amount of use, and the relationship is linear (e.g., 

more vehicles take more space).

• Some impacts are related to amount of use, but the relationship is non-linear; 

pioneering use causes the most impact, and additional use has proportionally less 

impact (Cole 1987; Hammitt et al. 2015; Kuss et al. 1990).

• Some impacts are related to type of use rather than amount of use (e.g., just 

one person using poor camp or human waste practices can cause unacceptable 

impact).

Successful management actions need to consider these relationships. For example, 

impacts with linear relationships to amount of use can be effectively managed with a 

carrying capacity. In contrast, impacts related to type of use rather than amount of use do 

not respond to capacities but do respond to other management actions (e.g., requiring fire 

pans or human waste carry-out). 

Education Is Not a Panacea; Persuasion Is Challenging

Education is sometimes seen as a panacea for human-caused impact problems As 

Roggenbuck (1992) states, “If people only understood the impacts they cause, we could 

get them to behave differently.” Education is also preferred by many managers because 

it is less obtrusive (Fish and Bury 1981), and users support the idea of “educating those 

other people who are causing the problem.” The complexities and pitfalls of education 

as a solution to environmental problems is well documented in Heberlein’s (2012) 

description of the “cognitive fix.” Education and enforcement may be important parts of 

any management program, but they require careful “engineering” and public support.

Managing for Ecological and Recreation Habitats

Managing for both ecological and recreation habitat is challenging. For example, 

higher intensity management (many boardwalks and fences in service of ecological 

habitat goals) may detract from naturalness (a recreation habitat goal). But in the long 

run, success with ecological habitat probably requires success with recreation habitat. We 

provide high quality experiences and nurture a stewardship ethic when people visit rivers 

and enjoy high quality riparian environments. 

References

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37-41.

Cardno Entrix. 2012. Merced River and riparian vegetation assessment. Concord, CA: Final Report 

prepared for Yosemite National Park, National Park Service. https://www.nps.gov/yose/learn/

nature/upload/Merced-River-Vegetation-Assessment-Cardno_Inc_June-2012.pdf/ [Accessed 

November 2018]. 
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Clark, R.N.; Hendee, J.C.; Campbell, F.L. 1971. Values, behaviors, and conflict in modern camping 

culture. Journal of Leisure Research. 3(3): 143-159.

Cole, D.N. 1987. Research on soil and vegetation in Wilderness: A state of knowledge review. 

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Research Station.

Cole, D.N. 1995. Experimental trampling of vegetation. I. Relationship between trampling intensity 

and vegetation response. Journal of Applied Ecology. 32: 203–214. 

Collins J.N.; Stein E.D.; Sutula, M.; Clark, R.; Fetscher, A.E.; Grenier, L.; Grosso, C.; Wiskind, 

A. 2007. California Rapid Assessment Method (CRAM) for wetlands. Version 5.0.1. 151 p. 

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November, 2018].

Fish, C.B.; Bury, R.L. 1981. Wilderness visitor management: diversity and agency policies. Journal 

of Forestry. 79(9): 608-612.

Foin, T.C.; Garton, E.O.; Bowen, C.W.; Everingham, J.M.; Schultz, R.O.; Holton Jr, B. 1977. 

Quantitative studies of visitor impacts on environments of Yosemite National Park, California, and 

their implications for park management policy. Journal of Environmental Management. 5(1): 1-22.

Forman R.T. 1995. Land Mosaics. New York: Cambridge University Press. 217 p. 

Gaines, W.L.; Singleton, P.H.; Ross, R.C. 2003. Assessing the cumulative effects of linear recreation 

routes on wildlife habitats on the Okanogan and Wenatchee National Forests. Gen. Tech. Rep. 

PNW-GTR-586, Portland, OR: U.S. Department of Agriculture, Forest Service, Pacific Northwest 

Research Station. 79 p. 

Hammitt, W.E.; Cole, D.N.; Monz, C.A., 2015. Wildland recreation: Ecology and management. 

Hoboken, NJ: John Wiley & Sons. 334 p. 

Heberlein, T. A. 2012. Navigating environmental attitudes. New York: Oxford University Press. 

228 p.

Heberlein, T.A.; Shelby, B., 1977. Carrying capacity, values, and the satisfaction model: A reply to 

Greist. Journal of Leisure Research. 9(2): 142-146.

Hendee, J.C.; Stankey, G.H; Lucas, R.C., 1978. Wilderness management. Miscellaneous Publication 

1365. Washington DC: U.S. Department of Agriculture, Forest Service. 363 p. 

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invertebrate fauna and patterns of abundance and biodiversity. El Portal, CA: U.S. Department of 

Interior, National Park Service, Yosemite National Park. Unpublished report.

Holmquist, J.G.; Schmidt-Gengenbach, J. 2008. Effects of experimental trampling addition and 

reduction on vegetation, soils, and invertebrates in Tuolumne Meadows. El Portal, CA: U.S. 

Department of Interior, National Park Service, Yosemite National Park. Unpublished report.

King, M.A.; Hansen, P.A., 2001. Assessment of shore angling impacts to Kenai River riparian 

habitats, 1998. Anchorage, AK: Alaska Department of Fish and Game, Division of Sport Fish, 

Research and Technical Services. 76 p. 

King, M.A.; Clark, R. 2004. Kenai River riparian habitat assessment. Report to the Alaska Board of 

Fisheries, 2005. Anchorage, AK: Alaska Department of Fish and Game. 42 p.

Knight, R.L; Cole, D.N., 1995. Wildlife responses to recreationists. In: Knight, R.L.; Gutzwiller, 

K.J., eds. Wildlife and recreationists coexistence through management and research. Washington, 

DC: Island Press: 51-69.

Knight, R.L. 2000. Forest fragmentation and outdoor recreation in the Southern Rocky Mountains. 

Boulder, CO: University Press of Colorado: 135-154.
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Kuss, F. R.; Graefe, A. R.; Vaske, J. J. 1990. Visitor impact management: A review of research. 

Volume 1. Washington, D.C.: National Parks and Conservation Association. 217 p.

Larson, L.L.; McCracken, B.W., 1998. Assessment of angler impacts to Kenai River riparian habitats 

during 1996. Anchorage, AK: Alaska Department of Fish and Game, Division of Sport Fish, 

Research and Technical Services. 49 p. 

Leung, Y.F.; Newburger, T.; Jones, M.; Kuhn, B.; Woiderski, B., 2011. Developing a monitoring 

protocol for visitor-created informal trails in Yosemite National Park, USA. Environmental 

management, 47(1): 93-106.

Leung, Y-F.; Bigsby, K.; Kollar, C. 2011. Developing methods for integrated analysis of meadow 

condition and informal trail data in Yosemite National Park. Technical report submitted to USDI 

National Park Service, Yosemite National Park. 170 p. 

Liepitz, G.S. 1994. An assessment of the cumulative impacts of development and human uses on fish 

habitat in the Kenai River. Technical Report No. 94-6, Anchorage, AK: Alaska Department of Fish 

and Game, Habitat and Restoration Division. 68 p. 

Lindenmayer, D.B.; Fischer, J. 2006. Habitat fragmentation and landscape change: An ecological 

and conservation synthesis. Washington, DC: Island Press. 317 p. 

Madej, M.A.; Weaver, W.E.; Hagans, D.K. 1994. Analysis of bank erosion on the Merced River, 

Yosemite Valley, Yosemite National Park, California, USA. Environmental Management. 18(2): 

235-250.

Manfredo, M.J. 2002. Wildlife viewing in North America: A management planning handbook. 

Jefferson, MO: Modern Litho. 299 p. 

Manning, R. 2001. Programs that work. Visitor experience and resource protection: a framework for 

managing the carrying capacity of National Parks. Journal of Park and Recreation Administration. 

19(1): 93-108.

Manning, R.E. 2010. Studies in outdoor recreation. Corvallis, OR: Oregon State University Press. 

468 p. 

McGarigal, K.; Marks, B.J. 1995. FRAGSTATS: Spatial pattern analysis program for quantifying 

landscape structure. Gen. Tech. Rep PNW-GTR-351, Portland, OR: U.S. Department of 

Agriculture, Forest Service, Pacific Northwest Research Station. 122 p. 

Milestone, J.F. 1978. The influence of modern man on the stream system of Yosemite Valley. San 

Francisco CA: San Francisco State University. Thesis. 227 p.

National Park Service. 1997. VERP: The Visitor Experience and Resource Protection (VERP) 

Framework – A handbook for planners and managers. Denver, CO: Department of Interior, 

National Park Service, Denver Service Center. 108 p.

O’Toole, A.C.; Hanson, K.C.; Cooke, S.J. 2009. The effect of shoreline recreational angling 

activities on aquatic and riparian habitat within an urban environment: Implications for 

conservation and management. Environmental Management. 44(2): 324-334.

Roggenbuck, J.W. 1992. Use of persuasion to reduce resource impacts and visitor conflicts. In M. 

J. Manfredo, ed. Influencing human behavior: Theory and applications in recreation tourism, and 

natural resources management. Champaign, IL: Sagamore Publishing: 149-208.

Schmidt, J.C.; Webb, R.H.; Valdez, R.A.; Marzolf, G.R.; Stevens, L.E. 1998. Science and values in 

river restoration in the Grand Canyon. BioScience. 48(9): 735-747.

Shelby, B.; Heberlein, T.A. 1986. Carrying capacity in recreation settings. Corvallis, OR: Oregon 

State University Press. 164 p. 

Shelby, B.; Vaske, J.J; Harris, R. 1988. User standards for ecological impacts at wilderness 

campsites. Journal of Leisure Research. 20(3): 245-256.
USDA Forest Service RMRS-GTR-411. 2020. 249

Stankey, G.; Cole, D.; Lucas, R.; Peterson, M.; Frissell, S.; Washburne, R. 1985. The Limits of 

Acceptable Change (LAC) system for wilderness planning. Gen. Tech. Rep. INT-176, Ogden, 

UT: U.S. Department of Agriculture, Forest Service, Intermountain Forest and Range Experiment 

Station. 37 p.

Whittaker, D.; Shelby, B. 2012. Boats, beaches, and river banks: Visitor evaluations of recreation on 

the Merced River in Yosemite Valley. Final study report to Yosemite National Park. Corvallis, OR: 

Confluence Research and Consulting. 74 p. 

Whittaker, D.; Shelby, B.; Meldrum, B.; DeGroot, H.; Bacon, J. 2012. September. Transportation, 

recreation, and capacities in Yosemite National Park. The George Wright Forum. 29(3): 338-350. 

Whittaker, D.; Shelby, B. 2010. Kenai River Recreation Study. Final study report to Alaska State 

Parks. Corvallis, OR: Confluence Research and Consulting. 149 p.
250 USDA Forest Service RMRS-GTR-411. 2020.

Chapter 10. Intended Versus Unintended Effects During 

Riparian Restoration Create High Quality Recreation 

Habitat 

Bo Shelby and Doug Whittaker

Introduction

Many of the same features that are good for riparian ecology are good for people. 

Varying flow regimes, diversity of plants and animals, dynamic geomorphology that 

produces open beaches and shaded/sheltered areas, and most importantly, the presence 

of water—all are attractive for recreation, so human use can be an explicit part of the 

restoration equation. 

But when we consider restoration of riparian ecosystems, it is easy to overlook its 

effects on recreation. The “recreation habitat” concept was introduced in Chapter 9 in this 

volume. While that chapter explored ways to minimize recreation impacts on riparian 

habitat through management, this chapter asks whether restoration can create recreation 

habitat as part of its design. To balance restoration goals with goals for recreation resources, 

we need to recognize the activities people are trying to pursue, the riparian characteristics 

they seek, and how restoration choices may impact recreation habitat.

This chapter discusses a typology of anthropogenic effects on riparian habitat, with 

one dimension considering whether impacts are intentional or unintentional, and another 

considering whether impacts are positive versus negative. We then discuss information 

from users about recreation habitat needs, preferences, and decision-making, using 

camping as an example. This helps evaluate outcomes from restoration efforts in terms of 

specific recreation resources. A final section provides other examples, using photos with 

extended captions. 

Descriptive and Evaluative Components

 Recreation use typically creates changes in biological and physical characteristics. For 

example, walking across a meadow may crush plants or compact soils, and these changes 

can be described objectively. But an evaluative dimension is also needed: What are the 

goals for recreation opportunities and ecological conditions, and are the specific biophysical 

changes acceptable within those frameworks? An 18-inch-wide band of lost vegetation 

and compacted soil may be acceptable if defined as a trail, while a wider band or multiple 

bands may not. Recreation management frameworks distinguish between these kinds of 

descriptive and evaluative information, encouraging resource managers to explicitly define 

the conditions they are trying to provide or avoid (Graefe et al. 1990; Manning 2001; 

National Park Service 1997; Shelby and Heberlein 1986; Stankey et al. 1985). 

 Applying such a framework to riparian changes, the descriptive component focuses 

on scientific relationships between restoration actions and different conditions (e.g., 

measurable changes in species diversity, erosion rates, or percent of bare ground), 
USDA Forest Service RMRS-GTR-411. 2020. 251

independent of whether those conditions meet goals or standards. An evaluative 

dimension (e.g., standards for “acceptable” species composition, channel dynamism, or 

amount of bare ground) is considered separately to assess whether a restoration project 

accomplishes specific ecological goals or end-states.

This concept seems simple, but in practice often it is not. Although ecologists may 

share implicit restoration goals (e.g., increase species diversity or geomorphic dynamism), 

there are often tradeoffs among outcomes, especially when those outcomes are tangible 

and specific (Schmidt et al. 1998). In Grand Canyon, for example, operational changes 

for Glen Canyon Dam could partially restore flood flow regimes that benefit some species 

(e.g., humpback chub and other native fishes, sub-tropical song birds), but adversely affect 

others (e.g., trout, the eagles that feed on them, and other birds such as the southwestern 

willow flycatcher). 

Clarifying Goals For Restoration

Aside from these specific examples, several different restoration endpoints are 

possible, some of which are difficult to define (Cole et al. 2008). A commonly expressed 

goal is to restore “natural” conditions, but increasing consensus in the scientific literature 

indicates such conditions are dynamic (Wu and Loucks 1995) and often conflated with 

current and/or historical human activities (Cole and Landres 1996). We can no longer 

assume ecosystems are clearly defined, balanced, or self-regulating, or that previous 

conditions are the most appropriate target for “active management” (Leopold et al. 1963). 

Cole et al. (2008) examine a range of possible restoration goals from an ecological 

perspective. They conclude that historical fidelity, biodiversity, species conservation, 

ecological integrity, and ecological resilience are distinct goals (albeit with some overlap) 

that may be more or less appropriate, realistic, or measurable in different restoration 

situations. 

Given this, projects to recover “damaged” or “degraded” ecosystems would benefit 

from explicit definitions of those terms, along with specific goals for improvements. 

Even the term “restoration” implies a value judgment that the existing condition is “bad” 

and the restored condition will be “better.” In some cases, restoration goals refer to 

some idealized equilibrium that may never have existed or cannot be replicated. Cole et 

al. (2008) argue that other terms such as “intervention” or “redirection” better describe 

intentional actions to alter an ecosystem’s path. Regardless of the terminology, these 

different end-states are human-defined and could include recreation habitat goals. 

Restoration can occur at different scales, ranging from specific sites to segments, the 

entire river, or the watershed. For example, restoration may include rehabilitating specific 

locations like camps or visitor attraction sites; improving the connectivity of meadows; 

altering dam-regulated flow regimes that have reduced flooding or sediment deposition; 

changing widespread grazing; altering trophic cascade effects from species removal; or 

introducing nonnative species such as tamarisk beetles to reduce an earlier invasive. 

This chapter adopts a more flexible definition of riparian areas than some of the 

technical delineations described in other chapters. On a desert river like the Colorado in 

Grand Canyon, for example, users may camp above the narrow riparian zone defined by 

dam-altered hydrologic regimes, often because riparian changes such as tamarisk invasion 

or beach erosion have constrained camp-able areas. In contrast, wetter environments may 
252 USDA Forest Service RMRS-GTR-411. 2020.

encourage camping close to the river on sand or gravel bars that are free of insects or 

vegetation. For recreation users, the riparian area may include sand beaches, gravel bars, 

rock cliffs, upland benches, and associated vegetation ranging from grasses to shrubs to 

large trees. People seek recreation habitats that support their activities, and they may not 

care about or even recognize formal ecological definitions.

Riparian Changes Due to Human Activity

Many riparian changes are caused by human activities, ranging from dams, grazing, 

or urbanization, to restoration projects designed to mitigate these kinds of stressors. 

The typology presented in figure 42 evaluates two orthogonal dimensions: whether 

the effect was intentional or unintentional, and whether it was positive or negative. We 

have included several phrases that might be used to describe actions in each quadrant, 

illustrating the range of ways scientists, practitioners, policy makers, or stakeholders 

account for different outcomes. 

The typology recognizes that some negative changes can occur through unintentional 

oversight or neglect, while other manipulations have well-understood negative 

consequences that were intentional but considered acceptable because of other benefits or 

tradeoffs. For example, good snail habitat on the dam face or in downstream fluctuation 

zones is a fortuitous unintentional effect of Glen Canyon Dam, just as shade and 

screening within campsites are unintentional positive effects of tamarisk introduction 

(although pruning may be required). One could discuss similar questions about different 

aspects of tamarisk beetle introduction. 

In all these cases, changes caused by humans (either positive or negative) are forms of 

“riparian engineering”; this term recognizes that humans are part of the equation and are 

responsible for many outcomes, regardless of intentions. As we work to restore riparian 

Negative Positive

Intentional

Unintentional

Stewardship,																																			

best	management	practice,

good	professional	management,

management	by	design

Trade-off,																	

sacrifice,												

collateral	damage

Oversight,	

ignorance,	

blunder

Fortuitous,	

lucky	

Figure 42—A typology of 

riparian effects due to 

human activity.
USDA Forest Service RMRS-GTR-411. 2020. 253

ecosystems from more negative states to more positive ones, we have the opportunity to 

consider, and perhaps intentionally improve, characteristics that benefit human activities 

like recreation. The result is recreation habitat “by design rather than by default” (Shelby 

and Shindler 1995; Shelby et al. 1992), with increased likelihood that outcomes will be 

intentional and positive rather than negative. 

Creating High Quality Recreation Habitat: The Campsite Example

Campsites are an important riparian feature, particularly for rivers with multiday trips. They are an excellent example of recreation habitat that can be affected by 

restoration projects on the orthogonal dimensions shown in figure 42. 

Several studies have examined campsite choice in a range of settings to determine 

characteristics that are important to recreation users. A review developed three categories 

of campsite attributes (Brunson and Shelby 1990): Necessity attributes such as flat 

ground, nearby water, and shelter/shade are minimum requirements. Experience attributes 

such as screening from other sites, being out of sight/sound of other groups, or good 

scenery are secondary, although they enhance the overall experience. Amenity attributes 

are a final category of less-important characteristics that may be tiebreakers in the choice 

process; examples include the amount of bare ground or presence/size of fire rings. 

 To examine the choice model, Brunson and Shelby (1990) collected data from 

floaters on Oregon’s Deschutes River, a desert canyon stream east of the Cascade 

Mountains. Summer flows are clear, cold, and fairly constant as a result of an upstream 

reservoir with power-generating and re-regulating dams. Camping most often occurs on a 

segment with several Class II-III rapids and a nationally known trout fishery that attracts 

highly specialized anglers. 

As hypothesized, necessity attributes (flat ground, good boat tie-up, and shade) were 

rated most important. Experience attributes (good fishing water, screened site, and out 

of sight/sound of other sites) were less important and associated with desired experience 

outcomes such as “getting away from others.” Amenity attributes (free of grazing, little bare 

ground, and distance from a railroad that runs along the river) were rated least important. 

The study also explored alterations in the campsite choice process due to: differences 

in outcome goals (e.g., anglers care more about fishing water than non-anglers); the need 

to choose quickly because of competition from other users, fatigue, or threatening weather; 

and lack of information about sites that lie ahead (because non-motorized boaters approach 

sites sequentially and cannot go back upstream). Under ideal conditions, users go through 

the longer three-stage choice process, but constrained conditions may cut the process short 

any time minimum thresholds are met (similar to Simon’s [1956] satisficing concept). 

Campsite attribute needs and preferences suggest several priorities for restorationrelated riparian engineering. Necessity attributes should be the first concern; the ability 

to access the site and tie-up, flat ground, and shade protection were most important to 

Deschutes River users. These attributes are clearly related to restoration options that affect 

the amount of beach, the type and size of vegetation, or availability of flat bench uplands.

 A 1980s restoration project on the Deschutes provides an interesting example. Camps 

in Sections 1 and 2 are scarce because the left bank of the river is closed to camping by 

the Warm Springs Tribes. Many of the right-bank sites lack shade, a necessity attribute 

during hot dry summers. To create more and better camps, agencies and conservation 
254 USDA Forest Service RMRS-GTR-411. 2020.

groups cooperatively planted trees along river-accessible upland benches flat enough for 

camping. But with a long dry summer, growing trees required water. The solution was 

placing 5 gallon buckets with interpretive signs explaining the need to water trees until 

they developed deep roots. Tree growth was successful at several sites, although the 

choice of some tree species (e.g., Russian olive) introduced exotics with other restoration 

consequences, and cottonwoods created widow-maker hazards as they matured (L. Ripley 

2016, personal communication). In short, there were some negative effects in spite of 

good intentions and a specific recreation habitat goal.

Riparian engineering may also affect experience attributes. Good fishing water was 

important for many Deschutes River floaters, and this is related to fish habitat goals for 

game and nongame species. Vegetation is another restoration variable that may provide 

screening within and between camp sites. Interestingly, amenity attributes such as bare 

ground and absence of grazing impacts were less important to Deschutes users, supporting 

the argument in our other chapter that recreation habitat may have dimensions that are 

different from biological habitat. 

A similar campsite selection study of boaters in Hells Canyon on the Oregon/Idaho 

border showed similar results, with additional attributes related to daily flow fluctuations 

from power generation releases at Hells Canyon Dam (Shelby et al. 2002). Although 

floaters and power-boaters had some different needs for parking boats, both groups were 

concerned about getting stranded if water levels dropped over short periods (a negative 

outcome from an intentional change in flow regime). 

Other Examples

To summarize, human-caused riparian restoration generally alters physical and 

biological ecosystems, probably changing characteristics that are important for recreation. 

This means that as we engineer physical or biological habitat, we are also engineering 

recreation habitat, either by default or by design. It makes sense to choose design, 

although we don’t always. Using photos with extended captions (figs. 43-58), the 

following section provides a variety of examples, ranging from site-specific to watershed 

and regional scales.
USDA Forest Service RMRS-GTR-411. 2020. 255

Figure 43—Granite Camp on the Colorado River (top) is a 

heavily-used “tamarisk camp,” where trees have been 

pruned over the years to provide shade and screened 

sleeping areas. The restoration project described by this 

onsite sign (center left) is one of the few places in Grand 

Canyon where trees have been extensively planted and 

cultivated (center right, and in top photo) to replace 

tamarisk dying from introduced beetles (some trees have 

been cut and removed, bottom left). But it will be some time 

before these new trees reach the maturity of “old growth” 

tamarisk or replace its recreation benefits in terms of shade, 

shelter, or screening.
256 USDA Forest Service RMRS-GTR-411. 2020.

Figure 44—These tamarisk at 

Carbon Creek camp on the 

Colorado River are the best 

available shade and shelter 

for this kitchen setup, even 

though they are blackened 

and dying. In contrast to 

the cultivation at Granite 

Camp, most Grand Canyon 

camps (like this one) have 

no “engineered” vegetation 

to replace beetle-killed 

tamarisk. New vegetation 

may regrow spontaneously 

but may not provide good 

recreation habitat. 

Figure 45—This Google Earth 

satellite photo shows the areas 

just above Phantom Ranch 

in Grand Canyon, with the 

Kaibab Trail and bridge crossing 

immediately downstream. There 

is high demand for Cremation 

Camp (on river left) because 

the 12-mile Inner Gorge area 

upstream has few camps per 

mile, and many parties want 

to camp near the Kaibab and 

Bright Angel Trails that provide 

access for passenger exchanges. 

In addition to resource 

protection goals, the site has 

been engineered to provide 

good recreation habitat for 

multiple groups. The “one party 

per camp” norm that exists in 

most of the Canyon is explicitly 

suspended here to increase 

capacity (insets show different 

landings that allow for campsharing).
USDA Forest Service RMRS-GTR-411. 2020. 257

Figure 46—This carefully engineered 

tent site at Crystal Camp 

in Grand Canyon provides 

excellent camping habitat. Data 

show visitors appreciate cues 

that identify where to camp 

and indicate that their activities 

are not causing unacceptable 

impacts. Such improvements 

protect the riparian area by 

keeping camps in the more 

resistant post-dam riparian zone 

and discouraging use in the 

more fragile native vegetation 

higher up the bank.

Figure 47—Funneling use to a 

single area that can then 

be hardened is an effective 

strategy to get visitors through 

sensitive riparian areas (Little 

Nankoweap camp on the 

Colorado River in Grand 

Canyon shown here). Top 

photos show steps from the 

boat landing to the camp area, 

soon after construction in 2005 

(top left) and needing some 

work in 2016 (top right). Lower 

photos show the trail from 

camp up to Little Nankoweap 

Canyon and the Nankoweap 

granary, some of it in good 

shape (bottom left) and some 

not (bottom right).
258 USDA Forest Service RMRS-GTR-411. 2020.

Figure 48—Even in remote 

locations like the bottom 

of the Grand Canyon, 

riparian restoration may 

need to accommodate large 

amounts of recreation use, 

as suggested by parked 

boats at Deer Creek (top), 

Havasu (bottom left), and 

Elves’ Chasm (bottom right). 

Almost all Grand Canyon 

trips stop at these popular 

attraction sites.

Figure 49—Good trail design provides hiking habitat by identifying routes that get people where they want to go (trail to upper 

Deer Creek in Grand Canyon shown here). At the same time, trails protect restoration projects by keeping vegetation loss and 

soil compaction at acceptable levels. 
USDA Forest Service RMRS-GTR-411. 2020. 259

Figure 50—Techniques for identifying trails, keeping people on them, and minimizing impacts include those shown here: border 

rocks (top left), marking routes with cairns (top right and bottom left), and steps in steeper segments (bottom right). Photos are 

from Nankoweap and Carbon Creek areas in Grand Canyon. 
260 USDA Forest Service RMRS-GTR-411. 2020.

Figure 51—This Google Earth satellite image of Nankoweap on the Colorado River in Grand Canyon shows where the Park 

Service developed and clearly marked a “best” trail for the popular hike to the Anasazi granary high on a cliff (lower inset), 

with a spectacular view down the Canyon (upper inset). Before this restoration, which included designating camps and day 

use areas, river runners created a web of dozens of trails.
USDA Forest Service RMRS-GTR-411. 2020. 261

Figure 52—Tamarisk has overtaken other riparian vegetation on many Southwestern rivers, creating dense stands that typically 

eliminate open space and flat ground needed for camping. Resource agencies have released tamarisk beetles to combat the 

species, a program claimed to be among the most carefully studied introductions ever. Although this is controversial, beetles 

have clearly reduced the amount of living tamarisk and probably its spread to other areas. But as shown in this photo from 

the Green River in Stillwater Canyon, impenetrable dead trunks and branches prevent recreation use and create fire hazards.
262 USDA Forest Service RMRS-GTR-411. 2020.

Figure 53—Piles of cut tamarisk (top photo) in Three Canyon camp at Trin Alcove Bend in the Green River’s Labyrinth Canyon. 

Despite the extensive restoration effort, remaining stumps and uneven ground (bottom left) interfere with good camping 

habitat, and dense willow may quickly take over the area (bottom right). Some tent spots have been cleared in nearby oak 

uplands (center) and future high water may deposit enough sand to reduce the problem.
USDA Forest Service RMRS-GTR-411. 2020. 263

Figure 54—A Park Service/Canyon Country Youth Corps crew removes thickets of beetle-killed tamarisk at the confluence of the 

Green and Colorado Rivers (top left), part of what one crew-member described as a career goal to rid the area of this invasive. 

Trees up to 12–inches in diameter were cut with chainsaws (bottom left) and thrown in the river (top right and center right). 

Downstream, driftwood with saw-cut ends float on the river (center and bottom right), piles up on shore, or eventually winds 

up in Lake Powell. It is not clear whether this “new” form of driftwood is considered good, bad, or just a tradeoff of clearing 

invasive tamarisk, but regulations in some jurisdictions prevent river runners from using slash for firewood. 
264 USDA Forest Service RMRS-GTR-411. 2020.

Figure 55—On a cold and stormy day at Spanish Bottom in Cataract Canyon below the Green/Colorado River confluence, even 

dead tamarisk with blackened trunks provide better camping habitat (inset) than a more aesthetically pleasing open beach 

(top photo) that was lashed by wind-driven sand and rain. Lower photo shows planted cottonwoods protected by staked wire 

cages; tamarisk stumps like those in the foreground remain after the trees have been cut and the slash removed. Although 

some cottonwoods were planted soon after beetle introduction, it will take many years before shade and shelter from these 

newer trees replace the “old growth” tamarisk (something this restoration project appears to recognize in its design).
USDA Forest Service RMRS-GTR-411. 2020. 265

Figure 56—A beach area recently protected from grazing on Oregon’s John Day River (top photo, taken at a higher spring flow) 

shows willows establishing along the lower summer water line: willows may eventually take over or cut off access to the 

beach, decreasing recreation habitat. Mitigation for such effects can be included in riparian engineering during restoration 

projects. Pioneering restoration research has been done in the John Day and Deschutes River watersheds east of the Oregon 

Cascades. This work showed how riparian vegetation in historically grazed locations (center left) was markedly different from 

areas protected from grazing by roads or railroads (center right). Cottonwood plantings (background of lower right photo) are 

part of the effort, but a pleasant surprise has been dramatic recovery after removing livestock or simply installing watering 

tanks (on trailers with solar-powered pumps) higher in the riparian zone.
266 USDA Forest Service RMRS-GTR-411. 2020.

Figure 57—This whitewater park in Salida, Colorado, is a large-scale river restoration that includes extensive channel 

modifications as well as bank improvements. Specifically designed to provide hydraulic features for whitewater boating 

(ColoradoKayak.com photo, bottom left), the park creates diverse recreation habitats on and off the water, including pathways 

(top left), seating areas (top right and bottom right), and a “downtown restoration” setting for shops and restaurants (out of 

sight upstream on river right). Ironically, on-shore areas provide for much larger numbers of recreation-days, compared to 

boating and other water activities.
USDA Forest Service RMRS-GTR-411. 2020. 267

Figure 58—Signs marking restoration areas take many forms: short vs. long, less vs. more informative, friendly vs. more 

demanding instructions, and messages that sometimes appear incongruous (like “not a trail” with an obvious trail behind 

it, bottom right). The detailed sign at the Salida Whitewater Park (top) conveys an interdisciplinary message about habitat 

benefits for boaters, aquatic life, and riparian vegetation.
268 USDA Forest Service RMRS-GTR-411. 2020.

References

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