Planning for Biodiversity: Bringing 

Research and Management Together

Proceedings of a Symposium for the South Coast Ecoregion

February 29 - March 2, 2000 Pomona, California

United States 

Department of 

Agriculture

Forest Service

Pacific Southwest 

Research Station

General Technical 

Report

PSW-GTR-195

December 2005
Abstract

Kus, Barbara E., and Beyers, Jan L., technical coordinators. 2005. Planning for 

Biodiversity: Bringing Research and Management Together. Gen. Tech. Rep. 

PSW-GTR-195. Albany, CA: Pacific Southwest Research Station, Forest Service, U.S. 

Department of Agriculture; 274 p.

Southern California, recognized as a major center of biodiversity, contains some of the most 

diverse habitats of any landscape in North America. The ever-expanding human population of the region desires land, water, resources, and recreation, creating conflict with the 

habitat requirements of many rare species. Managing resources in a way that maximizes 

biodiversity in remaining habitats, while providing opportunities for other appropriate uses 

of the land, presents a formidable challenge, requiring coordination between scientists and 

resource managers. The papers in this proceedings volume reflect the breadth of issues 

facing the science and management communities in southern California, ranging from the 

threats of fire, air pollution, grazing, exotic species invasion, and habitat loss on native habitats and sensitive species, including birds, mammals, reptiles and amphibians, to the role of 

mycorrhizal fungi as indicators of biological change. 

Retrieval Terms: air pollution impacts, biological diversity, California south coast ecoregion, 

fire management, habitat management, rare species

Technical Coordinators

Barbara E. Kus is Research Ecologist with the U.S. Geological Survey Biological Resources Division, Western Ecological Research Center, San Diego Field Station, 4165 Spruance Road, Suite 200, San Diego, California 92101. 

E-mail: Barbara_kus@usgs.gov

Jan L. Beyers is Plant Ecologist with the USDA Forest Service, Pacific Southwest Research 

Station at the Forest Fire Laboratory, 4955 Canyon Crest Drive, Riverside, CA 92507.

E-mail: jbeyers@fs.fed.us
Planning for Biodiversity: Bringing 

Research and Management Together

Proceedings of a Symposium for the South Coast Ecoregion

Barbara E. Kus and Jan L. Beyers

Technical Coordinators

February 29 – March 2, 2000

Pomona, California
Disclaimer

Papers were provided by the authors/technical coordinators in camera-ready form for printing. 

Authors are responsible for the content and accuracy. The use of trade or firm names in this 

publication is for reader information only and does not imply endorsement by the U.S. Department 

of Agriculture of any product or service. Opinions expressed may not necessarily reflect the 

position of the U.S. Department of Agriculture.
i

Planning for Biodiversity: Bringing Research and Management Together

Contents

iii Introduction

iv Acknowledgement

 Invited Presentations

1 Assessing Estuarine Biota in Southern California

	 	 	 Kevin D. Lafferty

17 Up, Down, or Stable: Populations of Endangered Birds in Beach and Estuarine Areas 

 in Southern California

	 	 	 Abby N. Powell

29 Considerations for the Management of Vernal Pool Faunal Communities

	 	 	 Marie A. Simovich

43 Biodiversity of Mycorrhizal Fungi in Southern California 

	 	 	 Michael F. Allen, Louise Egerton-Warburton, Kathleen K. Treseder, Cara Cario,

	 	 	 Amy Lindahl, Jennifer Lansing, José Ignacio Querejeta, Ola Karen, Sharon 	 	

	 	 	 Harney, and Thomas Zink

57 A New Look at Coastal Sage Scrub; What 70-year-old VTM Plot Data Tell us about 

 Southern California Shrublands 

	 	 	 Robert S. Taylor, Jr.

79 Air Pollution and Vegetation Change in Southern California Coastal Sage Scrub: A 

Comparison with Chaparral and Coniferous Forest 

	 	 	 Edith B. Allen, Abby G. Sirulnik, Louise Egerton-Warburton, Sheila N. Kee, 	 	 	 	 	 Andrzej Bytnerowicz, Pamela E. Padgett, Patrick J. Temple, Mark E. Fenn, Mark

	 	 	 A. Poth, and Thomas Meixner

97 Fire as a Threat to Biodiversity in Fire-Type Shrublands 

	 	 	 Jon E. Keeley

107 Recovery Research for the Endangered Pacific Pocket Mouse: An Overview of 

Collaborative Studies 

	 	 	 Wayne D. Spencer

127 Herbaceous Responses to Livestock Grazing in Californian Oak Woodlands: A 

Review for Habitat Improvement and Conservation Potential 

	 	 	 Barbara Allen-Diaz and Randall D. Jackson

145 Air Pollution Impacts in the Mixed Conifer Forests of Southern California 

	 	 	 Patrick J. Temple, Andrzej Bytnerowicz, Mark E. Fenn, and Mark A. Poth

165 Sensitive Species of Snakes, Frogs, and Salamanders in Southern California Conifer 

Forest Areas: Status and Management 

	 	 	 Glenn R. Stewart, Mark R. Jennings, and Robert H. Goodman, Jr.

199 The Spotted Owl in Southern California: Ecology and Special Concerns for 

Maintaining a Forest-Dwelling Species in a Human-Dominated Desert Landscape 

	 	 	 William S. LaHaye and R. J. Gutiérrez
ii

GENERAL TECHNICAL REPORT PSW-GTR-195

211 Bats in the South Coast Ecoregion: Status, Conservation Issues, and Research Needs 

	 	 	 Karen L. Miner and Drew C. Stokes

229 The Santa Margarita River Arundo donax Control Project: Development of Methods 

 and Plant Community Response 

	 	 	 Dawn M. Lawson, Jesse A. Giessow, and Jason H. Giessow

 Extended Poster Abstracts

245 Diptera Community Composition and Succession following Habitat Disturbance by 

Wildfire 

	 	 	 Michael A. Patten, Jutta C. Burger, Thomas R. Prentice, John T. Rotenberry,

	 	 	 and Richard A. Redak

249 Patterns and Processes of Arthropod Community Succession after a Fire 

	 	 	 Jutta C. Burger, Michael A. Patten, John T. Rotenberry, and Richard A. Redak

253 Monitoring the Effects of Natural and Anthropogenic Habitat Disturbance on the 

Ecology and Behavior of the San Diego Coast Horned Lizard, (Phrynosoma 	 	

	 	 	 coronatum blainvillei)

	 	 	 Tandora D. Grant and Allison C. Alberts

257 Fire Management in Some California Ecosystems: A Cautionary Note 

	 	 	 Hartmut S. Walter, Teresa Brennan, and Christian Albrecht

261 Spatial and Temporal Variation in Ephemeral Pool Crustacean Communities 

	 	 	 Janette Holtz, Marie A. Simovich, and Thomas Philippi

265 Habitat Ephemerality and Hatching Fractions of a Diapausing Anostracan 

(Crustacea: Branchiopoda)

	 	 	 Marie A. Simovich, Thomas E. Philippi, Ellen T. Bauder, and Jacob A. Moorad

267 Bee Diversity Associated with Limnanthes Floral Patches in California Vernal Pool 

 Habitats 

	 	 	 Joan M. Leong and Robbin W. Thorp

269 Atmospheric Nitrogen Deposition and Habitat Alteration in Terrestrial and Aquatic 

Ecosystems in Southern California: Implications for Threatened and Endangered 

 Species

	 	 	 Mark Fenn, Mark Poth, and Thomas Meixner

273 Surveys for California Red-Legged Frog and Arroyo Toad on the Los Padres National 

Forest 

	 	 	 Valerie K. Hubbartt and Thomas G. Murphey
iii

Planning for Biodiversity: Bringing Research and Management Together

Introduction

Southern California contains some of the most diverse habitats of any landscape in 

North America and is recognized as a major center of biodiversity. It is also home 

to an ever-expanding human population, with needs for land, water, resources, and 

recreation that conflict directly with the habitat requirements of many species. As a 

result, this region has more endangered and threatened species than any other area 

in the continental United States, and once-extensive natural communities, many of 

which are unique to the ecoregion, have been reduced to mere remnants.

Managing resources in a way that maximizes biodiversity in remaining habitats 

while providing opportunities for other appropriate uses of the land presents a formidable challenge requiring coordination between scientists and resource managers, yet 

few opportunities exist for regular exchange of information and ideas between these 

groups. In response to this need, the U.S. Geological Survey and the USDA Forest 

Service partnered to co-host a symposium on “Planning for Biodiversity: Bringing 

Research and Management Together,” held February 29–March 2, 2000 at the Kellogg West Conference Center, California State Polytechnic University in Pomona, 

California. The goal of the 3-day conference was to identify the current status of our 

knowledge and gaps in our understanding of regional biodiversity and ecosystem 

processes, present and future threats to species and habitats, and effective monitoring strategies for southwestern and central coastal California resources. Through a 

program of 52 invited presentations, 18 contributed posters, and 10 focused discussion groups, the conference created an environment for formal and informal communication among the 300 attendees about the results of scientific studies and their 

application to resource conservation and management, as well as the information 

needs of managers responsible for determining and implementing management on 

the ground.

Of the 45 technical papers presented at the conference, 14 are included in this 

volume. Authors were asked to synthesize the current state of knowledge regarding 

their topic and identify areas needing future research. Each paper was assigned to 

an editor for review and received one to three additional peer reviews. Expanded 

abstracts of nine posters also were reviewed by the editors and included. The topics addressed in the papers and poster abstracts reflect the breadth of the conference 

presentations and the issues facing the science and management communities, ranging from the threats of fire, air pollution, grazing, exotic species invasion, and habitat 

loss on native habitats and sensitive species of birds, mammals, reptiles and amphibians, to the role of mycorrhizal fungi as indicators of biological change.
iv

GENERAL TECHNICAL REPORT PSW-GTR-195

An outgrowth of the symposium was the recognition that periodic conferences, while important for facilitating information exchange, are by themselves 

not enough to maintain the ongoing dialog among conservationists needed to effectively apply resources to accomplish mutual goals. As a result, the “San Diego 

Partners in Biodiversity” was formed, a network of scientists, field biologists, 

natural resource managers, and others working toward conservation of biodiversity. Building upon the relationships established at the conference, this group now 

meets monthly to inform and educate each other about activities, opportunities, 

and needs relating to resource conservation. It is our hope that this type of coordination and collaboration continues to expand and that this volume contributes to 

that spirit.

ACKNOWLEDGEMENTS

The work and financial support of many individuals and organizations made the 

conference and resulting volume possible. We are grateful to our cosponsors for 

financial contributions: California State Parks, The Irvine Company, The Nature 

Conservancy, San Diego State University Field Stations, LSA Associates, Inc., and 

Southern California Edison. The Kellogg West Conference Center provided an excellent venue for the event. Jan Beyers and Patrick Temple, USDA Forest Service, 

and Robert Fisher and Barbara Kus, U.S. Geological Survey, served on the Steering Committee, assisted by a Program Committee including Michael Arbaugh and 

Steve Anderson, USDA Forest Service; Wendy Picht, Metropolitan Water District 

of Southern California; Mark Reynolds, San Diego State University Field Stations; Ray Sauvajot, National Park Service; Trish Smith, The Nature Conservancy; 

and John Stephenson, U.S. Fish and Wildlife Service. Penelope Randall served as 

Symposium Coordinator, ably turning vision into reality. 

The conference would not have been a success without the participation of 

numerous individuals who volunteered their time to serve as session chairs and 

dis-cussion group leaders, run the registration desk, provide audio-visual support, 

and assist with other logistical activities: Adam Backlin, Cameron Barrows, 

Alan Bate, Kathy Baumberger, Shawna Bautista, Melanie Beck, Peter Beck, 

Ann Berkeley, Anne Bradley, Chris Brown, Karen Carter, Gina Castro, Bonnie 

Corcoran, Gary Davis, Edward Ervin, Andrew Fisher, Liz Gallegos, Tony Gomez, 

Karen Gonzalez, Chris Haas, Stacie Hathaway, Anita Herring, Geary Hund, 

Mark Jennings, Boulderson Lau, Kevin Lafferty, Jaqueline Leonard, Steve Loe, 

Jeffrey Lovich, Rob Lovich, Claudia Luke, Lisa Lyren, Melanie Madden, Judith 

Mitchell, Jud Monroe, Marcelo Pires, Therese O’Rourke, Molly Pohl, John Rieger, 
v

Planning for Biodiversity: Bringing Research and Management Together

Carlton Rochester, Tom Scott, Drew Stokes, Andrew Suarez, Ted St. John, Jennifer 

Turnbull, and Dustin Wood.

We tappreciate the efforts of and thank the following referees who generously 

shared their time and expertise in reviewing manuscripts for this volume: Edith 

Allen, Richard Ambrose, Michael Arbaugh, Ellen Bauder, Denton Belk, Andrew 

Bohonak, Patricia Brown, Mark Capelli, Jeanne Chambers, Art Chappelka, 

Norman Christensen Jr., Denny Constantine, Frank Davis, Sandra DeSimone, Janet 

Franklin, Catherine Gehring, Gordon Gould, Nancy Grulke, Robert Hansen, Dan 

Holland, Ron Jurek, Barbara Massey, Leroy McClenaghan, Paul Miller, Richard 

Minnich, Gary Page, James Patton, David Peterson, Elizabeth Pierson, David 

Pritchett, John Robinson, Paula Schiffman, Norman Scott, Darlene Southworth, 

John Stephenson, Brent Takemoto, Antonia Wijte, and Paul Zedler.

Barbara E. Kus Jan L. Beyers

U.S. Geological Survey USDA Forest Service

Western Ecological Research Center Pacific Southwest Research Station

San Diego Field Station Forest Fire Laboratory

San Diego, CA 92101 Riverside, CA 92507

Assessing Estuarine Biota in Southern 

California1

Kevin D. Lafferty2

Abstract 

In southern California, most estuarine wetlands are gone, and what little habitat remains is 

degraded. For this reason, it is often of interest to assess the condition of estuaries over time, 

such as when determining the success of a restoration project. To identify impacts or 

opportunities for restoration, we also may want to know how a particular estuary, or area 

within an estuary, compares with neighboring areas. Comparisons among wetlands require

knowledge of different estuary types. The seven types of estuaries described in this paper can

be easily grouped into two functional types, fully tidal and seasonally tidal, based on a simple

biotic index: presence of horn snails. A description of the distribution, diversity, and

abundance of organisms in estuaries is one way to assess resources, determine the success of a

habitat restoration, and compare estuaries to evaluate the biotic consequences of degradation.

In this review, I summarize techniques that may be useful for managers charged with biotic

inventory and monitoring, emphasizing techniques to categorize wetlands and quantify plants, 

invertebrates, fishes, birds, and trematode parasites. 

Key words: bird, estuary, fish, indicator, invertebrate, trematode

Introduction 

Although California has a Mediterranean climate, it contains many types of wetlands. 

Along the coast, estuaries are the most familiar. In this paper, I first describe 

estuaries, provide examples of the functions that they provide, and discuss impacts,

losses, and restoration. I then describe methods for characterizing and comparing the 

biota of southern California’s estuaries. 

Cowardin and others (1979) define estuaries as subtidal habitats and their

adjacent tidal wetlands that are usually partially enclosed by land but have open, 

partly obstructed, or sporadic access to the open ocean and in which ocean water is at

least occasionally diluted by freshwater runoff from the land. A combination of

ample nutrients and sunlight make estuaries productive, though easily perturbed, 

habitats. Often, species found in estuaries are specially adapted to and, therefore, 

dependent on estuaries. From an ecosystem perspective, estuaries provide the

following functions: (1) food chain support and nutrient cycling functions such as 

primary production, decomposition, nutrient export, and nutrient utilization; (2) 

habitat functions for native plants, invertebrates, fish, birds, mammals, and 

herpetofauna; (3) hydrological functions including flood conveyance, sediment 

control, ground water recharge and discharge, and shoreline protection; and, (4) 

water quality functions including water supply, waste-water treatment, detoxification 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 Research Marine Ecologist. Western Ecological Research Center, USGS. c/o Marine Science Institute.

University of California, Santa Barbara, CA, 93106 (email: klafferty@usgs.gov) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 1 
Assessing Estuaries—Lafferty

of toxic substances, and modification of pollution from nutrient enrichment (Ferren

and others 1995). 

Why should people care about estuaries? To some, they are ugly swamps, unproductive lands, or mosquito breeding areas. Through most of history, governments 

around the world actively worked to eliminate estuarine habitats or convert them to

more “productive” uses. A more modern view considers that estuaries, in addition to 

the ecosystem functions described above, provide the following socio-economic

functions: (1) non-consumptive values such as recreation, research, and education;

(2) consumptive values of “harvestable” physical resources such as water, gravel, and 

petrochemicals, and biotic resources such as fisheries, lumber, and crops; (3) cultural 

values by Native Americans, including activities such as basketry, gathering of food 

and medicinal plants, and acquisition of house building materials; and, (4) aesthetic 

and natural heritage values including natural landscape features, setting, and other

aspects of California’s natural heritage (Ferren and others 1995). 

One reason that wetlands are an area of concern is that we have lost ~90 percent 

of the original coastal wetland habitat in southern California as a result of filling or 

dredging during the last century (NOAA 1990). What is left has been degraded 

through fragmentation, water quality degradation, introduction of invasive plants and 

animals, unregulated public access, and other forms of environmental perturbation. In

addition, damming, agriculture, and urbanization have greatly altered the magnitude 

and timing of freshwater inputs. Even in a degraded state, however, estuaries continue to provide important ecosystem functions. Of these functions, perhaps that of

habitat for rare and endangered species is most important because of the many plants

and animals restricted to the remaining estuary habitat (Zedler and others 1990).

Ferren and others (1995) group anthropogenic impacts to Californian estuarine

ecosystems as: (1) agricultural development (for example, deforestation, draining,

increased sedimentation, fragmentation, nutrient enrichment, pesticide pollution); (2) 

urbanization (for example, filling for development, landfill, excavation, 

fragmentation, diking, impoundment, runoff, effluent pollution, nutrient enrichment, 

and sedimentation); (3) agency activities (for example, channelization, flood control, 

ditching, and chemical treatment for mosquito abatement); (4) resource extraction 

(for example, petrochemicals, sediments, water, salts, fisheries); (5) public access

(for example, compaction of soil, removal and disturbance of organisms); and (6)

introduction of invasive exotic species. 

The sad state of our coastal wetlands has led to increased efforts to enhance, 

restore, create, and recreate estuarine habitat. Such attempts are either due to 

conservation actions or as required mitigation for the destruction of wetland habitat 

elsewhere. Throughout southern California, many estuaries have been subject to

habitat restoration. Unfortunately, studies of the effectiveness of estuarine restoration 

have been sobering (Zedler 1984, Zedler and Langis 1991). At best, most restoration 

has converted one type of wetland to another, often arguably resulting in an overall 

net loss in wetland function (Zedler and Langis 1991). Thus, mitigation as a 

technique to compensate for losses or impacts should presently be viewed with 

concern, and the ability to assess success becomes critical to wetland conservation. 

2 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Assessing Estuaries—Lafferty

Assessment 

There are many reasons to assess estuarine functions. In some cases, we may wish to 

evaluate the success of a restoration project. We also may try to determine the extent 

to which a particular wetland is degraded or to obtain baseline information to be able

to detect changes in function in the future. If the assessment is related to a specific

function, its goals may be quite narrow. Such assessments might attempt to determine

the presence of an endangered or fisheries species, the ability of a wetland to hold 

storm runoff, the quality of water exiting an estuary into the littoral zone, or the 

presence of mosquitoes. The habitat role of estuarine wetlands may be the most 

comprehensive function of all and the most difficult to assess. The following sections

provide ideas for sampling plants and animals that use the estuary habitat. 

For any assessment, it is important to consider that we can categorize southern 

California’s estuaries along many different descriptive axes. Because the biotic 

community that occurs in an estuary may be influenced by the estuary’s type as much 

as by the quality of the habitat, Ferren and others (1996) provide the following 

hierarchical categorization scheme for the seven distinct types of estuaries that occur

in southern California: 

River Mouth Estuaries: These estuaries occur at the mouths of large rivers that 

generally have some perennial flow, producing nearly permanently brackish

conditions in the estuary. The Santa Clara River Estuary, the Ventura River Estuary 

and the Tijuana River Estuary are representative. 

Canyon Mouth Estuaries: Emergent portions of the southern California coastline

have canyons that drain watersheds. Creeks flowing through these canyons empty 

into the ocean through small estuaries that vary greatly in size, frequency of tidal

flushing, salinity regimes and biota. Arroyo Burro and Devereux Slough in Santa 

Barbara County are examples of canyon mouth estuaries. 

Lagoonal Estuaries: These estuaries have mouths closed by sand bars most of

the year, usually have brackish fringe-marshes rather than vegetated flats, often have 

salinities that can approach fresh water, and support fauna typical of brackish to fresh 

water conditions (this is especially true where freshwater input has been enhanced by

wastewater discharges and agricultural and urban runoff). Although many estuaries in

southern California have lagoons at their mouths, lagoonal estuaries, such as Buena 

Vista Lagoon, primarily occur in San Diego County. 

Coastal Dune-Creek Estuaries: These estuaries develop when wetlands near

dunes occasionally connect with the ocean but have a seasonal sand bar. They are

generally slightly brackish most of the year. Examples include the San Antonio Creek

Estuary and Shuman Creek Estuary in Santa Barbara County. Dune encroachment

into the wetlands keeps these wetlands small. 

Bay Estuaries: These estuaries have large areas of subtidal habitat (bays) and

low elevation salt marsh. Bolsa Chica, Anaheim Bay, Upper Newport Bay and 

Mission Bay are examples. There is a strong marine influence in bay estuaries

because there is a consistently open and generally wide mouth. Intertidal mud flats,

cord grass, and eelgrass characterize bay estuaries (Zedler 1982, Zedler and others 

1992). 

Structural Basin Estuaries: Tectonic activity that leads to down-faulted and 

down-folded geologic structures below sea-level may create estuaries of moderate

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 3 
Assessing Estuaries—Lafferty

size that are most common in Santa Barbara County. Down-faulted basins include 

Goleta Slough, and down-folded (synclinal) basins include Carpinteria Salt Marsh. 

Artificial Drain Estuaries: Agricultural and urban drains along the coast can

form small, narrow estuaries when they reach the coast. They may be all that remains

of a historically larger estuary or they may be new creations. Some of these exist in 

the form of small wetlands on sandy beaches at the mouths of culverts and storm 

drains, such as along Cabrillo Blvd. in downtown Santa Barbara. Others reach the

coast as canals and have tidal channels with fringe marshes. Agricultural or urban

runoff often pollutes these small estuaries. 

An estuary is not a homogeneous habitat; each has a unique composition of

landforms characterized by a specific origin, geomorphic setting, water source, and

water regime (Ferren and others 1996). Associated with the large water body that 

defines the estuary may be channels, ditches, and ponds. Seeps and springs may

provide freshwater. The water bodies may cover bottoms, beds, bars, and reefs (such 

as may be formed by oysters and mussels). Affiliated with the margins of water

bodies may be shores, beaches, and banks. Tidal (mud) flats and deltas can be quite

extensive in West Coast estuaries. They may form with the accumulation of 

sediments and are exposed by the retreating tide. Although natural hard substrate is 

uncommon in estuaries, woody and shelly debris and artificial structures such as

pilings, levees, docks, and buoys can provide habitat for attached invertebrates and 

algae (Ferren and others 1996). The landforms an estuary possesses may affect the 

sorts of animals and plants that the estuary supports. In addition, because community

composition may vary from landform to landform, characterizing landforms at each

sampling site within an estuary will facilitate comparing communities among sites. 

Another important factor associated with habitat function is how salinity varies 

in space and time. Two physical features influence salinity in an estuary: (1) the 

amount and seasonality of freshwater input and (2) the presence of a bar that can

close off the mouth of the estuary from the ocean. As discussed later, invertebrates

seem to be the biota most sensitive to these factors. Bay and structural-basin estuaries 

tend to have year-round tidal influence and relatively little freshwater input. River 

mouth estuaries have perennial freshwater inputs and relatively less tidal influence.

Canyon mouth, lagoonal, coastal dune-creek, and artificial drain estuaries have seasonal freshwater inputs and seasonal bars that isolate them from the sea. These latter 

estuaries have periods of low salinity, high salinity, and tidal flushing that vary with 

the season. Their water levels also can fluctuate well beyond the normal tidal range. 

A final physical aspect of wetlands to consider in an assessment is size.

Although the technical boundaries of an estuary are based on salinity, it is more

practical to use the plant community to indicate the extent of the estuary. The main

reason for measuring wetland size is that it can act as a useful covariate when 

comparing assessments among wetlands. For example, the general ecological 

association between species richness and the size of a habitat sampled predicts that 

large estuaries should have higher species diversity by means of their size, 

independent of their quality of habitat. If this ecological rule holds for estuaries, large 

estuaries, all else being equal, generally could appear to be in better shape than small 

ones. For example, to determine the extent to which different estuaries are degraded, 

one could plot a measure of the estuary, such as species richness of a particular set of 

organisms, against estuary size. The prediction is that richness should increase with

size, but that more degraded estuaries should fall below the line that describes the 

size-richness curve. Preliminary efforts show that the richness of both fishes and 

4 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Assessing Estuaries—Lafferty

birds increases with wetland size but that estuaries that have few fish species for their 

size do not have few bird species for their size, suggesting that these two biota

respond independently to environmental factors. I have also found it useful to

separately measure the flooded, unvegetated area of the marsh (the sum of pans, 

channels, and mudflats) because fishes and birds seem to respond more to flooded

area than wetland size per se. Vegetated marsh and flooded habitats can be mapped

from aerial photographs. 

Sampling the Biota 

An enjoyable way to understand an estuarine community is to put on waders, grab

sampling gear, and count the plants and animals that live there. Plants, invertebrates, 

fishes, and birds all require different sampling methods. The biggest challenge when 

sampling wetlands is to obtain data that are comparable over time or across estuaries.

This means that substantial thought must be given to standardized timing, habitat,

methods, and efforts. An excellent primer on sampling estuaries is from the Pacific 

Estuarine Research Laboratory (PERL 1990), and interested readers should consult 

this reference. 

Sampling strategies will vary according to the types of questions being asked. 

Here I briefly describe the basic rationale behind the most common types of sampling

in wetlands. Central to sampling design is the selection of sampling sites. It is useful 

to avoid the temptation of simply sampling adjacent to access points. In most cases, it

will be necessary to stratify sampling by habitat type (for example, vegetated marsh, 

pans, narrow channels, and mudflats adjacent to the shores of large channels) so that 

comparisons among sites are not confounded by differences in habitat type. 

It is also important to account for gradients that may exist within habitat types. 

For example, when targeting a single tidal channel, we often stratify sampling so that 

our sites are spaced along the upstream-downstream gradient (mouth, upper extent of

tidal influence, and one or more points in between). Within a channel site, we may

also stratify sampling according to tidal height. 

Site selection will obviously be affected by the goals of the study. If the goal is 

to follow some aspect of the community over time, sites can be targeted that are

home to the species or community of interest. In this case, it is usually permissible to 

choose sites that can be conveniently revisited for monitoring. Another common goal 

is to assess a distinct location (for example, an impact or restoration site) and 

compare it to similar sites (often these comparisons are made over time). In this case,

the location of the target area is often predetermined, and comparison sites are chosen 

in similar habitat either along a gradient of impact or free from the variable of 

interest. If the goal is to assess the distribution and abundance of organisms in an 

entire wetland, it is necessary to randomly select multiple sample sites. For this, we 

have used an aerial photograph of a marsh, and divided it into discrete habitat types. 

From each of these types, we randomly select a subset of sampling locations by using

GIS or a grid system (we employ a rule so that sites are no closer than 50 m). 

The remainder of this section gives a rapid introduction to some of the sampling

methods that I have found most useful. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 5 
Assessing Estuaries—Lafferty

Plants 

Estuaries often support thick vegetation that forms marshes and provides important 

habitat for animals (see Zedler 1982). Because plant assemblages are often zoned

according to tolerance to inundation and soil salinity, these two characteristics can be 

useful estuarine indicators. Cord grass (Spartina foliosa) and pickleweed (Salicornia

virginica) often dominate the lower salt marsh. The middle salt marsh is generally 

irregularly flooded and has varying salinities. Pickleweed usually dominates here in 

association with other species such as salt grass (Distichlis spicata), alkali heath 

(Frankenia spp.), jaumea (Jaumea carnosa), and sea lavender (Limonium 

californicum). Parish's glasswort (Salicornia subterminalis), shore grass 

(Monanthochloe littoralis), and pickleweed (Salicornia spp.) dominate the high

marsh with its irregular flooding and euryhaline or hyperhaline salinity. Cattails 

(Typha spp.) and rushes (Scirpus spp., Juncus spp., and so forth) occur where salinity

is persistently low. Zedler and others (1999) argue convincingly that the elevationbased terms (“low,” “middle,” and “high” marsh) should be replaced by a system that

also considers landscape position and conspicuous species. They suggest three 

alternative habitat designations: (1) the high marsh—a 30- to 70-cm elevation range 

with Salicornia subterminalis; (2) the marsh plain—a 30-cm elevation range with

heterogeneous topography and a mix of several species; and (3) cord grass habitat—

the bayward portion of the marsh plain and lower elevations, all occupied by

Spartina foliosa. In addition, vegetation adjacent to channel edges tends to be distinct 

from vegetation away from edges.

Emergent plants are food for a number of terrestrial arthropods, several of which

are salt-marsh specialists. The contribution of salt-marsh plants to the diet of marine 

animals, however, is thought to be primarily through detritus. Fully aquatic 

vegetation is less apparent in estuaries but includes flowering plants such as widgeon 

grass (Ruppia spp.), eelgrass (Zostera spp.), and algae such as epipelic diatoms and 

the mat-forming filamentous alga, Enteromorpha spp. 

Sampling estuarine plants can be relatively straightforward because plant

species lists are not long, the topography is simple, and canopy heights are short 

enough that percent cover estimates are usually adequate. For example, in our field

work, we run a series of 10-m transects in which we make point-contact assessments

of the plant species composition every 10 cm and canopy height estimates every 

meter. Good aerial photographs are often available of southern California estuaries, 

and one can usually discern large patches of different plant communities. However,

one almost always needs to go to the field to ground-truth these maps and determine 

the composition of vegetation mosaics. There are good field guides and keys to salt 

marsh plants (Abbott and Hollenberg 1976, Faber 1996, Jepson 1993, 

http://elib.cs.berkeley.edu/photos/flora/). Probably the single best visual indicator of 

an estuary in southern California is the ubiquitous and easy to identify pickleweed. It

is difficult to find an estuary in southern California that lacks this plant. 

Epipelic micro-algae (for example, diatoms) and mat forming macro-algae are

important components of estuarine food webs. Micro-algae can be assessed 

microscopically or through the chlorophyll content of the sediment. Macro-algae can

be quantified in transects and quadrats (like those used to quantify invertebrates; see

the next section). Some macro-algal flats are extensive enough to be quantified 

through aerial photographs. The nutrient content of algae can be used as an indicator 

of nutrient enrichment in an estuary (Fong and others 1998). 

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Invertebrates 

Estuarine invertebrates live in the water, on vegetation, on the mud, and in the mud.

Most species reach peak abundance in the summer. They are relatively rare in winter 

after high freshwater flows. One can visually survey for crabs and snails at low tide.

Some invertebrates, particularly crustaceans, will come to baited minnow traps, while 

some small invertebrates will seek out litter traps (see PERL 1990). Many 

invertebrates are out in the open or give away their presence with burrows of

distinctive shapes (burrows can be difficult for novices to identify with any 

certainty). Estimating density requires setting a sampling unit, usually in the form of 

a quadrat, transect, or core. A factor that complicates density estimates is that most

animals are neither uniformly nor randomly distributed across a mudflat or tidal 

channel. Since many species are patchy, large numbers of sampling sites are

necessary to estimate density. In addition, many species are zoned according to tidal 

height, meaning that tidal height should be accounted for when making comparisons 

among sites. One method to sample invertebrates in a tidal channel is to extend band

transects from the vegetated margin of the salt marsh down into the intertidal zone,

noting the change in elevation and density of organisms at regular intervals of the 

transect. 

Sediment cores are useful for sampling shallow-burrowing species but often 

miss the deeper-dwelling clams, shrimp, and worms. Labor increases with the length 

and diameter of the core taken and decreases with the size of the screen the sediment

is passed through. To accommodate this difference, we take large diameter (for

example, 15 cm), deep cores (for example, 40-50 mm in depth) which are passed 

through a coarse mesh (for example, 3-5 mm) in the field and small diameter (for 

example, 5 cm), shallow (5-10 cm deep) cores that are preserved and then passed

through a fine mesh (1-0.5 mm) in the laboratory. Large cores can be sieved in the 

field. In cases where many cores are taken or the site is far from water, we find that a

water pump running off a small generator greatly speeds the task. Once captured, 

most of the larger marine invertebrates are relatively easy to identify to species.

Unfortunately, there are many small species that can be quite a challenge to identify 

even using the several good keys available (for example, Smith and Carlton 1975,

Morris and others 1980). 

Because coring is labor intensive, it is usually necessary to limit the number of

cores taken. Randomly selecting coring locations provides the best estimate of 

densities, but, due to the patchy distribution of animals, a large number of cores is 

needed to obtain a representative sample of even the most numerically dominant 

species. Patchiness is accentuated if there is significant variation in tidal height 

within the sampling area. Stratifying coring effort by tidal height helps to distribute 

effort across zones. Nonetheless, variation in the depth profile from site to site limits

comparisons among sites. More recently, we have experimented with mapping the

presence of beds of benthic infauna (as indicated by clam, crab, and shrimp burrows, 

and so forth) within transects and then randomly subsampling these beds with cores. 

Such targeted sampling intensifies effort in the areas where animals are most likely to 

occur, thereby increasing the efficiency of coring. 

In the subtidal areas of larger estuaries, coring is relatively difficult to 

accomplish and not frequently undertaken. Using SCUBA (self-contained underwater

breathing apparatus), divers can obtain cores similar to those obtained in the intertidal

zone. However, in most estuaries, diving conditions can be challenging due to low 

visibility and strong tidal flux. Alternatively, sediment samples can be taken by grabs 

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Assessing Estuaries—Lafferty

deployed off boats. Unfortunately, these usually only capture animals on the surface

or in relatively shallow parts of the sediment. 

One component of the estuarine invertebrate community that is often 

overlooked, particularly by those with training in marine biology, is the insect fauna.

Aquatic insects (many of which are larval forms) can dominate the invertebrate

community of some estuaries. Terrestrial insects are abundant in the warm months in 

and around estuaries and may play important, unappreciated roles in estuarine 

ecology. There are several insects, such as the pygmy blue butterfly caterpillar

(Brephidium exilis), that feed specifically on estuarine plants. 

Although assessing invertebrate communities can be very difficult (one needs to 

consider the resources available before embarking on an ambitious sampling

program), this does not mean that one should ignore invertebrates when assessing

wetlands. In particular, the invertebrate community can be a useful indicator of the 

type of tidal inundation that an estuary receives. As mentioned earlier, invertebrates 

are particularly sensitive to variations in salinity (especially compared with fishes and 

birds). Sandoval and Lafferty (1995) found that the invertebrate community of

estuaries with regular tidal influence is dominated by relatively marine species such 

as crabs, shrimp, polychaete worms, clams, mussels, and horn snails. In estuaries

with variable salinity, these species are usually absent. Instead, aquatic insects, 

amphipods, isopods, crayfish, small snails, and oligichaete worms can be abundant.

One species, the California horn snail (Cerithidea californica), is abundant and is 

easy to see and identify because it often carpets the mudflats and channel banks

where it occurs (Lafferty 1993). For example, out of the approximately three dozen

estuaries in Ventura, Santa Barbara and San Luis Obispo Counties, the only four

natural wetlands with horn snails (Mugu Lagoon, Carpinteria Salt Marsh, Goleta 

Slough and Morro Bay) are also the only four with consistent tidal flushing (nearly 

all the others have records of tidewater gobies, see below). The broad distribution and 

abundance of C. californica within its range make it an excellent reflection of regular 

tidal flushing and the associated marine invertebrate community. Horn snails may

even provide a historical record of tidal influence: their shells are preserved in 

sediment and can be retrieved from cores. The presence of horn snails also provides 

the possibility for using larval trematode parasites as indicators (as discussed later). 

One exception to the “horn-snail rule” is Elkhorn Slough in Monterey County, a fully

tidal estuary. Here, instead of horn snails, one finds an introduced ecological 

analogue, the mud snail, Batillaria attramentaria. A second exception is Campus 

Lagoon (in Santa Barbara County), an artificial marine pond without tidal influence 

(the seawater source is the outflow of the UCSB marine lab) where horns snails were

planted and survived for several years. 

Fishes 

A few dozen fishes use southern California estuaries (a dozen are common).

Particularly common or apparent species include killifish (Fundulus parvipinnis),

topsmelt (Atherinops affinis), arrow gobies (Clevelandia ios), longjaw mudsuckers 

(Gillicthys mirabilis), tidewater gobies (Eucyclogobius newberryi), and mullet (Mugil 

cephalus). Diversity is relatively low, and although species are well represented in 

field guides for the area, it is useful to have a guide for both freshwater (for example, 

McGinnis 1984) and marine fishes (Miller and Lea 1972 or Eschmeyer and Herald

1983). Some, such as sharks and rays, are more typical of the open ocean but will

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Assessing Estuaries—Lafferty

enter estuaries at high tide. Juveniles of reef species that are common in the near

shore may enter an estuary and stay if patches of hard substrate are present.

Historically, steelhead and lamprey migrated seasonally through estuaries. 

There are several methods for sampling fishes in estuaries. Like invertebrates, 

the abundance of fishes is higher in the warmer months, and sampling in the late 

summer/early fall yields the greatest catch. Because estuaries are relatively calm and 

shallow and often have smooth bottoms, seining, when conducted properly, can give 

a very good indication of the diversity and relative abundance of fishes at a particular 

location. If one is interested primarily in obtaining a list of common fishes, the 

following approach works fairly well with a crew of two to five people: At slack tide,

when currents are weak, we select three or more shallow locations (<1.5 m deep)

within the estuary. We then take relatively short seine hauls (<30 m) that can

terminate on a gently sloping, unvegetated bank or shore (avoiding areas with

excessive submerged rocks or other hazards). We find that a seine 10-20 m long by 

2 m high with a 1/8 – 1/4 inch (3-6 mm) mesh works well. Small mesh is better for

capturing small gobies, but the extra drag associated with small mesh necessitates 

using a shorter seine (which is not as good for capturing large fish or schools of fish). 

We fish until three to five successive seine hauls do not add any additional fish 

species to the accumulated species list. 

Minnow seines will not adequately sample larger wetlands with substantial

subtidal areas. Traps can be used in areas that are too deep to seine, because most

estuarine fishes will enter minnow traps (tidewater gobies are a notable exception). 

Traps require relatively little effort (except that they require two visits to a site) and 

do a better job at capturing large individuals that are able to out-swim seines. They

work best when baited and left out overnight; dry dog food makes a convenient bait. 

However, in shallow areas fish may die if traps are left out too long because low tides 

may expose traps to the air; anoxia may kill fish at night in traps set in deep water; or 

the captives may eat one another. In general, traps seem to have a high variance in 

what they catch: some traps come up empty next to traps filled with fish. There is 

also a tendency for some fish species to enter traps more than others (Layman and

Smith 2001). Although trapping is not a substitute for seining, often traps can add

new species not seen in a day’s seining. Gill nets are similar to traps in the sense that 

they are set and later retrieved. They are able to catch larger fish than minnow traps 

but mostly yield dead fish. 

In larger wetlands, benthic trawls (and, to a lesser degree, midwater trawls) are

the most tractable way to sample substantial subtidal areas. Trawling in an estuary 

requires a small boat, preferably one with a shallow draft. When trawling, it is 

important to keep track of the depth of the trawl and the area over which the trawl

was pulled. Trawls can also sample subtidal epibenthic invertebrates. 

Sampling fish for abundance is challenging. For traps and gill nets, the area 

sampled is not known and certain fishes are better able to avoid seines and trawls (by 

swimming or burrowing) than others. Rozas and Mellor (1997) advocate enclosure 

devices such as drop traps and throw samplers instead of seines. One way to obtain 

more accurate densities with seines is to first set up blocking nets that trap fish into a

moderately sized area and then seine repeatedly (for example, at least five times)

inside the area until no more fish are captured on three successive passes. Then, pull 

in the blocking nets carefully, one behind the other, to capture fishes that have sought

refuge in the nets. Blocking nets 20 m long are sufficient to block off most narrow 

channels. A smaller seine (5-10 m long) can be used for the successive passes. In

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Assessing Estuaries—Lafferty

areas where small gobies are abundant, 20 or more passes may be necessary to

adequately sample the area due to the habit of gobies to rapidly burrow. Narrow 

channels are easy to sample this way because blocking nets can be run from bank to 

bank. Areas adjacent to open shores can be blocked by forming a triangle, with the 

shore acting as one side and the blocking nets meeting at the apex offshore. 

A very useful indicator that an estuary is closed off to tidal flushing for at least

part of the year is the presence of the tidewater goby, Eucyclogobius newberryi. 

Tidewater gobies are often the most common fish in wetlands where they occur, 

particularly in the summer and fall. Unfortunately, tidewater gobies are an

endangered species and have been extirpated from large parts of their range, 

particularly in Los Angeles and Orange Counties (Lafferty and others 1996). 

Therefore, the absence of tidewater gobies does not indicate year-round tidal 

flushing; it may simply represent a local extirpation (Lafferty and others 1999). 

Birds 

Birds use estuaries in large numbers for food and rest during migration periods. Some 

species breed in southern California wetlands. Due to declines in the abundance of 

several species (for example, California least tern [Sterna antillarum browni], lightfooted clapper rail [Rallus longirostrus levipes], Belding’s savannah sparrow 

[Passerculus sandwichensis beldingi], western snowy plover [Charadrius 

alexandrinus nivosu], and brown pelican [Pelecanus occidentalis] are federally

protected, and several shorebirds are considered to be in decline), there is a growing 

need for understanding their status and trends. In addition, because they are 

charismatic and easy to view, bird communities appeal to the general public. The low

topographical relief in most estuaries makes it possible to observe most bird species 

in a particular field of view. Some wetlands in southern California are sufficiently 

small that all birds can be counted from just a few fixed points. With a few 

exceptions (such as female ducks, dowitchers (Limnodromus spp.), some terns,

immature gulls, and some sandpipers), most species of birds that use estuaries can be 

identified given a good look, basic birding experience, and a field guide (for 

example, Sibley 2000). Often data on birds exist in the form of notes from 

experienced birders, who often visit estuaries for their rich bird communities. Such 

notes can be useful but usually cannot replace a systematic sampling effort. However, 

amateur birders represent a potential pool of expert volunteers and, when using an 

appropriate sampling protocol, can provide excellent data. 

Because migration and breeding alter the sorts of birds that visit our estuaries

from month to month, it is necessary to limit comparisons among estuaries to 

samples taken in the same season. In addition, the bird community can be affected

over short periods of time by weather, tide, and time of day. For example, as the tide 

retreats, birds may move from roosting sites in the upper marsh or beach to forage on 

exposed mudflats. Tide and other variables should be controlled for in sampling or

accounted for in comparisons. 

Sampling in the spring provides a chance to detect breeders such as least terns,

while sampling in the fall yields a community augmented by migrants. Limiting 

sampling effort to aquatic birds (for example, gulls and terns, shorebirds, waterfowl, 

divers, rails, kingfishers, herons, and egrets) and birds of prey, while ignoring

terrestrial passerines, is one means to reduce effort and simplify analyses. However, 

in many cases, terrestrial passerines may be of particular importance if they are of

10 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Assessing Estuaries—Lafferty

conservation concern or use the margins of the estuary. Due to their smaller size and

more cryptic habits, terrestrial birds may not be effectively sampled by the techniques

described below. 

One may be able to survey a small (<50 acre) estuary in a few hours, but large 

estuaries may require subsampling or substantial sampling effort. In a small estuary,

a simple and repeatable approach is to count and identify all birds visible within

defined perimeters from fixed locations and fixed periods of time (for example, 30 

minutes). Choosing locations that provide good vantage of the mouth, middle and 

back sections of an estuary provides a good estimate of abundance by species. Then,

to determine a more complete total species list, one can roam throughout the estuary 

for 30-minute viewing periods until a period passes without detecting any new bird 

species. This facilitates the inclusion of the presence/absence of uncommon birds. 

To quantitatively subsample birds in an estuary, we have employed the 

following sampling method: Sites are selected from a map of the estuary (in our case, 

these sites are the randomly chosen locations for sampling other taxa). A 100-m

radius is then drawn around the sampling point. Using a GPS unit and a map of the 

site, observers approach the perimeter of this circle, mapping the location of birds

once they have a vantage of the interior of the plot. Then, over a 20-minute time 

interval, observers walk to and along a transect located along the circumference of a 

50 m radius around the central point. Birds are counted and mapped when observed.

This approach has the advantage of standardizing sampling time and transect length. 

It also tends to flush cryptic species that would otherwise not be observed from a 

fixed point. 

Recently, (with R. Hechinger) we have used remote video surveillance to 

determine bird visitation rates of small areas (on the scale of 10 by 10 meters). We 

have had success by taking five seconds of video every minute over a two week

period. Although this is the only way that we know how to effectively sample birds

in small areas, it has various limitations such as the expense and effort required to 

sample several remote sites simultaneously. Improvements in remote surveillance 

technology may soon increase the practicality of this approach for larger scale 

studies. 

New Approaches 

The Environmental Protection Agency is presently funding over 30 investigators in 

the exploration and development of new ecosystem indicators through the Pacific 

Estuarine Ecosystem Indicator Research (PEEIR) Consortium. The objective of this 

project is to develop indicators of wetland ecosystem health through evaluations of 

stressor-response patterns in wetland biota. The goal is to create an integrated suite of

indicators to evaluate impacts of toxic contaminants, nutrient enrichment, microbial

contamination, and exotic species invasions. The ecosystem-indicators component of 

PEEIR (led by S. Morgan) is investigating new ways of sampling the biota that may

some day be added to the more traditional methods discussed previously. Some

examples are described below. 

Measures of microbial community composition and diversity (led by T. Holden) 

using DNA from the T-RFLP (terminal restriction length polymorphism) method are 

highly transferable among sites and ecosystems, making them potentially valuable 

indicators of pollution and enrichment. Assessment of nitrogen isotope ratios in

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Assessing Estuaries—Lafferty

selected biota at southern California sites (led by M. Page) has promise for assessing 

ecosystem responses to nutrients. Refinement of methods to describe

ammonification, decomposition, and nitrification rates (led by S. Williams) may

generate information on trophic support and biogeochemical performance indicators. 

One method to investigate impacts to ecosystems is to evaluate changes in growth 

and reproduction of individual organisms by linking contaminant effects in dynamic 

mass-energy budget models. For example, analysis of fish otoliths to indicate growth

and rates (led by W. Bennett and A. Brooks) may provide a means to indirectly 

assess habitat suitability for fishes under different contaminant regimes. This 

information might then be integrated into mathematical models designed to assess

impacts at the population level. Parallel to the development of ecosystem indicators is 

research into other new technologies for assessment. For example, more sophisticated 

application of remote sensing, such as high spatial resolution airborne hyperspectral

imagery (led by S. Ustin), could provide quantitative mapping of canopy pigments,

water content, and dry standing litter. A value to the collaborative approach in 

developing indicators is that there may be important links among indicators that 

would not emerge from the development of each alone. 

Trematodes 

Using parasites as indicators of ecosystem function is the aspect of PEEIR that I 

oversee. Trematodes are parasitic worms with complex lifecycles. Several species 

use the common horn snail as a first intermediate host. There are 17 described species 

in California, and it is not unusual for half or more of the snails in an estuary to be 

infected by larval trematodes. The trematodes leave the snail in a free-swimming

cercarial stage and (with the exception of a blood fluke) seek out a second

intermediate host. This is usually a fish, but it can also be a crab or a mollusk, 

depending on the species of trematode. Most trematodes encyst inside their second 

intermediate host, but some encyst on the outside of snails or crabs. Mammals (such

as raccoons) and birds become infected when they eat infected second intermediate 

hosts. Trematodes mature and reproduce in birds and mammals, their eggs being

dispersed along with the final host’s feces. 

Because each trematode species has a unique life cycle that depends on many

interacting free-living animals, an estuary with a rich food web of fishes, birds, and

invertebrates can support a diverse and abundant community of trematodes. 

Therefore, the trematode community within a snail population provides, in a glance, 

the sum effect of several interacting communities of free-living animals. It is 

relatively easy to sample a population of snails for trematodes, all of which can be 

identified to species with a practiced eye using a dissecting scope. Because snails and 

their trematodes are long lived (up to more than 15 years), the community of 

trematodes in a population of snails provides an integrated history of the free-living

animals that have used the estuary on the scale of months or years (depending on the 

age of snails sampled). This is potentially more satisfying than the single day 

snapshot surveys that are usually the only possible way to sample birds, fishes, and 

invertebrates because these are subject to seasonal and, especially for birds, even 

daily fluctuations in abundance. It also makes trematodes ineffective for assessing

rapid changes occurring on short time frames. 

An observation of an urban population of horn snails first suggested to me that 

parasites could indicate habitat function (Lafferty 1993). I was sampling a site 

12 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Assessing Estuaries—Lafferty

separated from a marsh by a parking lot and bordered by a busy highway intersection. 

Birds were conspicuously absent (presumably due to the proximity of vehicle traffic), 

and the snails were completely uninfected compared with a prevalence of 25 percent 

in the adjacent marsh, suggesting that degradation had reduced trematode prevalence.

Other authors have also speculated that the prevalence of trematodes in snails 

corresponds to the degree of habitat degradation (reviewed by Lafferty 1997). Cort

and others (1960) made the first such comparison in Michigan. They found that larval 

trematode diversity and species richness had declined since studies conducted 20 

years before. They also noted an increase in human disturbance and a reduction in the

shorebird population. 

With Todd Huspeni, I have used trematode communities to assess the success of

restoration at Carpinteria Salt Marsh. As one would expect, degraded areas showed 

significantly reduced diversity and abundance of trematodes compared with reference 

areas. While trematode communities sampled from reference sites did not vary much 

from year to year, those in the degraded area rapidly increased in prevalence after

restoration. Newly created habitat produced trematode communities similar in

prevalence to reference sites. Still, at present, the reference sites and the created sites 

are not yet identical in their species diversities or compositions; restored areas have 

relatively more trematodes that use mollusks as intermediate hosts and relatively few 

trematodes that use fishes (Huspeni and Lafferty, unpublished manuscript). This 

underscores previous studies that indicate that newly restored estuarine habitat is not 

automatically the same as natural estuarine habitat. To make this approach broadly 

useable will require the additional ground-truthing of methods and their eventual 

publication and dissemination. 

Conclusion 

In summary, there are several ways that we can describe estuaries when we seek to

better understand their function. The most appropriate method varies with the goals 

and expertise of the evaluator. It is important to understand that there are a variety of 

types of estuaries that have unique properties, and these may be inappropriate to 

directly compare with one another. For example, when choosing an estuary or 

estuaries to act as a reference site, it would be important to be sure that size and tidal 

regime were similar to the target estuary. The horn snail makes an excellent indicator

of regular tidal flushing. Because sampling a single taxon will likely provide an 

incomplete picture of an estuarine community, attempting a variety of assessments is 

important. New technologies and ideas may provide innovative methods to assess 

estuary biota. A description of a larval trematode assemblage is a single measure that

captures the ability of an estuary to support a rich and functioning community of 

invertebrates, fishes, birds, and mammals. However, sampling must be consistent in

methodology, and effort must be comparable among sites. 

Acknowledgments 

Wayne Ferren developed most of the estuarine classification scheme I use. Mark

Page introduced me to estuaries, Armand Kuris introduced me to trematodes, and 

Mark Holmgren taught me birds. Todd Huspeni has furthered our efforts to use 

trematodes as indicators. Richard Ambrose supported and assisted much of the

sampling effort where these techniques have been used. Many technicians and

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 13 
Assessing Estuaries—Lafferty

volunteers have contributed time and expertise in the field, especially Nicole Mode 

and Jessica Altstatt. Nick Kalodimos and Ryan Hechinger proofread the manuscript. 

In addition, the paper has benefited from support received from the National Science 

Foundation through the NIH/NSF Ecology of Infectious Disease Program (DEB0224565). This research also has been supported by a grant from the U.S. 

Environmental Protection Agency's Science to Achieve Results (STAR) Estuarine

and Great Lakes (EaGLe) program through funding to the Pacific Estuarine 

Ecosystem Indicator Research (PEEIR) Consortium (U.S. EPA Agreement

#R-882867601). Although the research described in the article has been funded

wholly or in part by the U.S. Environmental Protection Agency, it has not been

subjected to any EPA review and, therefore, it does not necessarily reflect the views

of the Agency, and no official endorsement should be inferred. 

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State of California, Dept. of Fish and Game; 235 p. 

Morris, R.H.; Abbott, D.P.; Haderlie, E.C. 1980. Intertidal invertebrates of California. Palo

Alto, CA: Stanford University Press; 690 p. 

NOAA (National Oceanic and Atmospheric Administration). 1990. Estuaries of the United 

States, statistics of a national resource base. Special NOAA 20th Anniversary Report,

Strategic Assessment Branch, Office of Oceanography and Marine Assessment, 

Washington, DC.; 79 p. 

Pacific Estuarine Research Laboratory (PERL). 1990. A manual for assessing restored and 

natural coastal wetlands with examples from southern California. California Sea 

Grant Report No. T-CSGCP-021. La Jolla, California. 105 p. 

Rozas, L.P.; Minello, T.J. 1997. Estimating densities of small fishes and decapod

crustaceans in shallow estuarine habitats: A review of sampling design with focus on 

gear selection. Estuaries 20: 199-213. 

Sandoval, C.P.; Lafferty, K.D. 1995. Invertebrate communities. In Ambrose, R.F., editor.

Coastal wetland resources: Santa Barbara County Mainland. Final Report to the County

of Santa Barbara; 39-45. 

Sibley, D.A. 2000. The Sibley guide to birds. New York: Knopf; 545 p.

Smith, R.I.; Carlton, J.T. 1975. Light’s manual: Intertidal invertebrates of the central

California coast. Berkeley, CA: University of California Press; 716 p. 

Zedler, J.B. 1982. Salt marsh algal mat composition: Spatial and temporal comparisons.

Bulletin of the Southern California Academy of Sciences 81: 41-50.

Zedler, J.B. 1984. Salt marsh restoration: A guidebook for southern California. California 

Sea Grant College Program, Institute of Marine Resources, University of California, A032, California Sea Grant Report no. T-CSGCP-009; 46 p.

Zedler, J.B. 1982. The ecology of southern California coastal salt marshes: A community

profile. (Second printing, with corrections 1984). U.S. Fish and Wildlife Service, 

FWS/OBS-81-54; 110 p. 

Zedler, J.B.; Huffman, T; Josselyn, M., editors. 1990. Pacific regional wetland functions:

Proceedings of a workshop held at Mill Valley, California, April 14-16, 1985. 

Amherst, MA: The Environmental Institute, Univ. of Massachusetts at Amherst; 162 p.

Zedler, J.B.; Langis, R. 1991. Comparisons of constructed and natural salt marshes of 

San Diego Bay. Restoration and Management Notes 9: 21-25. 

Zedler, J.B.; Nordby, C.S.; Kus, B.E. 1992. The ecology of Tijuana Estuary, California: A 

national estuarine research reserve. Washington DC: National Oceanic and 

Atmospheric Administration Office of Coastal Resource Management, Sanctuaries and 

Reserves Division; 151 p. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 15 

Up, Down, or Stable: Populations of 

Endangered Birds in Beach and Estuarine 

Areas in Southern California1

Abby N. Powell2

Abstract 

The coastal beach-dune ecosystem in California supports two federally listed threatened or

endangered species: California least tern (Sterna antillarum browni) and western snowy 

plover (Charadrius alexandrinus nivosus). This ecosystem has become highly stressed due to

shoreline development, invasion of exotic plants, beach stabilization, and heavy recreational

use. Least tern populations have increased significantly since the 1980s. Snowy plover

populations, however, show continued decline along California’s coast. Management 

practices to protect least tern nesting colonies appear to be successful for terns but do not 

offer protection for plovers. More research is needed to monitor reproductive success and 

survival of snowy plovers in order to improve management practices and preservation of 

nesting areas. Impacts to estuarine ecosystems have been severe and include loss, degradation, 

and fragmentation. The most imperiled species within this system is the endangered lightfooted clapper rail (Rallus longirostrus levipes). Rail populations have ranged from a low of 

142 pairs in 1985 to a high of 325 pairs in 1996 during the 22-year period they have been

monitored. Most alarming is that only one estuary consistently supports more than 50 percent

of California’s rails, and only three sites support more than 80 percent. Although Belding’s 

savannah sparrow (Passerculus sandwichensis beldingi) is not federally listed as endangered, 

populations in estuaries are stable or declining, but monitoring is erratic at best. Belding’s

savannah sparrows are area-sensitive, and reproductive success appears to be low in

fragmented marshes. Research on productivity and survival of rails and sparrows is needed. 

Habitat enhancement and creation should be the highest priority for beach and estuarine birds 

as well as migratory shorebirds, but predator management and restrictions on recreation beach 

use are also necessary. 

Key words: Belding's savannah sparrow, California least tern, endangered, estuarine, lightfooted clapper rail, western snowy plover

Introduction 

The beach and estuarine ecosystems of southern California have suffered significant 

degradation and loss of natural habitats. Beach ecosystems are altered by heavy 

recreational use, urban development, beach stabilization efforts, river channelization, 

and non-indigenous plant species (Federal Register 1993, Page and Stenzel 1981,

Powell 1998, U.S. Fish and Wildlife Service 1985, Veirs and others 1998). Loss of 

beach habitats resulted in declines of beach-nesting birds, including the California 

least tern. In 1970, the California least tern was listed as federally endangered, with 

the California population estimated at 600 pairs (U.S. Fish and Wildlife Service 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2 Research Wildlife Biologist, Alaska Cooperative Fish and Wildlife Research Unit, USGS, Department 

of Biology and Wildlife, University of Alaska, Fairbanks, AK 99775-7020 (e-mail: ffanp@uaf.edu) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 17 
Endangered Birds in Beach and Estuarine Areas—Powell 

1985). More recently, in 1993, the Pacific Coast population of western snowy plover 

was federally listed as threatened (Federal Register 1993). 

California has lost an estimated 91 percent of its wetlands since pre-settlement

times, and remaining estuaries are fragmented. Estuarine systems in southern

California have been highly altered by urban development, filling, river

channelization, changes in freshwater flow, and invasion of exotic species. Loss of 

total wetland area and degradation of remnant estuaries have caused many wetlanddependent species to be listed as special concern by the State and Federal

governments (Veirs and others 1998). These include the light-footed clapper rail, 

federally listed as endangered in 1970 (U.S. Fish and Wildlife Service 1979), and the 

Belding’s savannah sparrow, listed as endangered by the State of California in 1974

(State and Federal endangered and threatened animals of California and their listing 

dates, unpublished report, State of California, 1994).

What do we know about the populations of these bird species that depend on

beach and estuarine ecosystems? How are resource agencies using population 

information to manage these endangered and threatened birds? Although three of the 

four species mentioned above were listed more than 25 years ago, information on

population status varies in quality, quantity, and availability. This paper summarizes 

the current knowledge of the four species’ populations in the south coast ecoregion of

California. In addition, I include recommendations for future management and 

research for each species. 

California Least Tern 

Migratory California least terns are present in California from the Mexican border 

north to San Francisco Bay only during their breeding season, which is mid-April

through September. Since being listed as endangered in 1970, management actions 

have included fencing colonies, limiting access to nesting areas, and, after the mid1980s, managing predators. Monitoring of least tern colonies to track changes in

population trends and productivity began in 1973; efforts to cover all areas have 

increased over the past 20 years (Patton 2002). Population monitoring, however, is 

not uniform at each colony, and methods range from non-invasive surveys of terns on 

nests to intensive efforts that include marking each nest and banding chicks. 

Monitoring is conducted by State and Federal personnel at some sites and by

biological consultants at others. There has been statewide coordination of population

data compilation since the mid-1970s (Ron Jurek, California Department of Fish and

Game, personal communication). 

The California least tern population was estimated at 600 breeding pairs in 1970 

(Fancher 1992). For removal from the endangered species list, there must be 1,200 

breeding pairs distributed among at least 20 coastal areas and an average over 5 years 

of one fledgling produced per breeding pair per year (U.S. Fish and Wildlife Service 

1985). The population of California least terns has increased dramatically since 1988, 

growing from approximately 1,000 pairs to more than 4,000 pairs within 10 years 

(fig. 1). Least terns nest at approximately 34-39 sites in southern California, and the

number of breeding pairs has exceeded 2,000 since 1992 (Keane 2001, Patton 2002). 

Although estimates of fledgling production have fluctuated, particularly in El Niño 

years (Massey and others 1992), the total number of fledglings produced

18 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Endangered Birds in Beach and Estuarine Areas—Powell 

0

250

500

750

1000

1250

1500

1750

2000

2250

2500

2750

3000

3250

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4000

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4500

4750

1970 1972 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000

Year

Number of Least Terns

Pairs

Fledglings

Recovery Goal = 1200 Breeding Pairs

Figure 1—Number of breeding pairs and fledglings of California least terns in

California, 1978-2000. Data are from Keane (1998) and from unpublished data on file

from California Department of Fish and Game, Sacramento, CA.

increased in the same time period (fig. 1). However, dividing the estimated total 

fledglings by the estimated number of pairs per year reveals that the recovery goal of 

one fledgling per pair is not being met (mean fledglings/pair = 0.68 ± 0.21 SD, years 

1987-2000). It is unknown what proportion of the population and productivity

increases can be attributed to increased monitoring efforts over the same time period. 

Regardless, increased breeding populations and productivity have been

attributed to protection of nesting colonies and predator management (Fancher 1992).

Fencing varies from site to site and may exclude people or both people and

mammals. Predators are managed by exclusion and direct removal; however, 

information on the numbers of predators removed is difficult to obtain. Predators 

include, but are not limited to, common raven (Corvus corax), American crow (C.

brachyrhynchos), American kestrel (Falco sparverius), peregrine falcon (F.

peregrinus), burrowing owl (Athene cunicularia), coyote (Canis latrans), domestic 

cat (Felis catus), gray fox (Urocyon cinereoargenteus), long-tailed weasel (Mustela

frenata), and raccoon (Procyon lotor). In addition, increased monitoring efforts

facilitate adaptive management of tern colonies (Keane 1998, Patton 2002). Of the 

four species described in this paper, we know the most about California least tern

populations, which have increased exponentially since 1988. 

Western Snowy Plover 

Although snowy plover habitat overlaps with the habitat of California least terns, 

there are several important differences in the biology, management, and population 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 19 
Endangered Birds in Beach and Estuarine Areas—Powell 

status of these two species (Powell 1998, 2001). Unlike least terns, snowy plover 

populations in California are a mixture of migratory and resident birds (Powell and 

others 2002, Stenzel and others 1994); thus conservation is an issue year-round. In 

winter, snowy plovers from northern California migrate to southern California. In 

addition, a proportion of southern California breeders overwinter in the same region. 

Snowy plovers begin nesting in early March, well before least terns arrive from their

wintering grounds in mid-April. 

There are large information gaps about snowy plovers that breed and winter in

southern California. The Pacific Coast population of western snowy plovers was 

federally listed as threatened in 1993 because statewide surveys showed overall 

declines in breeding and wintering populations (Federal Register 1993). Until 1994,

data on numbers of plovers, productivity, and survival in the south coast region was 

sporadic at best. Page and others (1991) estimated the number of individual breeding

plovers in California in 1989 (1,386) was 89 percent of the number found from 1977 

to 1980 (1,565). Coastal lagoons in southern California historically supported

breeding snowy plovers, but habitat loss and alteration of water flows have degraded

these sites; they now only sporadically support breeding pairs. Snowy plovers no 

longer breed within Los Angeles County, and only one nesting area, Bolsa Chica 

Lagoon, remains in Orange County.

Most of what is known about breeding populations of snowy plovers in Ventura, 

Santa Barbara, and San Luis Obispo counties comes from “window” surveys 

conducted approximately every three years by the Point Reyes Bird Observatory.

These surveys take place during a specific “window” of time, usually 1-2 days during 

peak nesting (last week of May). The most recent window survey in 2002 found

1,379 adult plovers in California (unpublished data on file, Point Reyes Bird

Observatory, Stinson Beach, CA). The largest numbers of breeding plovers found in 

1991, 1995, 2000, and 2002 were at Vandenberg Air Force Base (AFB; 242, 213, 

106, and 179, respectively), the Channel Islands (199, 196, 89, and 79), San 

Francisco Bay (176, no data for 1995, 96, and 78), and northern San Diego County 

(48, 49, 63, and 80) (unpublished data on file, Point Reyes Bird Observatory, Stinson 

Beach, CA.). In addition, Point Reyes Bird Observatory has long-term data on snowy

plovers in the Monterey Bay area. Their data indicate that the average number of 

fledglings per adult ranged from 0.85-0.86 annually, but after nests were actively

protected from people and predators in 1992, fledging rates increased to over 1 

fledgling per adult (unpublished data on file, Point Reyes Bird Observatory, Stinson

Beach, CA). A comprehensive study on the population dynamics of snowy plovers in 

San Diego County was initiated in 1994 and continued through 1999 (Powell and 

others 2002), and the U.S. Fish and Wildlife Service began monitoring snowy plovers 

at Bolsa Chica Lagoon in Orange County in 1997. On average, 50 percent of all 

snowy plovers in San Diego and Orange counties nested at one site, Marine Corps 

Base Camp Pendleton. Overall reproductive success was low, with estimated nest 

success in San Diego County typically less than 55 percent with fewer than 0.5 

fledglings produced per nest (Powell and others 2002). 

The creation of nesting habitat for least terns from dredged materials has been a 

popular component of habitat restoration to partially compensate for wetland loss in 

southern California. Five new nesting areas were created at Batiquitos Lagoon in 

1994 and 1995. This restoration project provided a unique opportunity to study

natural recolonization and subsequent success of snowy plover populations from 

1994 to 1998 (Powell and Collier 2000). Snowy plovers traditionally nested at the 

20 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Endangered Birds in Beach and Estuarine Areas—Powell 

lagoon along its perimeter (5 nests in 1994 prior to construction). After the creation 

of new least tern nest sites, snowy plovers began to use these areas immediately.

Although the number of nesting attempts at Batiquitos Lagoon increased from 5 in 

1994 to a high of 38 in 1997, fledge rates and nest success declined after peaking the 

first year the created sites were used, suggesting that the area attracted nesting 

plovers but ultimately became a habitat sink (Pulliam 1988). Predation pressure and

quality of habitat were important in determining habitat use and reproductive success 

on the newly created areas. As more birds, including least terns, used the created

sites, more predators were observed there. In addition, the amount of cover on the 

created sites may have contributed to lowered nest and chick survival. Areas devoid 

of cover provided little visual or thermal protection, while other areas became 

covered with invasive weeds, which made sites less attractive to nesting plovers

(Powell and Collier 2000). In summary, snowy plovers easily colonized newly

created sites and were highly productive immediately after colonization. However,

productivity decreased with time and was related to habitat quality and predation 

pressure. 

More monitoring and research on the effects of California least tern 

management on snowy plovers is needed in the region. Plovers are more vulnerable 

to human disturbance than least terns. In addition, estimating population growth by

monitoring nests or pairs of breeding plovers is not enough to determine population 

status, because some areas may serve as population sinks (Pulliam 1988). It is

necessary to obtain information on annual productivity in addition to number of 

breeding plovers. No area in San Diego County or Orange County produces enough

snowy plovers to be considered a source population, and knowledge about survival 

and productivity of snowy plovers north of these counties is sparse. 

Light-Footed Clapper Rail 

In the United States, this subspecies of clapper rail is a year-round resident of salt 

marshes from Tijuana Estuary on the Mexican border north to Santa Barbara County

(U.S. Fish and Wildlife Service 1985). Considerable research has been done on 

habitat requirements and behavior of light-footed clapper rails (Massey and others 

1984, Massey and Zembal 1987, Zembal and Fancher 1988, Zembal and others 1985, 

1989). In addition, genetic analysis of the subspecies indicated there was low genetic 

variability and a metapopulation with local populations that experienced extirpation, 

recolonization, and limited dispersal (Fleischer and others 1995). 

Light-footed clapper rail populations have been monitored annually since 1980.

Marshes that potentially support clapper rails are visited in spring, and clapper rail 

calls are counted. Clapper rails use several distinct calls during the breeding season 

that can be used to distinguish single males, single females, and mated pairs (Massey

and Zembal 1987). During a census, people walk slowly through the marsh at dawn

or dusk and mark locations of calls on a map. In addition, taped calls may be played

to elicit responses (Zembal 1998). Each year a breeding survey report is submitted to

the California Department of Fish and Game. In 1980, the first year of the survey, the

light-footed clapper rail population consisted of an estimated 203 breeding pairs; a 

high of 325 breeding pairs was counted in 1996. The number of estimated pairs has 

varied around the 22-year mean of 231 pairs, but there has been no overall pattern of 

decline or increase during this period (fig. 2). 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 21 
Endangered Birds in Beach and Estuarine Areas—Powell 

0

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1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 2002

Year

Number of Breeding Pairs

California Total

Tijuana Estuary NWR

Upper Newport Bay

Seal Beach NWR

22-year mean

Figure 2—Number of breeding pairs of light-footed clapper rails in California, 1980-

2002. Data are from Richard Zembal, U.S. Fish and Wildlife Service, personal 

communication, unpublished data on file, California Department of Fish and Game,

Sacramento, CA. 

Predator management to remove non-native red foxes (Vulpes vulpes) resulted in

growth of the population at Seal Beach National Wildlife Refuge (NWR) from a low 

of 5 pairs in 1986 to 65 pairs in 1993 (Zembal and others 1998). In addition to

predator management, nesting rafts have been deployed at this site to increase nest

site availability since the late 1980s. However, the population at Seal Beach appears 

to be lower (range 5-24 pairs) from 1995-2002 than in the early 1990s (fig.2). It is 

postulated that rafts may increase the incidence of raptor predation on clapper rails at 

this site (Zembal and others 1998). 

Upper Newport Bay Ecological Reserve consistently supported more than 50

percent of California’s light-footed clapper rails, and together three sites (Upper 

Newport Bay Ecological Reserve, Tijuana Estuary NWR, and Seal Beach NWR)

supported more than 80 percent of breeding pairs in any given year. These three 

estuaries are the second, third, and fourth largest in size within the range of the clapper rail. Mugu Lagoon, the largest wetland, supported on average only four pairs of

rails. Of the remaining 21 sites where light-footed clapper rails are found, none 

supported more than 4 percent of the metapopulation, and 15 each supported less 

than 1 percent of the metapopulation. 

Light-footed clapper rails are associated with cord grass (Spartina foliosa) habitats within coastal marshes in southern California (U.S. Fish and Wildlife Service 

1979). Although it is unlikely that new estuaries will be created in this region, it is

possible to improve and expand cord grass coverage in existing marshes. Foin and 

Brenchley-Jackson (1991) suggested that cord grass habitat improvement within existing marshes could potentially triple the rail population. Despite the severe limitation of available habitat, there are captive breeding and reintroduction efforts 

22 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Endangered Birds in Beach and Estuarine Areas—Powell 

currently being initiated (unpublished data, California Department of Fish and Game, 

Sacramento, CA). It has been well documented that the key predictors of successful

translocations are habitat quality and the number of animals released. In general,

endangered species translocations are unsuccessful more than 50 percent of the time

(Griffith and others 1989). Likewise, if animals are released into habitats that are in

poor condition or have insufficient area, they are unlikely to persist (Wolf and others 

1996, 1998). Fleischer and others (1995) cautioned that translocations of light-footed 

clapper rails should be avoided until there is documentation that inbreeding depression is a problem for this species. Finally, more information is needed on natural recruitment into light-footed clapper rail subpopulations. Given the small number of 

light-footed clapper rails remaining in southern California and the limited and degraded condition of estuarine habitat, efforts to increase populations should emphasize habitat creation and enhancement rather than costly translocations with low 

potential for success. 

Belding’s Savannah Sparrow 

Like light-footed clapper rails, Belding’s savannah sparrows are non-migratory and

endemic to southern California salt marshes. Unlike rails, however, Belding’s

savannah sparrows are associated with pickleweed (Salicornia sp.) habitats, which 

are much more expansive than cord grass habitats (Powell 1993). Pickleweed habitats 

have been degraded by changes in tidal flow and freshwater inputs, invasion of nonindigenous plants, and fragmentation by trails and roads. Restoration of this habitat 

type is difficult and may not result in suitable nesting habitat for sparrows (Keer and

Zedler 2002). 

Belding’s savannah sparrows are censused in spring by counting singing males 

and observing breeding pairs. They have been censused sporadically in 26 to 30

coastal salt marshes in southern California, and intensity of efforts to count them 

varied among wetlands and years. All 30 marshes were surveyed in 1986, 1991,

1996, and 2001, and the total estimated number of breeding pairs ranged from 1,844

to 2,902 (Zembal and Hoffman 2002, Zembal and others 1988,). Mugu Lagoon, the 

largest salt marsh in southern California, supported the largest subpopulation of 

sparrows, followed by Tijuana Estuary NWR, Upper Newport Bay Ecological 

Reserve, and Seal Beach NWR. Except in Mugu Lagoon, these sites also supported 

the largest subpopulations of light-footed clapper rails despite the differences in

habitat associations of the two species. 

There is evidence that Belding’s savannah sparrow has limited dispersal and

exists as a metapopulation with extirpation and recolonization of local populations 

(Bradley 1994, Burnell 1996, Zembal and others 1988). Like grassland savannah

sparrows (P. sandwichensis), Belding’s savannah sparrows are area-sensitive: there is 

a positive relationship between size of wetland and indices of population size, and

sparrows are unlikely to occur in marshes less than 10 ha in size (Powell and Collier 

1998). Recent research on reproductive success of Belding’s savannah sparrows in

different-size wetlands suggested that small, isolated salt marshes supported breeding 

birds but functioned as population sinks because they supported little or no 

productivity at those sites (Powell and Collier 1998). Finally, recent work on

breeding biology of Belding’s savannah sparrows indicates that the effective

population size is likely much smaller than the number of males present; less than 50 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 23 
Endangered Birds in Beach and Estuarine Areas—Powell 

percent of males establish territories, and only 43 percent of the territorial males 

manage to attract mates (Burnell 1996). 

Ecosystem Importance 

There is considerable overlap of the estuaries and associated beaches most important 

to these four species, and many of the important areas are on federal or State lands 

(table 1). Mugu Lagoon, for example, is among the top five breeding areas for both 

snowy plovers and Belding’s savannah sparrows, and Santa Margarita River estuary 

and beach areas are critical breeding areas for California least terns and snowy

plovers. Both Mugu Lagoon and the Santa Margarita River estuary are within Federal 

lands managed by the military (Mugu Naval Air Station (NAS), Camp Pendleton 

MCB). Vandenberg AFB supports the largest number of breeding snowy plovers in 

southern California. Other important federal lands include National Wildlife Refuges 

(Tijuana Estuary NWR, Seal Beach NWR, and Bolsa Chica Lagoon) and National

Park Service (Channel Islands) lands. Upper Newport Bay, which is managed as an 

Ecological Reserve by the California Department of Fish and Game, supports the 

most light-footed clapper rails in the state as well as being the third most important

site for Belding’s savannah sparrow. 

Table 1—Top five sites, listed in descending order of population size, for breeding California

least terns, western snowy plovers, light-footed clapper rails, and Belding’s savannah 

sparrows in southern California. 

Species Site Management1

California least tern Santa Margarita Estuary Camp Pendleton MCB 

Venice Beach CDFG

Huntington Beach CDPR, City of Costa Mesa 

Coronado NAB Coronado

Mission Bay City of San Diego 

Western snowy plover Vandenberg Vandenberg AFB

Channel Islands NPS, TNC 

Santa Ynez River USAF

Mugu Lagoon Mugu NAS

Santa Margarita Estuary Camp Pendleton MCB 

Light-footed clapper rail Upper Newport Bay CDFG 

Tijuana Estuary NWR, CDPR

Seal Beach NWR

Kendall-Frost Reserve Univ. of California 

Carpenteria Marsh Univ. of California, CDPR 

Belding’s savannah sparrow Mugu Lagoon NAS 

Tijuana Estuary NWR, CDPR

Upper Newport Bay CDFG

Seal Beach NWR

Bolsa Chica FWS 1 AFB = Air Force Base, CDFG = California Department of Fish and Game, CDPR = California 

Department of Parks and Recreation, FWS = U.S. Fish and Wildlife Service, MCB = Marine Corps 

Base, NAB = Naval Amphibious Base, NAS = Naval Air Station, NPS = National Park Service, NWR = 

National Wildlife Refuge, TNC = The Nature Conservancy, USAF = U.S. Air Force.

24 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Endangered Birds in Beach and Estuarine Areas—Powell 

In addition to providing habitat to these four threatened and endangered species, 

these sites are important to wetland-dependent birds in general, particularly migratory 

species such as shorebirds. For example, Mugu Lagoon supports more than 65,000 

shorebirds during spring migration and more than 10,000 in fall (Page and Shuford

1999). Seal Beach NWR, Bolsa Chica Lagoon, and Upper Newport Bay Ecological 

Reserve all support more than 5,000 shorebirds in fall and winter, and Santa 

Margarita River estuary and Tijuana Estuary NWR each hold more than 1,500 

shorebirds during parts of the year (Page and Shuford 1999). The beaches associated

with these wetlands provide habitat for snowy plovers, least terns, and other species

of shorebirds. Restoration, enhancement, and creation of habitat at these sites and all 

remaining coastal wetlands will support biodiversity of beach and estuarine birds 

within the region. 

Conclusions 

Habitat degradation includes increased rates of predation related to human causes.

Nonnative red foxes have expanded their ranges and populations in California and

impact coastal ecosystems (Lewis and others 1999). There is evidence that common

raven (Corvus corax) populations have increased substantially in California since the 

1960s (Boarman and Berry 1995). Likewise, the proliferation of feral and domestic 

cats in urban areas has a significant impact on native birds (Ogan and Jurek 1997). A 

predator management plan is necessary to identify problem areas and control 

predation rates. 

In addition, the dense human population in southern California coupled with 

popular recreational beach use has led to high levels of human disturbance within 

breeding and foraging areas for terns and plovers and surrounding marshes. Humans 

can impact bird populations through direct (for example, destruction of nests by 

trampling, vehicles, and vandalism) and indirect disturbance (for example, causing

birds to abandon nests, separating parent birds from their young, bringing pet dogs 

into nesting areas, leaving trash that attracts predators). Limiting human access to

nesting areas during the breeding season is essential for the conservation of these

species. 

Habitat preservation and enhancement of both beach and estuarine systems are

essential for both long-term persistence of endangered birds and biodiversity along 

the southern California coast. Estuaries are natural islands of habitat, and principles 

of island biogeography and conservation biology should be applied to planning and 

management efforts. Light-footed clapper rail and Belding’s savannah sparrow 

populations have experienced isolation and local extirpations, and the smallest and

most isolated marshes are unlikely to contribute to productivity or overall recruitment 

for either species. Degradation and loss of surrounding native habitats have likely

affected survival and productivity by increasing predation rates and limiting dispersal

of these wetland birds. Nevertheless, these sites may still attract birds and thus

behave as ecological traps. 

To date there has been no organized effort to characterize existing estuaries 

across the region according to habitat type, patch size and shape, connectivity, and

isolation (Zedler 1996). This information is critical, especially for the management of 

the year-round residents (light-footed clapper rails and Belding’s savannah 

sparrows). Both subspecies are metapopulations endemic to southern California and

Baja California, Mexico, and should be managed as such. Planners and managers 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 25 
Endangered Birds in Beach and Estuarine Areas—Powell 

need to ask the following questions before designing habitat restoration and 

enhancement projects: (1) Are wetlands large enough to support self-sustaining 

populations over time? (2) Are patches of specific habitat types (for example, cord 

grass or pickleweed) large enough to support self-sustaining populations over time? 

(3) Are poor dispersers able to move between wetlands (will source/sink populations

even out) over time? Restoration and management planners need to consider these 

issues to maximize habitat quality and reduce the potential for creating ecological 

traps. 

Finally, a regional approach to wetland restoration is needed to enhance

metapopulations of sparrows and rails as well as migratory species such as snowy 

plovers, least terns, and other shorebirds. For example, given the limited overall area

within each wetland, increasing cord grass habitat to benefit clapper rails may be at 

the expense of pickleweed habitats required for Belding’s savannah sparrows and 

vice versa. Converting salt pan or dredged areas to salt marsh may reduce the amount 

of habitat available to least terns, snowy plovers, and other shorebirds. 

Considerations should be giving to the status of the target species, probability of

success of habitat restoration, and overall ecosystem functioning. 

Acknowledgments 

I would like to thank the people who have helped me with research on Belding’s 

savannah sparrows and western snowy plovers over the years: Christine Collier,

Bonnie Peterson, and Jill Terp. Richard Zembal has kindly shared data on surveys of 

light-footed clapper rails and Belding’s savannah sparrows over the years, and Kathy 

Keane, Carolee Caffrey, Robert Patton, and Ron Jurek have provided information on

California least terns. 

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Powell, A.N.; Collier, C.L. 2000. Habitat use and reproductive success of western snowy 

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Jennings, M.R.; McEachern, K.; Moyle, P.B.; Stefani, R.A. 1998. California. In: Mac,

M.J.; Opler, P.A.; Puckett Haecker, C.E.; Doran, P.D., editors. Status and trends of the

nation's biological resources. 2 vols. Reston, VA: U.S. Department of the Interior, U.S. 

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Reanalysis with phylogenetically independent contrasts. Biological Conservation 86:

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1154. 

Zedler, J.B. 1996. Coastal mitigation in southern California: The need for a regional

restoration strategy. Ecological Applications 6: 84-93. 

Zembal, R.; Fancher, J.M. 1988. Foraging behavior and foods of the light-footed clapper 

rail. Condor 90: 959-962. 

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light-footed clapper rails indicated in part through regular censusing. California Fish

and Game 71: 164-171. 

Zembal. R.; Hoffman, S.M. 2002. A survey of the Belding’s savannah sparrow

(Passerculus sandwichensis beldingi) in California, 2001. Species Conservation and

Recovery Program Report 2002-01. Sacramento, CA: Calif. Dept. Fish and Game,

Habitat Conservation Planning Branch; 12 p. 

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assessment, 1997. Species Conservation and Recovery Program Report 1998-01.

Sacramento, CA: Calif. Dept. Fish and Game; 23 p. 

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sparrows in California. American Birds 42: 1233-1236. 

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light-footed clapper rail. Journal of Wildlife Management 53: 39-42. 

28 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Considerations for the Management of 

Vernal Pool Faunal Communities1

Marie A. Simovich2

Abstract 

The faunal community of California’s vernal pools is less obvious and has been less studied

than the floral community but is similarly diverse, endemic, and endangered. Species richness

and composition vary between pool types, between pools of a given type, and within

individual pools between years and through the season. The physiology, life history, and

ecology of the pool organisms contribute to this complexity. Management considerations

must take into account a large number of factors including maintaining pool hydrology and

water chemistry, avoiding crushing cysts in the soil, avoiding the input of toxins, and 

maintaining genetic variability. Decisions must be made at the landscape level and focus on

functioning communities, not just individual species or single pools. 

Key words: crustacean communities, management, restoration, vernal pools 

Introduction 

California’s diverse climate and topography support a wide variety of ephemeral 

wetlands. Among these are the vernal pools of the Central Valley and southern 

coastal mesas. These pools are inherently biphasic. In most areas, the pools are dry 

for the majority of the year, but they are temporarily converted into aquatic habitats 

when they fill from winter rains or spring snow melt. The pools then support a 

diverse flora and fauna, specifically adapted to their oscillating extremes. 

The floral communities of California’s vernal pools are quite well known (for

example, Bauder 1987, Holland 1978, Holland and Jain 1988, Hoover 1937, Stone 

1990, Zedler 1987); however, the faunal communities have been less studied. 

Although amphibians and waterfowl are quite visible, the much smaller invertebrates 

are not. These include crustaceans, rotifers, insects, worms, and snails. 

The crustacean contingent is the more visually apparent of the invertebrates and 

includes branchiopods (fairy shrimp, tadpole shrimp, clam shrimp, and water fleas),

ostracods (seed shrimp), and copepods. These obligately aquatic organisms survive

the terrestrial phase of the pool as desiccation resistant, diapausing propagules (cysts, 

embryos, or eggs) in the soil. Upon hydration, they hatch, quickly develop, mate, and 

produce new propagules. The life cycle is compact because the pools are ephemeral. 

In some areas pools may last for months; in others they may last a few weeks or

sometimes just days. 

Ephemeral pools only exist in areas that are relatively flat and underlain by 

impermeable soils. Generally, the pools occur in clusters or complexes, creating a

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2 Department of Biology, University of San Diego, 5998 Alcala Park, San Diego, CA 92110 (e-mail 

simo@sandiego.edu)

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 29 
Vernal Pool Fauna—Simovich

spatial patchwork across the landscape. Unfortunately these are also the areas utilized 

for agriculture and urban development. It is currently estimated that for these reasons, 

90 percent of the pools have been lost statewide (Holland 1978, Stone 1990). In

rapidly growing areas such as San Diego County, the estimates are upwards of 97 

percent (Bauder and Wier 1990). This loss of vernal pools and their associated

populations has resulted in the listing of several species of vernal pool plants and 

animals as endangered (Federal Register 1980, 1991, 1992, 1993a, 1993b, 1994,

1997). Unfortunately, a lack of information regarding obligate vernal pool 

invertebrates complicates this problem. Many more organisms are taxonomically

undescribed and may merit listing as well, but their distribution, ecology, and life 

history are not well known. Only a very few have had primary research targeted

toward them specifically. Furthermore, symbiotic interactions within the community

and roles of community members in ecosystem cycles are poorly understood. 

Consequently, decisions regarding management, protection, restoration, and recovery

have been difficult. 

Here, I briefly discuss a few key findings from the research on vernal pool 

organisms (focusing on crustaceans) and suggest some of the implications which are 

relevant to management decisions. The majority of the research to date concerns

branchiopods, but most of the implications for management are expected to extend to

other taxa as well. 

Key Research

Diversity and Endemism 

California’s vernal pool invertebrate communities exhibit quite high alpha (within

site) and beta (across space) diversity. 

• Single pools may have as many as 27 species of crustaceans, which exceeds 

the diversity seen in most permanent lakes (King and others 1996). 

• Pools of different types (due to location and soil chemistry) can differ

significantly in the species that are present (King and others 1996, Simovich

1998) (table 1). 

• Within a pool complex, the faunal community can vary in richness (number 

of species) and species composition (which species) between pools (King 

and others 1996, Ripley and Simovich unpublished data, Simovich 1998). 

• Many species appear to be highly endemic and are limited to one or a few 

types of pools (Eriksen and Belk 1999, King and others 1996). 

• Sampling over a six-year period has shown that the same pool can differ in

richness and composition between years and over the season within a year 

(Ripley and Simovich unpublished data). Similar variability is seen in

ephemeral pool plants (Holland 1987). Thus, richness and composition vary

spatially and temporally. 

30 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Vernal Pool Fauna—Simovich

Table 1—Summary of crustacean diversity for two surveys. Only three species were found in

both studies. 

 North/Central Calif. Southern Calif. 

 (King and others 1996) (Simovich 1998)

# sites 14 1 

# pools 58 66

# species 67 27 

Range of # species per sites 4-39 NA

Range of # species in single pools 1-27 2-22 

Physiology and Life History

Work done with the more visible vernal pool crustaceans, the anostracans or fairy

shrimp, points to differences in physiological tolerance and life history associated 

with the species distributions. 

• Some species appear to be broadly tolerant of a variety of water chemistry 

conditions. Others however, are restricted in their temperature, pH, 

alkalinity, turbidity, and salinity tolerance (Alexander 1976, Belk and Cole 

1975, Belk and Nelson 1995, Bernice 1972, Brendonck 1996, Brown and 

Carplan 1971, Donald 1983, Eng and others 1990, Forro 1989, Gonzalez and 

others 1996, Hartland-Rowe 1972, Horne 1967, 1971, Lanway 1974, Moore 

1967, Prophet 1963, Sam and Krishnaswamy 1979, Thiéry 1991, or see 

Eriksen and Belk 1999 for a review). This restricts the elevation 

(temperature) and location (soil chemistry) of the pools in which they can

occur (table 2). 

• Some species take several weeks to grow, mature, and reproduce while

others can do so in as little as 10 days (Eng and others 1990, Hamer and 

Appleton 1991, Hamer and Sawchyn 1968, Hathaway and Simovich 1996, 

Mura 1991, Patton 1984, Pennak 1989, Rettalack and Clifford 1980). These 

life history characteristics restrict species’ distributions to areas where

precipitation and pool size/depth result in pool longevity consistent with their

needs. Although precipitation (and consequently pool duration) can vary 

among years, pools must last long enough, often enough for reproduction.

These factors contribute to endemism, beta diversity, and the differences

within pools over time and among years. 

• Some vernal pool organisms do not spend their entire life in the pool. Insect

pollinators and amphibians are dependent on surrounding uplands for most of 

their life cycle. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 31 
Vernal Pool Fauna—Simovich

Table 2—A comparison of three species of fairy shrimp with differing habitat requirements 

(data reviewed in Eriksen and Belk 1999). 

Hatching TDS Pool Example species 

temp. oC ppm duration 

0-15o <300 >3 weeks Eubranchipus serratus

5-20o <600 >3 weeks Branchinecta conservatio

17-30o 300-600 <3 weeks Thamnocephalus platyurus

Incomplete Hatch 

In areas where precipitation is extremely variable, pools fairly frequently may fill and 

dry before some species can reproduce. If all the propagules in the soil hatched in 

such a filling, the species would be extirpated. 

• As an apparent adaptation to such occurrences, fairy shrimp species in 

unpredictable habitats exhibit incomplete hatch (al-Tikrity and Graninger

1990, Belk 1977, Belk and Nelson 1995, Brendonck 1996, Brendonck and 

Persoone 1993, Dexter 1973, Hildrew 1985, Mossin 1986, Philippi and

others 2001, Simovich and Hathaway 1997). As a hedge against failure, only 

a portion of the cysts hatches in a given pool filling. As predicted by theory

(Cohen 1966, 1967, 1968) and demonstrated in the lab (Philippi and others

2001, Simovich and Hathaway 1997), this portion approximates the 

proportion of time the pool fills sufficiently for reproduction; in other words, 

the probability of success. Thus, a cyst bank in the soil is crucial to the

persistence of the population over time. 

• There are several theoretical consequences of incomplete hatch. For example,

if a number of cysts are transported to a new pool, only a small portion will 

hatch. When these reproduce, their offspring (cysts) will dominate the cyst

bank, thus dominating the subsequent gene pool. Essentially there are fewer

actual founders of the population than cysts transported. This can contribute 

to genetic drift and inbreeding. 

Population Genetics

Vernal pools generally occur in clusters or complexes. Distances between pools 

within a complex are generally on the order of meters. Distances between complexes

are commonly on the order of kilometers. 

• As vernal pool crustaceans are obligately aquatic, they must rely on external 

means of dispersal. Within a complex, cysts may move between pools via 

overland water flow in high rainfall years. For dispersal between pools in 

drier years, or between complexes, the species must rely on vectors. The 

cysts can be transported on the legs of grazing animals or through the 

digestive tracts of waterfowl and amphibians (Bohonak and Whiteman 1999, 

Krapu 1974, Maguire 1963, Proctor and others 1967, Reid and Reed 1994). 

Thus, these species are distributed spatially as metapopulations, numerous 

subpopulations linked by varying levels of migration and gene flow. 

• Genetic studies have shown evidence of low gene flow and consequently

high differentiation between populations (Bohonak 1998, Boileau and Her32 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Vernal Pool Fauna—Simovich

bert 1988, Crease and others 1990, Davies and others 1997, Fugate 1992, 

1998; Hann and Herbert 1986, Herbert 1974, King 1996). The narrowly

endemic San Diego fairy shrimp (Branchinecta sandiegonensis) shows low 

genetic variability (allozymes) within subpopulatons and high differentiation 

between populations, even over short distances (FST > 0.50 at less than 50 

km) (fig. 1) (Davies and others 1997). A similar result was found for a Rocky 

Mountain species (Branchinecta coloradoensis) (Bohanak 1998). A more 

detailed analysis shows the differentiation by distance relationships of the 

two species to be almost identical (Bohanak and Simovich unpublished data). 

Figure 1—Isolation by distance. FST versus physical distance for populations of 

Branchinecta sandiegonensis in San Diego County, California (data from Davies and

others 1997). FST can be considered an indicator of differentiation due to genetic drift. 

r=0.57 

Missing Knowledge and Missing Species 

Although some physiological and ecological information is available for some fairy

shrimp species, these data are largely missing for most vernal pool crustaceans and

other invertebrates. 

• It appears that approximately 50 percent of California’s vernal pool

crustaceans are undescribed species about which essentially nothing is 

known (King and others 1996, Simovich 1998). 

• Habitat loss models estimate that approximately 30 percent of California’s

vernal pool crustacean species have already been lost (King 1998). 

• Very little is known about the food web, symbiotic relationships, and nutrient 

and energy cycles of vernal pools. The algae, plants, and invertebrates are 

food for birds and amphibians (Bohonak and Whiteman 1999, Krapu 1974,

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 33 
Vernal Pool Fauna—Simovich

Morin 1987, Proctor and others 1967, Simovich and others 1991). Some

anostracans consume mostly algae while others are predators (see for review

Parsick 2002). However, there are no studies available for most vernal pool 

invertebrates.

Additional information 

• Anostracan cysts crush very easily, especially when wet (Hathaway and 

others 1996).

• Commonly used chemicals such as Malathion can kill vernal pool 

crustaceans (Davis and others 2001). 

• At least some cysts (Wells and others 1997) and seeds (Cox and Austin 

1990) in pools survive after fire. 

Management Implications 

These findings have important implications for management including habitat 

preservation, habitat restoration, and species relocation efforts. 

Diversity and Endemism 

• Pools at different sites or of different types differ in richness and 

composition. Therefore, preserving pools at one site as mitigation for losses 

at another may not preserve all of the same species. 

• Pools within a site differ. Therefore, preserving a few pools at a site while 

losing the others may not preserve all species. 

• Pools can differ between years in richness and composition. Therefore, more

than one season of pre-manipulation surveying is necessary to establish baseline conditions, and several years of post-manipulation monitoring are

necessary to establish success. 

Physiology and Life History

• Many species, particularly the narrow endemics that are generally the species

of concern, have very specific requirements for water chemistry,

temperature, and pool longevity. Thus restoration or habitat creation must 

result in pools with appropriate chemistry and hydrology that mimic the 

range and variation of natural pools. 

• Any manipulation involving the transport or introduction of organisms must 

consider the appropriate source. Organisms from different areas or pool types

will probably not be appropriate and may not establish populations. 

• Preservation of vernal pool communities requires preservation of the 

surrounding uplands for watershed and for amphibian, pollinator, and vector 

habitat. 

34 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Vernal Pool Fauna—Simovich

Incomplete Hatch 

• Although incomplete hatch provides a hedge against cycles with low

reproduction, this life history strategy complicates simple management 

decisions. For example, after inoculation of a created or restored pool with 

soil, shrimp could hatch over several hydrations from the cysts in the 

inoculum. However, if the pool did not last long enough or hydrate often 

enough for sufficient reproduction, the cyst bank would not be replenished. 

Thus, initial monitoring would suggest a successful restoration, while in

reality the cyst bank was being depleted to extirpation. Consequently, 

successful reproduction must be documented over several years to establish

success. 

• Incomplete hatch of the cysts used to inoculate pools may result in lower 

genetic variability of the new population due to genetic drift. This can have 

adverse effects on population fitness and evolutionary potential over time. 

Use of inocula with low cyst density may compound the problem. 

Population Genetics

Particularly with narrow endemics, such as the San Diego fairy shrimp (B. 

sandiegonensis), genetic variation within populations and gene flow among 

populations (as indicated by allozymes) is apparently low. 

• Populations should be surveyed in order to preserve those with higher genetic 

variability, which could, and may now, serve as source populations for gene 

flow and colonists for other populations.

• Increased isolation of pools will reduce visitation by vector organisms. This 

will further reduce gene flow and subsequently result in further reductions in

genetic variation. This would be expected to put the species in jeopardy of 

reduced fitness due to inbreeding depression. It would also be expected to 

reduce evolutionary potential and the ability of the species to adapt over time 

and to reduce the possibility of rescue (re-colonization) if a population is 

extirpated or reduced. Consequently, preservation efforts must be considered

at the landscape level rather than pool by pool. Sufficient buffer and corridor 

access must remain to encourage the usage of the area by potential 

distribution vector species. 

Missing Knowledge and Species 

• If an organism is undescribed, it is difficult to make decisions necessary to 

preserve and protect it. Taxonomic description, surveys, and natural history

studies to establish species’ distributions and resource requirements are

critical. 

• Any management or manipulations must focus on the entire community, not

simply a single species. If a species’ food is missing, or for plants, if a 

pollinator is missing, the species simply will not survive long. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 35 
Vernal Pool Fauna—Simovich

Additional information

• Vehicles should be kept out of pools. 

• Pollution run-off from surrounding areas is a potential problem. 

• Fire should not be excluded from fire-type habitats, but excessive unnatural 

fuels in a pool (weeds) may increase fire temperatures and be a problem for

cyst survival. 

Summary 

When trying to preserve communities about which little is known, careful decisions 

based on the best scientific knowledge available and a clear understanding of population ecology and genetics are necessary. Preservation decisions must be at the landscape level and consider functional communities, not just individual species or single 

pools. For restoration attempts, a careful, scientific approach documenting pre- and 

post-restoration conditions and based on well-reasoned experimental design is critical. This can best be accomplished by closely linking research and management. 

Finally, results must be published, to increase our knowledge, increase the 

probability of making good decisions, and decrease the probability of repeating poor 

ones. 

Acknowledgments 

My thanks to Sharon Bunton, Diolinda Parsick, Janette Holtz, Mike Wells, Denton 

Belk, Andy Bohanak, Tom Philippi and Ellen Bauder. Much of the author’s work

referenced was supported by grants from the Department of Defense and the 

University of San Diego. 

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USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 41 

Biodiversity of Mycorrhizal Fungi in 

Southern California1

Michael F. Allen2, Louise Egerton-Warburton, Kathleen K. 

Treseder, Cara Cario, Amy Lindahl, Jennifer Lansing, José 

Ignacio Querejeta, Ola Karen, Sharon Harney, and Thomas Zink 

Abstract 

Mycorrhizal fungi are a diverse group of organisms and form an integral link between plants

and soil. These symbiotic fungi provide most of the soil resources to the plant in exchange for 

energy. However, their very diversity means that they can play many different roles and serve

as indicators of changing ecosystem dynamics. There are probably more than a hundred

mycorrhizal fungi associated with coast live oak alone in southern California. Data indicate 

that both the composition and functioning of these fungi change with the changing regional

environment. Further, as regulators of ecosystem processes, changes in mycorrhizal

composition and activity can ramify through the entire community. Understanding

mycorrhizal and soil microbial dynamics and incorporating this understanding into

management strategies will be crucial for sustaining viable preserves. 

Key words: diversity, ecosystem, fungi, global change, mycorrhiza 

Introduction 

Biodiversity is a broad and general term used to describe the variety of organisms

inhabiting a geographical or topological region of interest. For the most part, the

emphasis of biodiversity focuses on listing or describing organisms of interest for 

protection or utilization, or the extent of this variation in terms of species number and

abundance (Sisk and others 1994). However, there are vast arrays of organisms about 

which we know very little. According to recent estimates, we have described around 

50 percent of vertebrates, but only 5 percent or less of insects and fungi (for example, 

Fröhlich and Hyde 1999, Hughes and others 2000). For prokaryotic organisms, we 

don’t even know how to classify kingdoms, much less the concept of species. Of 

those we know, many are threatened with extinction, often before we even know who 

they are or their roles in ecosystems (Hughes and others 1997). 

Fungi make a useful test group for understanding the concept of biodiversity

and, in turn, understanding the environmental changes likely to affect the entire biota 

of a system (E. Allen and others 1995, M. Allen and others 1995). Fungi are found in

every terrestrial environment and are directly and indirectly exposed to every 

perturbation. They range in scale from microfungi growing in solitary colonies on a 

single root tip to a macroorganism, such as Hydnangium carneum, that extends

across an entire plant stand (Allen and others 1993). They also perform many

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2 Center for Conservation Biology, University of California, Riverside, CA 92521-0334 (email address: 

michael.allen@ucr.edu)

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 43 
Biodiversity of Mycorrhizal Fungi—M. Allen and others 

essential roles in ecosystems, such as decomposing detritus and providing much of 

the nutrients needed by plants for photosynthesis. Thus, they are susceptible to every 

form of environmental change. 

Of the vast array of fungi, mycorrhizal fungi represent a tractable group that is 

also sensitive to environmental perturbations. Mycorrhizae are mutualistic

associations between plants and fungi that are localized within the root or root-like 

structures (Allen 1991). These fungi extend not only outward into the soil to extract

nutrients and water, but they also extend into the plant roots where they exchange 

those resources for carbon. The only terrestrial plants not forming mycorrhizae are a 

limited number of weedy annuals in a few families such as the Chenopodiaceae or

Cruciferae (Allen 1991). Thus, any environmental factor that affects plants can also 

affect these fungi. Likewise, each change in the substrate that could influence 

mycorrhizal fungi will also affect the host plants. 

Mycorrhizal fungi are interesting and important in themselves. Many are 

commonly collected mushrooms for culinary purposes. They also make an important 

contribution to the diets of many mammals, especially rodents. Truffles are common, 

and searching for them is largely based on finding animal diggings. The hyphae of 

mycorrhizal fungi also provide much of the soil structure by producing glycoproteins 

called glomalin, which may make up as much as 25 percent of the soil organic matter 

and bind soil particles into macroaggregates (Rillig and others 1999, Treseder and

others 2003). This structure protects nutrients and increases porosity and waterholding capacity. Finally, mycorrhizal fungi are responsible for much of the nutrient 

turnover and most of the transport dynamics that move resources around plants and 

between patches. 

We have been studying mycorrhizal fungi in southern California vegetation 

types for over a decade. Such studies are vital because southern California is a 

designated hot spot for biodiversity conservation (Myers and others 2000), and the 

region is subject to intense perturbation from a variety of anthropogenic sources. Our 

observations have demonstrated some unique patterns (described below), but, of 

greater importance, they have indicated the need for long-term observations before

definitive trends can be conclusively determined. We have focused on observations 

extending through precipitation cycles and across edaphic gradients at a variety of

scales and on experiments evaluating responses to changing environments. Based on

these data and the measured environmental change already observed by physical 

scientists, we believe that southern California ecosystems are likely beginning to 

undergo important and quantifiable change in microbial communities. Here, we focus 

on measurable responses of mycorrhizal fungi. These data can form a baseline 

against which future changes can be measured. 

Mycorrhizal Fungi in Southern California: Biodiversity of 

a Group 

Mycorrhizal fungi are diverse in southern California. Little is known about some key 

variables that would be involved in any program to describe and conserve biodiversity in mycorrhizal fungi. Although seven groups of mycorrhizal fungi are recognized, we focused largely on two major groups of mycorrhizal fungi. The first are 

ectomycorrhizal fungi found in symbiosis with coast live oak (Quercus agrifolia),

pines, some other trees, and a few shrubs; ascomycetous and basidiomycetous fungi

44 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Biodiversity of Mycorrhizal Fungi – M. Allen and others 

form these mycorrhizae. The second are arbuscular mycorrhizal fungi (Glomales)

associated with other trees, forbs, most shrubs, and grasses. 

Structure of Diversity in Fungi 

Diversity is richness and evenness of species, genomes, and functions. Fungi are different from higher plants and animals in that variation can occur and be measured

within an individual as well as between individuals, populations, species, and clades. 

This variation can be detected at the scale of a small patch of ground (cm2

) as well as 

at larger land units. Variation can also occur through time. As environmental pressures change, not only species but also the relative genetic composition of an

“individual” mycelium can change. 

Characterizing both the current and anticipated changes in fungal diversity with 

environmental perturbation is challenging. Fungi have multiple, haploid (1N) nuclei. 

In zygomycetous and glomalean fungi, no cross-walls exist. Thus, groups of nuclei

can be found scattered throughout a developing mycelial fan. In ascomycetous and 

basidiomycetous fungi, most cells contain two separate 1N nuclei. The hyphae

expand as individual cells multiply and elongate. However, when two compatible

hyphae meet, they anastomose, forming a single mycelial network. As the mycelial

network extends outward and multiple hyphae connect, a single connected “individual” mycelium can contain many different nuclei, and an individual connected hypha 

can contain cells with different functional and sometimes incompatible genes. 

Ectomycorrhizal Fungal Diversity 

Ectomycorrhizal (EM) fungi are a highly diverse group. Based on a decade of collecting from San Diego State University mycology classes, we have identified 74 

species associated with coast live oak groves. Nevertheless, how diversity is characterized is as important as the data collection itself. To better evaluate this diversity, 

we have utilized a number of approaches. Because the real extent of demographic 

units in mycorrhizae is essentially unknown, as is the frequency of occurrence of

metapopulations, we used geographically and ecologically stratified samples to

evaluate patterns of both species and population diversity. 

Our experimental research has been undertaken largely at Lopez Canyon, in the

Western Riverside County Multispecies Reserve, and consists of evaluation of fungal

diversity at the morphological and molecular level. Our first assessment was to plant

oak seedlings in association with mature individuals (> 100 years old). At the end of

the growing season, we harvested seedlings and estimated the number of 

morphotypes. Morphotypes are morphologically distinct mycorrhizal root tips that 

can be distinguished on the basis of characteristics such as texture, color, and hyphal 

types. Morphotypes sometimes correspond to fungal taxa, although there can be more 

than one fungal taxon per morphotype. Because of this ambiguity, we also employed 

molecular differentiation of tips that were morphologically similar. DNA was then 

extracted and amplified from roots of some seedlings using the polymerase chain 

reaction (PCR) of the internal transcriber region (ITS) of the small subunit of the 

ribosomal DNA. This part of the DNA chain was cut using endonuclease enzymes,

leaving distinctive small units of DNA. These were run on gels that showed the size

patterns of these DNA fragments, called restriction fragment length polymorphisms

(RFLPs). These size units are distinctive to species and thus can be used to assess 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 45 
Biodiversity of Mycorrhizal Fungi—M. Allen and others 

diversity of mycorrhizal fungi at each individual root tip scale (Gardes and Bruns 

1996, Karen and others 1997). RFLP comparisons allowed us to assess the numbers 

of actual mycorrhizal fungal taxa per tree based on the composition of the DNA. 

During the winter rainy season of 2001, we found that the richness of EM morphotypes per seedling associated with mature coast live oak averaged 10; however,

the richness at the stand level was expected to be much greater. There is considerable

variability in the EM composition of seedlings from different oak patches resulting in

much greater stand richness. Our initial findings showed that it was common for only 

a few EM morphotypes to be the dominant colonizers on an individual seedling. 

Furthermore, not only was there variability in the EM composition between seedlings

from different patches, but also in which EM morphotypes were the dominant 

colonizers. 

When we expanded to the RFLP analysis, a greater number of types emerged. 

For comparative purposes, an average of seven morphotypes per seedling were 

found, in contrast with an average of 12 RFLP taxa per seedling and a maximum of

20 RFLP taxa per seedling. The same pattern as morphotypes per stand emerged, 

though. A few dominant taxa could be found, with many infrequent species. 

This finding is similar to EM community studies of pinyon pine (Pinus edulis) 

in Arizona (Gehring and others 1998) and New Mexico (Allen and others 2002). Pinyon pine trees act as isolated islands in which an individual tree has an average richness of 5 morphotypes but a stand of 8 trees has a richness of 20 to 30. Similarly, in 

coast live oak, only a few dominants per tree were found, but there was a great deal 

of variability among trees. As we sampled across the landscape, there was a 

continued replacement of species demonstrating a high level of β diversity (Allen and 

others 2002).

Our final effort was to investigate the diversity of species identified by sporocarps and the relationship between diversity and a stochastic measure of the rate of 

change in mycorrhizal diversity on a per area basis. At Camp Pendleton, we undertook a 4-year study monitoring the EM fungi fruiting in plots under stands of coast 

live oak and Engelmann oak. During this period, we identified 58 species of EM 

fungi. The species increment curve demonstrated a leveling off, suggesting that we 

had identified most of the sporulating fungi for those sites. 

We estimated the land area for sampling, allowing us to calculate a z-value that 

describes the rate of change in diversity in relation to sample area size (from s=caz

, 

where s is the number of species, or richness, c is a constant, and a is the area,

Rosensweig 1995). Although we recognize that there are many means to assess 

diversity, the z-value is independent of size and allows comparisons to be made

among habitats with varying areas. Since we are only able to assess mycorrhizal

fungal responses to environmental change on small plots, this relationship provides 

some insights into how our region and mycorrhizal fungi might be affected. 

Our results show that the z-value may be appropriate for comparison. For the 

individual seedling studies at Lopez Canyon, the z-value for root-tip morphotypes 

was 0.57 with an r2

 for the species area curve of 0.96. For the Camp Pendleton sporocarp assessment, the z-value was 0.58, with an r2

 of 0.975. These values are high 

compared with vertebrates, but they resemble the high values found in insect

communities. Given that this z-value is high at the scales of patches and sites, then as

the entire range of coast live oak is examined, the diversity of fungi has the potential

to become extraordinarily high. This suggests that there probably are, in reality, up to

46 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Biodiversity of Mycorrhizal Fungi – M. Allen and others 

several hundred taxa associated with coast live oak alone, as has also been found for 

other mycorrhizal plants (Allen and others 1995, Sanders and others 1995, Sharples 

and others 2000,). Thus, the potential biodiversity of EM fungi and richness of

valuable sources of genetic variation for a group of organisms important to the

survival of this tree are extraordinary.

Chamise (Adenostoma fasciculatum) normally forms arbuscular mycorrhizae 

with all genera found in the region. However, during wet years (El Niño), we found

EM associated with its roots and EM fungi in the stands (Allen and others 1999b). 

There was a high diversity of fungi ranging from Cenococcum and Balsamia spp. 

(ascomycetes) to a variety of basidiomycetes such as Pisolithus sp., Cortinarius spp., 

and Hysterangium separabile. We also found a new species of Rhizopogon: R.

mengei (Allen and others 1999a). This is an important finding, since all other known 

species of Rhizopogon are associated with conifers. In addition, we sequenced the 

dominant fungus found on the root tips of chamise, and this fungus was an unknown 

species of ascomycete, most closely related to Sarcocypha emarginata (97 percent 

similarity in the sequence alignment of the 5.8S region, 63 percent similarity in ITS1, 

and 59 percent similarity in ITS2), a common fire-following fungus. There are 

clearly plant/fungal symbiotic combinations that we do not understand and have yet

to explore. 

Although we have focused on coast live oak and chamise, it must be noted that

there are many other EM plants in southern California. These include all other 

species of oaks, all members of the Pinaceae, and chinquapin (Chrysolepus sempervirens). In our region, these never have been surveyed comprehensively. 

Endomycorrhizal Fungal Diversity 

Almost all other plants in southern California form endomycorrhizal associations. 

Most common is the arbuscular mycorrhizal (AM) type found in grasslands and 

shrublands. The fungal symbiont is a monophyletic group consisting of two distinct 

clades (phylogenetically-related species), the Glomaceae and the Gigasporaceae,

which split with the earliest land plants (Redecker and others 2000, Morton and 

Redecker 2001). These different clades have some quite distinct characteristics

important to mycorrhizal functioning and response to changing environments. Thus, 

while they are monophyletic, and species diversity is low (approximately 170 species 

are described worldwide), the different families diverged so long ago (350-400 

million years) that functional differences have resulted (Hart and others 2001). 

We conducted two relevant types of surveys. In the first, we examined the AM 

fungi associated with basin big sagebrush (Artemisia tridentata ssp. tridentata) 

throughout its distribution. This subspecies can be found throughout western North

America, ranging north from Baja California in Mexico to British Columbia in 

Canada and eastward as far as North Dakota and New Mexico. Using this large 

spatial scale, we found that species richness of AM fungi increased with greater area 

sampled (fig. 1). However, most of the diversity of AM fungi associated with the 

Mojave ecotype of big basin sagebrush could be accounted for in just seven sites (fig. 

2), because evenness actually declined as more sites were surveyed. 

If we examine the pattern of distribution, there are several fungi—Glomus 

aggregatum, Scutellospora calospora, and Paraglomus occultum—that are widely 

distributed and found throughout the range. Others, such as Sclerocystis rubiformis, 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 47 
Biodiversity of Mycorrhizal Fungi—M. Allen and others 

Glomus tortuosum, Archaeospora gerdemannii, Scutellospora pellucida, and S.

heterogama, are narrowly distributed, being found in only a few locations. In

expanding to all 99 sites where basin big sagebrush exists, the total richness (number

of AM fungal taxa) doubles, but the diversity continues to decline due to a lowered 

evenness. Clearly, each area has its specific combination of AM fungal species. The 

plant likely adapted to the local AM fungi but probably did not carry the AM fungi

with it as it dispersed. 

0

10

20

30

40

50

Southern CA plus Owens

Valley

plus Baja CA plus Central

Valley

plus SW UT allrange

cumulative locations

Species Richness

Figure 1—Richness of AM fungi associated with basin big sagebrush from southern

California across its range into central California, Baja California, and across the 

Great Basin. 

0

0.5

1

1.5

2

2.5

3

Southern

CA

plus

Owens

Valley

plus

Baja CA

plus

Central

Valley

plus SW

UT

all range

cumulative locations

Diversity and Evenness

H'

J'

Figure 2—Diversity (H’) and evenness (J’) of AM fungi associated with basin big

sagebrush from southern California expanding across its range into central California, 

Baja California, and across the Great Basin. 

Just as importantly, the same genera and most of the same species diversity can be

found associated with California sagebrush (Artemisia californica) and probably most 

other shrublands (Egerton-Warburton and Allen 2000). We do not yet have a large-scale

sampling of AM fungi associated with forblands and only limited sampling in perennial

Stipa grassland. However, in an undisturbed Stipa grassland, we found species of 

Glomus, Acaulospora, and Scutellospora (Allen unpublished data).

48 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Biodiversity of Mycorrhizal Fungi – M. Allen and others 

We know that, with few exceptions, AM fungi are not species-specific, but 

rather function as generalists and tend to infect whatever plant is nearby. We also

know that there are growth responses that occur preferentially between specific 

combinations of plants and fungi (Klironomos and others 2000). This leads to 

increasing stand productivity and plant diversity in response to increasing AM fungal

diversity (Klironomos and others 2000, van der Heijden and others 1998). However,

the specific combinations that are crucial to a native plant community remain elusive. 

Projections of Environmental Change: CO2, N Pollution, 

Temperature, and Precipitation 

Our global environment is changing dramatically and rapidly. This change translates

to direct effects on our local environments and the need for retaining and managing 

for biodiversity on the remaining open space reserves. The clear changes are in 

atmospheric CO2 and in NO3 and NH4 deposition. Less clear but potentially 

important are changes in temperature and precipitation (Rillig and others 2002).

CO2 has risen dramatically since the industrial revolution, largely in response to

increasing human demands for energy and increasing conversion of forest to 

agricultural lands globally (Houghton 1997). During the last ice age, atmospheric 

CO2 was below 200 ppm (parts per million). By the beginning of the industrial 

revolution, atmospheric CO2 was around 250 ppm. In the 1970s, it was up to 350

ppm and has rapidly increased, being currently about 370 ppm. Global projections 

range from 550 to 750 ppm over the next century. CO2 has direct effects on plant 

dynamics in that the need for a C sink increases to handle the increasing fixation. 

Further, nutrients become more limiting to photosynthesis, especially N and P.

Finally, with elevated CO2, water-use efficiency rises. 

Increased N deposition is also dramatic but concentrated at the regional rather 

than global scale. In southern California N largely comes from denitrification

associated with dairy cattle and vehicular exhaust. Although the levels and forms are 

changing dramatically (see E. Allen and others, this volume), the amount remains 

important. Within the Los Angeles basin, deposition can range from 30 to 45 kg

ha-1

y

-1, values equal to agricultural fertilization levels. 

Temperature and precipitation changes are much less predictable. However, 

only small changes can have dramatic impacts. Further, these can remain within the

bounds of evolutionary history and still affect the regional biota. According to most 

of the global circulation models (GCM), our region of southern California may be in

the area of maximum climate and vegetation change (Watson and others 1997).

Virtually every GCM shows increasing precipitation in southern California as a 

function of general warming caused by the greenhouse effect. With the heating of the

oceans associated with the greenhouse effect, the model predicts increasing

precipitation in a manner similar to the observed increases associated with El Niño 

phenomena. In this scenario, Neilson and colleagues (1998, and 2001 pers. comm.)

note a significant increase in both winter and summer precipitation. This increase is

coupled to increased water-use efficiency of desert plants associated with elevated 

CO2. Consequently, their projections suggest that this increasing precipitation would 

result in increasing grasslands and mixed forest and decreasing shrublands and 

deserts. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 49 
Biodiversity of Mycorrhizal Fungi—M. Allen and others 

Responses of Fungi to Change in Southern California 

Although all of these projections fail to include all relevant processes, clearly there 

are enough indicators for us to be concerned with retaining our current biodiversity 

and managing for the diversity that can protect species of concern. Mycorrhizal fungi 

are only one group of organisms on which the ecosystem depends. Likely, there is 

more diversity in these fungi than absolutely required for oaks and similar critical

species to survive. However, we are not very good at predicting which might be the 

crucial species. Below, we note some changes that we can see from our on-going 

experiments. 

Diversity

Diversity and composition responded to changing perturbations. In experimental 

studies, the total amount of AM increased in response to elevated CO2, and the 

increase was largely associated with the Gigasporaceae (Klironomos and others 1998, 

Treseder and others 2003; fig. 3). The amount of Glomaceae did not change even as 

CO2 increased from 250 ppm to 650 ppm. At 750 ppm CO2, the entire belowground

system collapsed, and we do not understand the possible mechanisms. Conversely, in 

response to N deposition, sporulation in the Gigasporaceae disappeared, although 

some vegetative hyphae remained in association with roots (Egerton-Warburton and 

Allen 2000). However, the Glomaceae remained relatively constant; moreover, some 

species of Glomus increased in abundance following N enrichment. These findings

indicate that N deposition promotes editing of the existing mycoflora, not an

invasion. It should be remembered that AM fungi evolved during a period when 

atmospheric CO2 was much higher than today (Allen 1996). CO2 per se is not likely a 

detrimental factor for AM fungi. However, N has generally been the major limiting 

element in most terrestrial ecosystems, and N fertilization can have rather dramatic 

effects on AM fungal dynamics (see E. Allen and others, this volume). 

There is little field experimental evidence on the interactions of CO2 and N 

deposition and the composition of the AM fungal community. In the only data set to 

date, we find that there was a shift from the normal mix of AM fungi to the fineendophyte form, Glomus tenue (Rillig and others 1999a). This may be a direct effect

or a response to changing soil pH. 

The abundance of ectomycorrhizae initially increased in response to N 

fertilization over a single growing season at Lopez Canyon. This area has not 

received long-term high levels of N deposition, as has the Los Angeles Basin. The zvalues were 0.57, 0.68, and 0.60 (not significantly different) for control, N-fertilized,

and weeded oak seedlings, respectively. In part, this demonstrates that in this region 

N remains the most limiting nutrient. N fertilization increased plant survival and 

growth, allowing a greater rate of both tip and EM formation. Also, these data

demonstrate the need for long-term data on the effects of perturbations such as N 

deposition. 

To our knowledge, no data on responses of EM fungal communities to elevated

CO2 have been published. This is important because increasing CO2 could push

ecosystems in high N depositional regions back to C/N equilibrium or propel the 

ecosystem along an entirely new trajectory.

50 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Biodiversity of Mycorrhizal Fungi – M. Allen and others 

AM infection

Glomaceae 

Gigasporaceae

Elevated N Ambient Elevated CO2

Figure 3—Changing AM activity along an atmospheric change gradient, ranging from

elevated N (N deposition) to elevated CO2. Data come from Egerton-Warburton and 

Allen (2000), Rillig and others (1999a), and Treseder and others (2003). 

Mycorrhizal Activity 

CO2 and N have many impacts on mycorrhizal fungi and the functioning of soil 

ecosystems. CO2 in particular has a very subtle effect on arbuscular mycorrhizae. In 

studies of newly disturbed soil, or pot cultures, AM fungi always appear to increase 

(Treseder and Allen 2000). However, in in situ studies, the total hyphal lengths per 

soil volume do not change (Allen in preparation). The hyphal lengths in soil 

aggregates and allocation of new C to soil via aggregates does increase as measured

using both isotopic data and glomalin concentrations (Rillig and others 1999b, 

Treseder and others 2003). These data indicate that more C is allocated to 

mycorrhizae but that the turnover also increases. The residual C in the form of

recalcitrant carbon builds up but probably not at the same rate as the increasing 

atmospheric levels that we are currently experiencing. 

There are no data on the effects of elevated CO2 on oak mycorrhizae from 

southern California. However, in a similar climate in Italy, ectomycorrhizal activity

tends toward increasing cycling rates in a manner observed for AM (Cario and Allen 

in preparation). 

N fertilization has mixed effects on total mycorrhizae. In N deficient situations, 

N fertilization can actually increase mycorrhizae. However, N can rapidly saturate

shrublands and forb communities, causing a decline in the abundance of mycorrhizae 

often within three years following the onset of N enrichment (Egerton-Warburton and 

Allen 2000). 

In forest ecosystems, there is a large capacity for absorbing N. We contrasted N

cycling in oaks, comparing two sites: the first was the Sky Oaks Field Station, 

managed by San Diego State University, near Warner Springs, CA (N limited and

low deposition); second was the San Dimas Experimental Forest (USDA Forest 

Service) (high deposition). The δ15N data from San Dimas shows that N cycling has 

changed markedly since the experiment station was established in the 1930s. The

values from San Dimas in 1935 (-2 percent) are equivalent to Sky Oaks, suggesting 

that Sky Oaks N-cycling patterns are less impacted by recent N deposition. Current 

leaf δ 15N is -4 percent. The values for the EM sporocarps for Sky Oaks and San

Dimas average 6.9 percent and 6.8 percent, respectively. These data suggest that the 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 51 
Biodiversity of Mycorrhizal Fungi—M. Allen and others 

functioning of the mycorrhizal fungi (utilizing NH4

+

 and organic N) may not have

changed but that depositional N is taken directly by the plant. In oak stands, initial

evidence suggests that the oaks are not N saturated and mycorrhizal fungi remain 

active. How long this will persist and whether the fungal composition changes as the 

plants depend less on mycorrhizae for N are open questions. 

Exotic Fungi 

Pisolithus tinctorius is a widespread fungus associated with many host plants. 

Because it is so widespread, it was also widely used in reforestation efforts and in 

urban plantings (Marx and others 1984). (It is important to note that when this project 

began, it was considered a single species, but recent systematic studies are in the 

process of dividing P. tinctorius into several new species.) We studied the diversity

of several sporocarps of P. tinctorius. Of relevance to us, this fungus is associated 

with oaks in California, pines in the southeastern US, and eucalyptus in Australia.

We wanted to determine whether the fungus associated with cultivated trees was 

different from fungus associated with eucalyptus or from the native oaks. 

On the basis of morphology, the sporocarps were virtually indistinguishable, and

the spores ranged in size from 9.3 to 10.3 μm in diameter. However, the median 

spore diameter did vary significantly among the different isolates. The ITS region 

was also examined to determine whether the molecular structure could be 

distinguished. Even without further RFLP analyses it was possible to discern three 

ITS size categories: 610, 635, and 650 base pairs. When the spore size and ITS 

regions are examined together, the different sporocarps could be broken into six

distinct groups (table 1). Importantly, there are four groups of sporocarps primarily

found with exotic trees, two in the city of San Diego, one at the Santa Margarita 

Ecological Reserve (planted in 1974 as part of a biomass production study), and a 

third mixed planting near the Sky Oaks site with eucalyptus. There are two distinct 

genotypes found with native trees. The first was found in southern California with 

Quercus agrifolia and with Q. turbinella in New Mexico. This second is found in 

association with many native plants and stands scattered across California. 

The important issue here is that there appear to be two native groupings and four

likely introduced exotic fungi, which may switch hosts. Although we focused on P. 

tinctorius, there are many other exotic mycorrhizal fungi found in eucalyptus groves, 

including Hydnangium separabile from Tasmania, a fungus in the Laccaria 

amethystina group from Australia (Allen and others 1993). At this time, we do not 

know the implications of this invasion, but in other studies, switching can result in a

less effective, or even detrimental, symbiosis for the plant. 

Implications for Southern California Ecosystems 

Mycorrhizal fungi are absolutely critical components of all terrestrial ecosystems in 

southern California and globally. These organisms provide the majority of nutrients 

to the plants, stabilize soils, sequester carbon, provide food for animals, and serve as 

the predominant microbial C sink for fixed photosynthate. They are extremely

diverse with the differing fungi undertaking slightly different roles in the highly 

spatially variable semiarid to arid ecosystems of our region. Most global change

models predict that our region will likely experience profound changes, potentially 

within our lifetimes. 

52 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Biodiversity of Mycorrhizal Fungi – M. Allen and others 

Table 1—Differentiation of taxa of Pisolithus tinctorius from collections in Southern

California based on spore size and ITS size fractions. 

Group Isolate Host ITS (bp) Spore size

(um) Location

I Pt 4 Eucalyptus sp. 610 9.3 SMER (Riverside County) 

Pt 2 Eucalyptus sp. 610 9.6 SMER (Riverside County) 

Pt 22 Pinus radiata 610 9.6 San Diego State University

II Pt 16 Eucalyptus sp. 610 9.8 Sky Oaks (San Diego Co.)

Pt 17 Eucalyptus sp. 610 9.8 Lake Skinner (Riverside Co.)

III Pt 24 Quercus turbinella 635 9.2 Sevilleta LTER, NM 

Pt 7 Quercus agrifolia 635 9.4

Boden Field (Sky Oaks, San 

Diego Co.) 

IV Pt 23 Eucalyptus sp. 635 9.6 San Diego State University

V Pt 11 Quercus agrifolia 635 9.7 Sky Oaks (San Diego Co.)

Pt 1 Chrysolepis sempervirens 635 9.9

San Bernardino National

Forest 

Pt 3 Chrysolepis sempervirens 635 9.9

San Bernardino National

Forest 

Pt 6 Chrysolepis sempervirens 635 9.9

San Bernardino National

Forest 

Pt 14 Quercus agrifolia 635 10 Sky Oaks (San Diego Co.)

Pt 18 Adenostoma fasciculatum 635 10 Lake Skinner (Riverside Co.)

Pt 10 Quercus agrifolia 635 10.1

Boden Field (Sky Oaks, San 

Diego Co.) 

Pt 13 Quercus agrifolia 635 10.3 Sky Oaks (San Diego Co.)

VI Pt 19 Eucalyptus sp. 650 9.4 San Diego State University

Pt 8 Eucalyptus sp. 650 9.6 SMER (Riverside County) 

Pt 15 Quercus agrifolia 650 9.6 Sky Oaks (San Diego Co.)

Pt 20 Eucalyptus sp. 650 9.7 San Diego State University

Pt 9 Eucalyptus sp. 650 9.9 SMER (Riverside County) 

Pt 21 Eucalyptus sp. 650 9.9 San Diego State University

Pt 5 Eucalyptus sp. 650 10 SMER (Riverside County) 

Pt 12\ Quercus agrifolia 650 10 Sky Oaks (San Diego Co.)

According to GCMs, our region generally will experience increasing

precipitation (IPCC 1998). The impacts of this potential change can be seen in the

effects of El Niño years on ecosystem dynamics. Moreover, the changing 

precipitation interacts non-linearly with N deposition, elevated CO2, human

population growth, and increasing development in many and complex ways. Our

ability to manage the diversity and integrity of wildland ecosystems may depend on 

our understanding the microbial groups such as mycorrhizal fungi and sustaining

their diverse and differing needs (Daily and others 2000). 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 53 
Biodiversity of Mycorrhizal Fungi—M. Allen and others 

Acknowledgments 

These studies were funded by the National Science Foundation Biocomplexity and 

Ecosystem Studies Programs, the Shipley-Skinner Endowment Fund for the MultiSpecies Habitat Reserve, and Camp Pendleton Marine Corps Base. We also thank the

Base, San Diego State University, and the USDA Forest Service (San Dimas

Experimental Forest) for permission to undertake much of this research. 

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56 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub: 

What 70-Year-Old VTM Plot Data Tell Us 

About Southern California Shrublands1

Robert S. Taylor, Jr. 2 

Abstract 

This is a preliminary report on a project to characterize the historic species composition of

coastal sage scrub vegetation in southern California using plot data collected between 1929

and 1934 by the USDA Forest Service's Vegetation Type Map (VTM) survey. Records and

maps from a large historic data set collected for the VTM survey were converted into a digital

database and GIS coverage. Historic data for 1481 shrub plots were combined with climate, 

modern land use, and land cover data. Preliminary analysis found patterns of species 

composition and geographic distribution in coastal sage scrub considerably more detailed than 

have been described in previously published regional descriptions. Coastal sage scrub

associations described in local studies are here shown to occur over larger areas in the 

southern California region than the areas within which they were originally described. Maps

of the regional distributions of these coastal sage scrub associations were produced. Summer

precipitation, elevation, and minimum winter temperature best predicted relative differences

in abundance of Eriogonum fasciculatum and Artemisia californica. Salvia apiana preferred 

interior, higher elevation sites with high annual precipitation and an extremely continental 

climate. Salvia mellifera’s distribution was not well predicted by any environmental factors

measured in this study. Rates of coastal sage scrub conversion were estimated by county and 

regionally for a number of species associations. 

Key words: classification, climate, coastal sage scrub, gradient analysis, ordination, vegetation associations

Introduction 

Coastal sage scrub is a drought-deciduous, soft-leafed shrub plant community

frequently dominated by Artemisia californica, Eriogonum fasciculatum, and/or

Salvia species (Mooney 1988). This vegetation has been drastically reduced in extent 

and highly fragmented by conversion to agricultural, industrial and residential land

use, grazing pressure, altered fire regimes, and air pollution. How much coastal sage

scrub has been lost is a matter of some contention. Westman (1981a) estimated that

perhaps 10-15 percent of the original distribution remained. Minnich (1998)

maintains that the Los Angeles basin contained very little of this vegetation at first

Spanish contact and suggests that a much higher fraction remains extant today.

Freudenberger and others (1987) concluded that the Los Angeles basin was once a 

mosaic of shrublands and grasslands, suggesting a number somewhere between 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 Biogeographer/GIS Specialist, Coast Mediterranean network, National Park Service, Santa Monica 

Mountains National Recreation Area, 401 W. Hillcrest Dr., Thousand Oaks CA 91360-4207 (e-mail: 

robert_s_taylor@nps.gov) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 57 
A New Look at Coastal Sage Scrub—Taylor 

Westman’s and Minnich’s opinions. O’Leary (1995) cited consultant Michael 

Brandman Associates’ estimate that as of 1991, 44 percent of historic coastal sage 

scrub remained. All agree that large cumulative losses of coastal sage scrub occurred

during the 20th century and that loss is ongoing. 

The remnants provide habitat for nearly 100 species of plants and animals

currently classified as rare, sensitive, threatened or endangered by Federal and State 

agencies (Davis 1994). Since 1990, coastal sage scrub in coastal areas from Los 

Angeles to San Diego has been the target of intensive conservation planning and 

management activities (O’Leary and others 1992). There is an immediate practical 

need for accurate information on the distribution and composition of coastal sage 

scrub (O’Leary 1990, 1995). 

The Natural Communities Conservation Program (NCCP) is a regional multiagency effort to conserve biological diversity across southern California through 

directed scientific research, regional land use planning, and mitigation banking

(O’Leary and others 1992). The coastal sage scrub preservation plan was the NCCP’s

first pilot project. The project area corresponds roughly to San Diego, Orange, 

western Riverside and western San Bernardino counties. The NCCP region represents

the geographic extent of research described in this report. 

The Vegetation Type Map (VTM) survey was a Depression-era public works 

project conducted between 1929 and 1934 by the California Forest and Range 

Experiment Station (USDA Forest Service), Berkeley, California (Wieslander and 

others 1933). Survey teams mapped vegetation (at 1:64,000 to 1:125,000 scale)

across about one third of California, including all of coastal southern California 

exclusive of the Los Angeles basin. They also collected quantitative data on

dominant plant species cover from about 18,000 plots, including several thousand 

shrubland plots. The shrub plots were 0.1 acre (0.04 hectare) in extent, 33.0 feet (10.1

meters) by 132.0 feet (40.2 meters). Tree plots were twice that size. They allocated

about 75 plots per 15-minute map, equal to a sampling density of about 1 plot per 

2000 acres (800 hectares). The plots were subjectively placed in locations intended to 

represent typical stands of the qualitatively determined vegetation associations 

indicated on the vegetation maps. The plots were allocated fairly evenly over nondesert areas that were undeveloped at the time. The data were collected at a rather

early point in the explosive urban development of southern California; thus they 

constitute a valuable historical snapshot of California plant communities when they

were still much more intact and expansive than they are today. Between 1930 and 

2000 the population of San Diego County went from 210,000 to 2,814,000, doubling

about every 14 years (U.S. Census Bureau 2003).

Axelrod (1978) and Westman (1981a, 1981b, 1983) described geographic limits

of generalized coastal sage scrub types. These pioneering studies represent a starting 

point for further research on regional distributions of coastal sage scrub species and 

species associations. Axelrod’s work was based on a qualitative analysis of VTM 

vegetation maps, weather station data and other data. Westman’s work was based on 

quantitative analyses of data from 99 vegetation plots subjectively placed in coastal

sage scrub stands from El Rosario in Baja California, Mexico, to Contra Costa 

County, California, including 28 plots in the NCCP region. Both researchers

described an interior type of coastal sage scrub called Riversidian and coastal

associations called Venturan and Diegan. Axelrod’s (1978) Diegan association 

reaches its northern limit near Torrey Pines State Park. Westman’s (1983) Diegan

coastal sage scrub association extends north along the coast all the way to the Los 

58 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

Angeles County line. Species associated with the two authors’ coastal sage scrub

types are roughly comparable. 

Although coastal sage scrub is generally recognized as a highly diverse plant 

community, it is often dominated by just one or a few species when viewed at a scale

of a few meters to tens of meters (Mooney 1988). Substantial floristic variation exists 

at scales larger than individual stands. In other words, coastal sage scrub seems to 

have moderate alpha diversity, but higher beta and gamma diversity (Whittaker 

1978). Westman (1981a) noted that this property of coastal sage scrub makes it 

necessary to sample a relatively large number of sites to characterize the full range of 

floristic variation in the vegetation type. He noted that his own sampling effort was 

distinctly inadequate to this task and encouraged further study. Nevertheless, 22 years 

later most coastal sage scrub conservation planning is still based on Westman’s very 

general regional characterizations of the resource (Holland and Keil 1986), although

more detailed sub-regional characterizations have been published (DeSimone and 

Burk 1992, Kirkpatrick and Hutchinson 1977, White and Padley 1997). 

Access to coastal sage scrub stands on private lands has only gotten harder since 

Westman did his field work, making the task of collecting a modern regional data set 

increasingly problematic. For example, White and Padley (1997) were only granted 

permission to conduct surveys on 4 percent of the coastal sage scrub in their 

Riverside County study area. The VTM survey certainly does not represent a 

complete inventory of prehistoric California vegetation. But VTM data are still very

useful because so much land has been developed since it was collected and because

many of the cumulative impacts of various anthropogenic disturbance regimes have 

become more pronounced over time. 

This report presents preliminary findings of a regional study of coastal sage

scrub in southern California based on VTM plot data. I used a newly created spatial 

database of VTM plot data to examine general organizing principles of shrubland 

plant communities in southern California at a regional scale, with a focus on coastal

sage scrub. Geographic patterns of species distribution, floristic composition of major 

species associations, and environmental correlates with species distributions in 

coastal sage scrub were examined in a regional context with a large, systematically

allocated data set. 

Preliminary research questions can be stated as follows: 

• To what extent is coastal sage scrub really a floristically distinct plant 

community separate from contiguous chaparral plant communities? 

• How strong are the relative contributions of climatic factors and disturbance 

history in determining the chaparral-coastal sage scrub boundary? 

• What types (associations) of coastal sage scrub are evident among the VTM 

plots? How do these coastal sage scrub associations compare to

classifications and habitat preferences reported by other authors from other

data sources?

• How are these coastal sage scrub associations distributed geographically? 

Do coastal sage scrub associations previously described in detailed, subregional studies occur beyond the boundaries of those studies? 

• How have regional development patterns impacted the coastal sage scrub

associations?

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 59 
A New Look at Coastal Sage Scrub—Taylor 

Methods

The original researchers recorded VTM plot data on paper forms and maps that are 

currently archived at U. C. Berkeley. All remaining plot records collected south of 

the greater Los Angeles area were photocopied and entered into a Paradox database 

(Borland 1996). Extensive error checking and data proofing were undertaken to 

create a reliable database from the handwritten notes on the historic paper forms.

Although the data are generally of high quality, many classes of errors were discovered, including redundant, ambiguous or undefined species codes, occasional 

reported occurrences of taxa in completely implausible locations (sometimes resolvable as typographic errors or misread handwriting), codes miscopied from one part of

data sheets to another, mistakes in counting occurrences on data sheets, and mislabeled data sheets. Many were correctable after careful examination, but about 50 

occurrence records (=0.04 percent of the total) had to be deleted because they could 

not be reliably resolved. Most of those were ground-layer species with low 

frequency. Archaic taxonomy was updated according to the new Jepson manual

(Hickman 1993). Using a geographic information system (ESRI 1999), plot locations

were digitized from archived VTM plot maps based on 15-minute USGS topographic

maps originally published in 1898. About 1,600 plot locations were indicated on the

maps, but a few were never sampled and some plot data sheets have been lost. The 

result was a database of 1,481 plot records and about 275 species with about 12,000 

records of individual species occurrences. 

VTM surveyors divided each plot into a 5 by 20 array of 100 square cells, 6.6 

feet (2.0 meters) on a side. For each cell, they recorded which shrub species had the

greatest absolute cover. If total plant cover was less than 50 percent they recorded the 

cell as bare. The number of cells dominated by each species was its percent absolute 

cover in the plot. They also often noted a number of additional subdominant and 

ground-layer species that did not appear in the cover summary, characterizing their 

absolute cover on an ordinal scale of 1 to 3 (“occasional, frequent, or abundant”) 

(Wieslander and others 1933). I converted the ground-layer scores to plausible quantitative absolute cover values based on Braun-Blanquet cover classes (1=3 percent,

2=15 percent, 3=38 percent) and combined them with cover data for dominant 

species. For some analyses, all quantitative data were converted to presence/absence 

data. All results presented in this paper are based on untransformed absolute cover 

values. 

Plots were sorted based on their species compositions by means of several ordination and classification methods, using methods for gradient analysis developed by

Whittaker (1978). To increase the robustness of conclusions, several different ordination and classification programs were used. Programs included TWINSPAN (a

divisive, polythetic, hierarchical classification program) (Gauch and Whittaker 1981, 

Hill 1979), Detrended Correspondence Analysis (DCA) (Hill and Gauch 1980), Nonmetric Multidimensional Scaling (NMS) (Clarke 1993), and an agglomerative cluster

algorithm (McCune and Mefford 1998). Forest and woodland plots were deleted by

inspection. Chaparral plots that were extreme statistical outliers were also deleted, 

reducing the data set to 1,252 shrub plots with 203 species. As Westman (1983) 

found, deleting records of all species with fewer than five occurrences resulted in a 

great reduction of data (reduced the species list to 97) without affecting ordination

and classification results noticeably. This preliminary analysis identified 529 plots

60 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

representing coastal sage scrub vegetation. Additional ordination and classification 

studies were conducted with this subset of plots. 

Excellent detailed reviews of ordination methods include Whittaker (1978), 

Greig-Smith (1983), Clarke (1993) and McCune and Grace (2002). A brief summary

of ordination and classification methods follows. 

Non-metric Multidimensional Scaling (NMS, MDS, NMDS, or NMMDS) is an

ordination method that is well suited to data that are non-normal or on arbitrary, 

discontinuous, or otherwise problematic scales. It is based on ranked distances, so it 

is robust with regard to the “zero truncation problem” that plagues most ordination 

methods for ecological data. It has been widely applied in disciplines including sociology, psychology, and, more recently, ecology. NMS is an iterative search for a 

ranking and placement of plots in species space in relatively few dimensions that

minimizes the stress of the full multidimensional configuration. “Stress” is a measure 

of departure from monotonicity in the relationship between the dissimilarity (distance) in the original multidimensional space and distance in the reduced, relatively

few-dimensional ordination space. 

TWINSPAN stands for Two Way Indicator Species Analysis because it simultaneously classifies species and samples. The classification program was written at 

Cornell University for ecological data. TWINSPAN is a divisive algorithm that iteratively splits a set of plots into smaller and smaller groups by most efficient dividing 

of a reciprocal averaging ordination space. It iteratively sorts plots based on the indicator species value of the species and then sorts species based on indicator species

value of plots. Recent improvements to code remove instability problems of earlier 

versions (McCune and Mefford 1998). It is good at representing a single 

environmental gradient but performs poorly when multiple, more or less independent 

gradients exist. It produces a two-way table with plots and species sorted by similarities in overall composition. It is often useful as a “first cut” analysis of a large data 

set, especially for identifying meaningful subgroups of similar plots for further 

analysis. 

Detrended Correspondence Analysis (DCA; Hill and Gauch 1980) is an

eigenanalysis ordination technique for ecological data also based on reciprocal averaging (RA). Like RA and TWINSPAN, DCA ordinates both species and samples simultaneously. It removes two known artifacts from RA ordination by dividing the 

first axis into segments and rescaling the results. RA is equivalent to an eigenanalysis

of distances in a space defined by the chi-squared distance metric. Recent improvements to code remove instability problems of earlier versions (McCune and Mefford

1998). 

DCA has been quite popular in ecology. Recently it has been found to be less 

robust than NMS in a number of circumstances. In the course of the current research, 

DCA tended to seek and find unimodal species distributions in data sets while NMS

tended to represent the most abundant species as unimodal and produce bimodal distributions of less abundant species. DCA is generally more sensitive to outliers than 

NMS. 

Cluster analysis classifies items into groups based on their similarities. Cluster

algorithms in PC-ORD are hierarchical, agglomerative, and polythetic. Hierarchical

means that large clusters are composed of smaller clusters. Agglomerative means that

the analysis proceeds by joining clusters rather than dividing them (as TWINSPAN

does). Polythetic means that many attributes of the items are used to decide the 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 61 
A New Look at Coastal Sage Scrub—Taylor 

optimum way to combine or divide clusters. Analyses presented in this report were

based on a relative Sorensen’s-distance matrix and a furthest-neighbor-linkage rule 

(McCune and Grace 2002, McCune and Mefford 1998). 

Agglomerative cluster analysis was used in this study to further reduce the data 

set and make general observations about the composition and distribution of many 

sites with similar species composition. A well-defined cluster of plots serves as a

meaningful and objective way to describe different types (associations) of coastal

sage scrub. This level of aggregation is useful to compare groups of plots and their 

geographic distributions to published descriptions of coastal sage scrub species assemblages. 

The most ecologically meaningful levels of hierarchical clustering can be determined by a series of Indicator Species Analyses and Monte Carlo tests for statistical

significance as suggested by Dufrene and Legendre (1997). Their method combines 

information on the concentrations of species abundance in particular groups and

faithfulness of species occurrences in groups. It produces indicator values for each

species in each group. These values are tested for statistical significance using a 

Monte Carlo technique. If groups are too finely divided then indicator values will be 

low. On the other hand, if groups are too large, then internal heterogeneity reduces 

indicator values. Indicator values peak at some intermediate level of clustering that

varies by species. An appropriate level of clustering that maximizes meaningful natural breaks between groups of plots in a data set is the goal. 

In this study, a hierarchical cluster analysis was performed. Indicator Species

Analysis (McCune and Mefford 1998) was performed for each cluster across many 

levels, and Monte Carlo tests of randomized plot data were performed for each indicator species and each cluster at each level to distinguish groups with statistically

significant indicator species. P-values for all clusters across each level of the cluster 

analysis were tabulated, and cluster levels with the largest numbers of highly predictive indicator species were identified. Finally, graphical and geographic analyses of 

cluster groups were undertaken. Species and groups of plots with similar species 

composition were mapped to examine their geographic distribution. Similarly, cluster 

groups were mapped in ordination space. 

Other sources of spatial data were collected, including climate data (Franklin

1998), modern land use, and land cover data for San Diego (SANDAG 2000),

Orange, and western Riverside counties (data from USGS, Carlsbad CA, 2000).

Point-in-polygon and grid-identity operations in Arc Info (ESRI 1999) were used to 

extract additional attribute data for each plot location from these data sources.

Summary statistics for land use conversions of VTM plots were tabulated. Several 

indices of continentality (Tuhkanen 1980) were calculated from climate data. As a 

simple estimate of relative insolation, an index of north-ness was calculated from 

slope and aspect attributes (Borchert and others 1989). Correlations between 

environmental parameters, and patterns of species composition represented by cluster

groups were assessed using Pearson and Kendall’s correlation coefficients. 

Results 

The Berkeley archive contained plot records for 1,481 shrub plots in the NCCP 

region. A total of 58 plot records have been lost, most of them in Riverside County.

These maps and plot records have been missing at least since 1974 (Rich Minnich, 

62 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

personal communication). Gaps in the plot distribution showed the extent of agricultural and urban development in the early 1930s (fig. 1). These areas are delimited

even more clearly for some areas on the published VTM vegetation maps. General 

patterns of shrubland biogeography documented by other authors could be seen. 

Coastal sage scrub and chaparral were quite intermixed when viewed at a regional 

scale, with coastal sage scrub generally occupying more coastal and lower elevation 

areas. In a portion of central San Diego County from La Jolla to Oceanside, chaparral

came down to the sea. Otherwise it generally occurred at higher elevation. Many 

more VTM plot locations have been developed in San Diego County than in Orange 

or Riverside counties. Most of the developed plot locations contained coastal sage 

scrub, not chaparral. Of 529 coastal sage scrub plots in 1931, only 332 (63 percent) 

remained substantially undeveloped in 1995. The highest remaining numbers of undeveloped coastal sage scrub plots occurred on the Camp Pendleton Marine Base in

northern San Diego County, at Crystal Cove State Park in Orange County, and on 

(mostly) private land in the Perris Plain of western Riverside County. 

Figure 1—VTM plots in the NCCP area. c=coastal sage scrub about 1931, 

z=chaparral about 1931, ±=developed by 1995, ?=plot record missing. 

An ordination and a divisive classification of plots showed convergent results 

(fig. 2). Plots displayed along the first two synthetic axes of an NMS ordination based 

on similarities in species composition clearly divided into two major groups. There 

was a high correlation between membership in these groups and membership in 

groups of plots determined by a first level split in TWINSPAN (eigenvalue = 0.5530). 

Species composition showed the two groups were coastal sage scrub and chaparral 

vegetation types, as described in community ecology literature (Hanes 1988, Mooney 

1988). Indicator species for the first TWINSPAN split included Artemisia californica

(+), Adenostoma fasciculatum (-), and Quercus berberidifolia (-). Positive preferentials (associated with coastal sage scrub) included Eriogonum fasciculatum, Lotus 

scoparius, Malosma laurina, Poaceae spp., and Salvia apiana. Negative preferentials 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 63 
A New Look at Coastal Sage Scrub—Taylor 

(associated with chaparral) included Arctostaphylos tomentosa, Ceanothus crassifolius, Ceanothus greggii var. perplexans, Ceanothus leucodermis, Ceanothus tomentosus, Cercocarpus betuloides, and Rhus ovata. Taxonomy follows Hickman

(1993). When ordered by floristic composition, coastal sage scrub and chaparral appeared as distinct, separate communities intergrading only slightly.

Figure 2—Coastal sage scrub and chaparral plots in NMS (non-metric multidimensional scaling) ordination space with joint plots of environmental factors. Plots ordered according to floristic composition measured as relative Sorensen’s distance.

Coastal sage scrub and chaparral plots distinguished by first level TWINSPAN split. 

Joint plots show direction and relative magnitude of correlation between environmental factors and synthetic ordination axes. ELEVATE=elevation (r2

=0.22), pmeanyr=mean annual precipitation (r2

=0.20), mnmean-j=mean minimum January temperature (r2

=0.17), Conrad_C=Conrad’s continentality index (r2=0.13), temprang= 

mean annual temperature range (r2

=0.12). 

Several environmental factors were significantly correlated with the environmental gradient that distinguishes chaparral from coastal sage scrub (fig. 2). Chaparral was positively correlated with (and coastal sage scrub was negatively correlated

with) mean annual precipitation, elevation, and two indices of continentality: mean 

annual temperature range and Conrad’s C. Conrad’s C is the most widely used index 

64 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

of continentality. It is based on annual temperature range and seeks to remove effects 

of latitude (Tuhkanen 1980). Conrad’s C is: 

 C = (1.7*A/sin[φ+10°])-14 [1]

Where A=mean annual temperature range, φ=latitude. Chaparral was negatively 

correlated with (and coastal sage scrub was positively associated with) mean minimum January temperature. A continental climate has colder winters and greater

evapotranspirative stress in summer; it apparently favors chaparral over coastal sage 

scrub. 

An indirect gradient analysis of the dominant species in chaparral and coastal

sage scrub was conducted to explore general patterns of community organization.

Running averages of absolute cover for the five most widespread and abundant

species in all shrub plots were plotted along the primary NMS ordination axis. Dominant species had generally unimodal distributions in this indirect gradient analysis, 

with the exception of Salvia mellifera, which was somewhat bimodal (fig. 3). 

Figure 3—Mean cover values of dominant shrub species and coastal sage scrubchaparral boundary along an NMS ordination axis correlated with several environmental gradients. Adenfasc=Adenostoma fasciculatum; Eriofasc=Eriogonum

fasciculatum; Artecali=Artemisia californica; Salvapia=Salvia apiana; Salvmell=Salvia 

mellifera. 

The 529 plots identified as coastal sage scrub in the initial classification (white 

plots in figs. 1 and 2) were examined more closely by ordination and agglomerative

clustering (fig. 4). The plots are displayed along the first two synthetic axes of another NMS ordination. Distance between plots is a function of their similarity in species composition. Clusters of plots having greater within-group similarities in floristic 

composition than between-group similarities are restricted to more or less discrete

areas of the space defined by the synthetic ordination axes. Total precipitation, which 

was a fairly strong predictor of distributions for chaparral and coastal sage scrub, had

no significant predictive power for floristic patterns within coastal sage scrub. The

relative distribution of Artemisia californica and Eriogonum fasciculatum was 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 65 
A New Look at Coastal Sage Scrub—Taylor 

Figure 4—NMS (non-metric multidimensional scaling) ordination of 535 coastal sage

scrub plots with joint plots of environmental factors. Plots ordered by floristic composition measured as relative Sorensen’s distance. Axis 1 captured 41.6 percent total 

variance. Axis 2 captured an additional 36.1 percent of total variance (cumulative

total=77.7 percent). Axes are 99.7 percent orthogonal. Joint plots show direction and

relative magnitude of correlations with environmental factors. 

Symbols show most frequently occurring coastal sage scrub associations, classified by agglomerative cluster algorithm using relative Sorensen’s distance and 

furthest neighbor linkage. ∆=-Eriogonum fasciculatum-dominant type, ▼=Artemisia

californica-dominant type, ○=Salvia mellifera-dominant type, Ä =Salvia apianadominant type). pcn_s/w=ratio of summer precipitation to winter precipitation

(Pearson correlation coefficient r2

=0.30); pmean-jj=mean precipitation June-Sept. 

(r2

=0.25); ELEVATE=elevation (r2=0.21); mnmean-j=mean minimum January temperature (r2

=0.22); Conrad_C=Conrad’s continentality index (r2=0.19); X_COORD 

=easting (r2

=0.18). 

66 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

correlated with summer precipitation, elevation, distance from the coast, mean

minimum annual temperature, and the continentality index Conrad’s C. The variation 

in mean annual summer precipitation over the range of E. fasciculatum was only 10-

12 mm. No measured environmental factors were significantly correlated with the 

second ordination axis, which was determined by the relative distributions of Salvia 

mellifera and S. apiana. 

Of all dominant species, Artemisia californica (the single most widespread and 

abundant coastal sage scrub species in this study) was geographically distributed

most like Westman’s (1981a) Venturan coastal sage scrub (fig. 5), with markedly

coastal, low elevation preferences. A few other strictly coastal species (Rhus

integrifolia, Encelia californica) shared this general distribution. 

Figure 5—Regional distribution of Artemisia californica in VTM plots. Symbol size

proportional to absolute cover. 

The second most abundant species, Eriogonum fasciculatum, had a broad

distribution that ranged from near the coast to far inland (fig. 6). It occurred from 

near sea level to considerably higher elevations than Artemisia californica could 

apparently tolerate. 

The third most abundant species, Salvia mellifera, was broadly distributed but 

also somewhat patchy (fig. 7). It generally co-occurred with Artemisia californica, 

but each species also occurred abundantly in a few gaps in the distribution of the 

other. S. mellifera was better represented than A. californica in Riverside County. Its

distribution was not significantly correlated with any of the environmental factors

measured in this study. 

Classification and ordination routines tended to form groups defined by 

dominance of the most common species. Their geographic distributions were similar

to those of the dominant species that define them (fig. 8). The four most common

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 67 
A New Look at Coastal Sage Scrub—Taylor 

Figure 6—Regional distribution of Eriogonum fasciculatum in VTM plots. Symbol size 

proportional to absolute cover. 

Figure 7—Regional distribution of Salvia mellifera in VTM plots. Symbol size

proportional to absolute cover. 

68 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

Figure 8—Regional distribution of common coastal sage scrub associations in VTM 

plots. y=Artemisia californica-dominant; ∆=Eriogonum fasciculatum-dominant; 

□=Salvia mellifera-dominant; Ä=Salvia apiana-dominant 

cluster groups, dominated by Artemisia californica, Eriogonum fasciculatum, Salvia

mellifera and Salvia apiana, respectively, were quite a bit more intermixed in

geographic space than they appeared in floristically-defined ordination space. 

A few other clusters represented less common coastal sage scrub associations

with more restricted geographic distributions (fig. 9). A group of plots dominated by

a regionally distinctive association of Rhus ovata, Salvia apiana, and Keckiella

antirrhinoides var. antirrhinoides was restricted to a small portion of western

Riverside County. Another group of plots dominated by Eriogonum fasciculatum var.

foliolosum was restricted to Riverside and Orange Counties. That group’s distribution

reflected the distribution of its dominant subspecies, which was quite different than

the widespread regional representation by Eriogonum fasciculatum. 

Another group of plots dominated by Eriogonum fasciculatum, Salvia apiana,

and Artemisia californica with a distinct but minor element of Adenostoma 

fasciculatum was largely confined to San Diego County. VTM surveyors noted no

higher incidence of recent disturbance for these plots than for other similar plots. 

All of the common coastal sage scrub associations were broadly similar to

associations described by one or more other published studies (table 1). Some of the

less widespread groups appear to be unlike anything previously described. 

Approximately 63 percent of coastal sage scrub sites sampled by VTM 

surveyors about 1931 were still undeveloped in 1995. Since VTM plots were fairly

evenly allocated over undeveloped areas in 1931, this should be a fairly good 

estimate of relative amounts of coastal sage scrub lost in different parts of the NCCP 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 69 
A New Look at Coastal Sage Scrub—Taylor 

Figure 9—Regional distribution of other coastal sage scrub associations in VTM plots. 

▲=Artemisia californica, Eriogonum fasciculatum, Adenostoma fasciculatum- dominant. 

○ =Eriogonum fasciculatum var. foliolosum-dominated. Ä=Rhus ovata, Salvia apiana,

Keckiella antirrhinoides var. antirrhinoides-dominant. 

Table 1—Coastal sage scrub associations (plot clusters) compared to broadly similar, 

floristically-defined associations of coastal sage scrub described in other published studies. 

Coastal sage scrub

associations 

(cluster groups)

Sawyer & 

Keeler-Wolf 

(1995)

White & 

Padley 

(1994)

DeSimone 

& Burk 

(1992)

Kirkpatrick & 

Hutchinson

(1977)

Davis & 

others 

(1994)

Artemisia californica 1 (+3) 1 1 (+1) 1 (+1) 2 

Eriogonum fasciculatum 1 1 - 1 1

Salvia mellifera 1 1 1 2 3

Salvia apiana 1 1 2

Artemisia californica,

Eriogonum fasciculatum, 

Adenostoma fasciculatum

- X - - X

Eriogonum fasciculatum

var. foliolosum 

- - - - X

Rhus ovata, S. apiana,

Keckiella antirrhinoides 

- - - - -

This table shows how many floristically-defined associations of coastal sage scrub described by other

published studies (in rows) are broadly similar to associations defined in the current study (in columns,

named by dominant indicator species). X = series overtly excluded by site selection criteria or analysis

in a published study. The terms “association” (Kirkpatrick and Hutchinson 1977), “subassociation” or 

“group” (DeSimone and Burk 1992), and “series” (White and Padley 1994, Sawyer and Keeler-Wolf

1995) are roughly equivalent. Davis and others (1994) defined “species assemblages” based on

frequently occurring dominant species in larger (>1km2

) mapping units than the other studies. 

70 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

region since that time. Fractions of VTM coastal sage scrub plots developed during

that time period varied by county (table 2). 

The fraction of VTM plots that were lost to land development between 1931 and 

1995 was different for different associations of coastal sage scrub (table 3). In several

cases losses were disproportionate to a type’s total representation in the plant

community. This reflected regional patterns in initial distribution of the plots and

regional patterns of land use conversion.

Table 2—Fractions of VTM coastal sage scrub plots developed 1931-1995, by county. 

County Orange Riverside San Diego Total

Number of plots sampled in 1931 91 123 315 529 

Number of plots undeveloped in 1995 71 91 170 332 

Percent of plots remaining 78 74 54 63

Discussion 

It is important to note that large amounts of coastal sage scrub were lost to agriculture 

(and to a lesser extent, urbanization) before the VTM plot data set was collected.

VTM vegetation maps show large areas of urban and agricultural land use. Various 

researchers believe those developed areas were formerly occupied by native grasslands, coastal sage scrub and other shrublands, or lost native forb communities, in unknown proportions. Most undeveloped areas sampled by the VTM crews had been

subjected to repeated burning, grazing, and other types of disturbance, with

imprecisely known effects on vegetation. The VTM survey cannot answer many 

questions about what happened in the preceding decades and centuries. The survey

does provide a large sample of consistent, quantitative observations of this vegetation 

about 1931. These data can be used to characterize vegetation patterns at that time

and to assess how vegetation has changed since that time. 

Table 3—Fractions of major coastal sage scrub species associations developed, 1931-1995.

CSS associations 

(cluster groups)

Number of plots 

about 1931

Number of plots still 

undeveloped in 1995

Pct plots 

 remaining, 1995 

Rhus ovata, Salvia apiana, 

Keckiella antirrhinoides

var. antirrhinoides 

4 4 100

Eriogonum fasciculatum

var. foliolosum-dominated

14 11 77

Salvia apiana-dominated 20 14 70

Artemisia californicadominated 303 193 64

Artemisia californica,

Eriogonum fasciculatum, 

Adenostoma fasciculatum

22 13 59

Salvia mellifera-dominated 53 29 55

Eriogonum fasciculatumdominated 105 53 50

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 71 
A New Look at Coastal Sage Scrub—Taylor 

The rate at which VTM coastal sage scrub plots were lost between 1931 and 

1995 is plausibly consistent with the previously published overall estimates of coastal 

sage scrub loss (63 percent of the VTM sites remained compared to Michael Brandman Associate’s estimate that 44 percent of original coastal sage scrub remained). 

Since 1931, much more shrubland in general (fig. 1) and coastal sage scrub in particular (table 2) has been lost in San Diego County than has been lost in Orange and 

Riverside Counties. This is partly because Orange and western Riverside Counties 

converted much more land to agriculture than San Diego County did prior to 1931.

Urban development throughout the region generally occurred on flat agricultural 

lands first and adjacent hillsides second. In San Diego, 20th century development 

moved onto hillsides (where coastal sage scrub still existed) to a much greater extent

than in the other two counties. 

Camp Pendleton and Crystal Cove State Park contain almost all the remaining 

coastal shrublands in the NCCP region. Without the large areas preserved at Camp

Pendleton, the loss would be even greater in San Diego County. 

Ordinations and classifications are generally driven by common species. This is

true in ordinations of coastal sage scrub and chaparral, since stands of these plant 

communities are generally dominated by just a few species. Coastal sage scrub and 

chaparral are floristically distinct and easily distinguishable by a variety of ordination 

and classification methods (fig. 2). 

Several correlated environmental factors related to seasonal water availability

make up an environmental gradient that separates coastal sage scrub and chaparral

communities. Although chaparral tends to receive more annual precipitation, it also

endures more extreme evapotranspirative stress in the summer. The fact that winter 

minimum temperatures are more strongly predictive of chaparral and coastal sage 

scrub distributions than summer maximum temperatures may partly be a reflection of 

wintertime temperature limits on efficient photosynthesis and growth for coastal sage 

scrub, which grows more actively during the winter months than does chaparral. 

One of the longest standing discussions in plant community ecology concerns 

the relative merits of a community unit model of vegetation (Clements 1916) versus

an individualistic (continuum) model (Gleason 1926). Over the years, a number of 

researchers have tested predictions of these two models against data from various 

sources (Hoagland and Collins 1997, Keddy 1987, McIntosh 1967, Whittaker 1975). 

Whittaker (1975) referred to the abundance curves of many species measured along 

an environmental gradient as a coenocline, which he used as a way to visualize plant 

community structure. A coenocline of dominant shrub species in southern California

shrublands along the most influential environmental gradient in those communities 

produced a typically ambivalent result (fig. 3). Species responses appeared more or 

less individualistic, but there was also some sense of a synchronous species turnover 

at the chaparral-coastal sage scrub boundary. Coastal sage scrub species occurred at 

the low precipitation, coastal end of the environmental gradient, and chaparral (represented by Adenostoma fasciculatum) occurred at the high precipitation, continental 

climate end of the axis, as has been reported in literature (Hanes 1988, Mooney 

1988). Most species exhibited a more-or-less unimodal distribution as predicted by

Whittaker (1978), even when ordinated by NMS, which does not assume unimodal

distributions for species. 

The general observation that Eriogonum fasciculatum prefers a more continental

climate than Artemisia californica is clear. The somewhat bimodal distribution of 

72 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

Salvia mellifera is probably artifactual, since that species was not found to be significantly correlated with these environmental factors (see fig. 4). The major coastal sage 

scrub species persist at low cover levels well into the portion of the coenocline dominated by chaparral species (represented in fig. 3 only by Adenostoma fasciculatum for 

clarity). A. californica is more restricted in distribution than the other species. The 

resolution of these species distribution curves was much higher than in other published accounts (see Westman 1981b, page 447), but even with hundreds of additional observations, this analysis did not produce a great amount of additional insight. 

Analysis of coastal sage scrub plots suggested that summer precipitation may 

possibly be an important environmental factor in determining the regional distribution of Eriogonum fasciculatum and Artemisia californica, but not of Salvia mellifera

and S. apiana. It seemed unlikely that the relationship would be causal, since summer 

rainfall anywhere in the coastal sage scrub zone is generally infrequent, unreliable,

and scant (John O’Leary, personal communication). That was the strongest regional

signal in this ecological data set, however (fig. 4). Summer rainfall patterns were invoked by Axelrod (1978) to explain the distribution and composition differences 

between Diegan (south of Torrey Pines State park in San Diego County) and Venturan coastal sage scrub (north of same). However the regional patterns of summer 

precipitation that he described (more summer precipitation south of Torrey pines in

coastal San Diego County) were not evident in this data set. 

The geographic distribution of Artemisia californica and the A. californicadominated coastal sage scrub type (fig. 8) corresponded more or less to the distribution of Westman's (1983) Diegan association (fig. 5). Other dominant species including Eriogonum fasciculatum (fig. 6), Salvia mellifera (fig. 7), S. apiana and 

coastal sage scrub associations (floristically-determined groups of plots) dominated

by these species (fig. 8) generally did not conform very well to the boundaries of regional coastal sage scrub types defined by Westman (1983) or Axelrod (1978). There

were general similarities between the distributions of these coastal sage scrub associations and maps of species assemblages derived from the California GAP database

(Davis and others 1994). These similarities were, to some extent, circularly referential because both studies were based at least in part on VTM data (plot data or vegetation maps derived from those data). 

In general, the most abundant floristically-defined groups of plots showed less 

distinct patterns when viewed in geographic space (fig. 8) than they did when viewed 

in species-attribute space (fig. 4). It appears that the geographic distribution of coastal 

sage scrub species associations evident in the VTM plot data are primarily 

determined by environmental variation occurring at a scale considerably smaller than

that of the NCCP region. Field observations by DeSimone and Burk (1992) support 

my own observations that stands of coastal sage scrub vary considerably in species

composition at spatial scales on the order of tens to hundreds of meters. When plot 

data are aggregated over the NCCP region, slope and aspect are poor predictors of 

species distributions, even when combined as an index of “north-ness.” These factors 

might have more predictive power if subsets of plots from smaller geographic areas 

were analyzed, because slope and aspect often appear to control local distributions of 

shrub species. Both DeSimone and Burk (1992) and the current study recognized a

Salvia mellifera-dominated coastal sage scrub type, and neither study found any

measured environmental factor that was correlated with its distribution. Salvia

mellifera’s distribution may possibly be controlled by biological or historical factors 

of site disturbance or site colonization not documented in these studies. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 73 
A New Look at Coastal Sage Scrub—Taylor 

Sample size and sample allocation methods clearly explain some differences in 

published accounts of regional coastal sage scrub variation. Since the VTM survey

sampled about 18 times as many coastal sage scrub sites in the NCCP area as Westman (1983) did, it is no surprise that VTM plot data revealed a more complex picture 

of coastal sage scrub associations. The main plot groups (coastal sage scrub associations) that emerged from this study’s classifications were generally determined by the 

dozen or so most frequent and dominant species. The main associations (cluster 

groups) appear to be dominated in turn by each of the most dominant species. Thus it

appears that the methods of the California GAP analysis (describing vegetation cover

by noting triplets of dominant species) may be a reasonably good way to make

general characterizations of variation in coastal sage scrub associations (Davis and 

others 1994). 

Some of the compositionally determined groups of plots found in this study resemble species associations described by Kirkpatrick and Hutchinson (1977), 

DeSimone and Burk (1992), and other more intensive studies of smaller areas

(table 1). The amount of variation found when studying coastal sage scrub is partially

a function of sampling intensity, either in a regional study like the VTM survey, in a 

mid-scale study (Kirkpatrick and Hutchinson 1977, White and Padley 1997), or in an 

intensive local survey (DeSimone and Burk 1992). Axelrod’s (1977) and Westman's 

(1983) generalized regional species associations begin to break down as sample size 

of a study increases, especially when sampling strategy aims to capture the full range

of compositional variation (as in the VTM Survey) rather than sampling only sites 

typical of the most abundant types. 

Previous studies (Kirkpatrick and Hutchinson 1977, 1980; Westman 1983) 

found that some understory species showed greater habitat specificity than the dominant shrub species did. There is no evidence of this in the VTM plot data. But since

VTM crews took only cursory, and sometimes incomplete, notes on understory

species, these findings should not be taken as definitive on this count. 

In summary, this regional scale study has largely substantiated findings of previous local studies of coastal sage scrub. Coastal sage scrub associations described in 

local studies are here shown to occur more broadly than the areas within which they

were originally described. Generalized maps of the regional distributions of coastal

sage scrub associations were produced. An intermediate coastal sage scrub-chaparral 

type with Adenostoma fasciculatum was found that did not appear to be dependent 

upon recent site disturbances. Several occurrences of a narrowly distributed coastal 

sage scrub type dominated by Rhus ovata, Salvia apiana, and Keckiella

antirrhinoides var. antirrhinoides were found on private land in western Riverside 

County. A new coastal sage scrub type dominated by Eriogonum fasciculatum var. 

foliolosum is described that was not described by previous authors, although it 

appears to occur in areas where they would have had opportunities to detect it. This 

may possibly reflect variable levels of attention to taxonomic details in previous 

surveys. It also highlights the importance of collecting samples for a large number of

sites when describing floristic variation in a diverse and patchy plant community. 

Coastal sage scrub characteristically is dominated by only a few species. The

strong control of community organization by the few most dominant species is one of

the most striking attributes of the community. Despite the widespread impression that

coastal sage scrub is a highly diverse community, most of the species occur 

infrequently and in rather low abundance. 

74 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
A New Look at Coastal Sage Scrub—Taylor 

The biggest regional trend observable in the dominant species is a transition 

from Eriogonum fasciculatum dominance at interior sites to Artemisia californica

near the coast. Of the environmental factors measured in this study, summer precipitation, ratio of summer to winter precipitation, and minimum winter temperature are 

most important in predicting the relative differences in abundance of E. fasciculatum

and A. californica. Salvia apiana prefers interior, higher elevation sites with high 

annual precipitation and an extremely continental climate. Salvia mellifera’s distribution is poorly predicted by any environmental factors measured in this study.

Different associations of coastal sage scrub were lost in varying amounts. 

Coastal sage scrub associations occurring frequently in coastal San Diego County and 

at low elevations suffered the highest rates of conversion, while those found in

Orange and Riverside Counties, or at high elevations throughout the region, remained

more intact. Associations of coastal sage scrub that have experienced high rates of 

loss (Eriogonum fasciculatum and Salvia mellifera-dominated associations, table 3) 

ought to be considered high priorities for inclusion in the NCCP preserve system.

Regionally distinct and geographically restricted associations (shown in fig. 9) may

warrant additional study as candidates for inclusion in a regional conservation plan. It 

would be useful to compare the proportions of coastal sage scrub associations currently included in preserves to these statistics to see how representative the current 

preserve system is of the diversity that existed in 1931 and the remaining diversity

existing today. 

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USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 77 

Air Pollution and Vegetation Change in 

Southern California Coastal Sage Scrub: 

A Comparison with Chaparral and 

Coniferous Forest1

Edith B. Allen2, Abby G. Sirulnik2, Louise Egerton-Warburton2, 5, 

Sheila N. Kee2

, Andrzej Bytnerowicz3, Pamela E. Padgett3, Patrick 

J. Temple3

, Mark E. Fenn3, Mark A. Poth3, and Thomas Meixner4

Abstract 

The coastal sage scrub (CSS) vegetation of southern California is rapidly converting to annual

grasslands, perhaps in part because of air pollution. By contrast, chaparral and coniferous 

forest are subject to equally high levels of air pollution but are relatively stable. A 

comparative analysis of ozone and nitrogen deposition on plants of CSS, exotic annual

grassland, chaparral, and coniferous forest shows these vegetation types have different

susceptibilities to each pollutant. Historically high concentrations of ozone in the local 

mountains weakened pines, contributing to tree mortality. Native shrub seedlings had 

decreased growth in chambers with current-day levels of 150 ppb ozone. Under natural field 

conditions the shrubs may escape ozone injury by being physiologically active early in the 

season when ozone concentrations are below phytotoxic levels. Summer-active pines are 

more susceptible to ozone than summer-deciduous CSS shrubs and senescent annual grasses. 

Nitrogen deposition has different impacts from ozone because N accumulates on leaf and soil

surfaces during the summer. Conifers are more susceptible to leaf-deposited nitric acid 

because they are physiologically active in summer, while chaparral may be less so because of 

thick cuticles and reduced summertime activity. Deciduous CSS and senescent grasses are

less susceptible to direct leaf damage. However, N becomes available for root uptake after the 

first fall rains. Soil accumulations up to 87 μg/g extractable N have been measured in surface 

soil of CSS shrubland, levels that have caused mortality in the greenhouse. Grasses may 

escape the deleterious effects of high soil N levels because of their annual habit. Coniferous

forest may have a higher threshold for N damage because of high stand biomass, high N 

immobilization in soil organic matter, and watershed N runoff. The resistance of chaparral to

high N is less well understood, but may be due to higher biomass and slower growth rates 

than CSS and also high leachate losses of N. 

Key words: chaparral, coastal sage scrub, coniferous forest, exotic annual grassland, nitrogen

deposition, ozone 

________________________ 

1 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2 Department of Botany and Plant Sciences and Center for Conservation Biology, University of

California Riverside 92521 (email: edith.allen@ucr.edu) 3 USDA Forest Service, Pacific Southwest Research Station, Forest Fire Laboratory, Riverside, 

California 92507

4 Department of Environmental Sciences, University of California Riverside 92521 5 Present address: Conservation Science Department, Chicago Botanic Gardens, 1000 Lake Cook Rd., 

Glencoe, IL 60022 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005 79 
Air Pollution and CSS Change—Allen and others 

Introduction 

We are currently witnessing a rapid vegetation type-conversion of the coastal sage 

scrub (CSS) of southern California to exotic annual grassland, a change that has been

occurring over the past three decades (Allen and others 1998, Minnich and Dezzani

1998). Local botanists report that mountainsides, such as the Box Springs Mountains 

and Mount Rubidoux, that were covered with CSS only 20-25 years ago are now

dominated by less diverse stands of Mediterranean annual grasses and forbs. In other 

areas where shrubs still occur they have exotic-dominated understories, and the 

native herbaceous species that used to occur have largely been replaced. This

directional change in vegetation composition has been attributed to such factors as

urbanization, fragmentation and corridors that increased weed movement (Zink and 

others 1995), and historic domestic livestock grazing that removed native species and

facilitated weed seed dispersal (Burcham 1957). Frequent fire is creating a positive 

feedback of increased dominance of highly flammable annual grasses, compared to

CSS shrubs that burned on a 25-30 year fire cycle (Minnich and Dezzani 1998). Air 

pollution, which has less often been considered a factor in vegetation change in CSS, 

is the focus of this review. Understanding the impact of air pollution for conservation

of CSS is critical because it impinges upon vegetation even in those reserves that 

have been protected from other impacts. Air pollution is greatest in areas dominated 

by Riversidean sage scrub (Westman 1981), so this CSS association will be 

emphasized. 

As part of this discussion, we compare the type-conversion of CSS to annual

grassland with the comparatively stable vegetation in two other major vegetation 

types of southern California, chaparral and mixed coniferous forest. CSS harbors 

some 200 sensitive species and is the least widespread of these three vegetation types 

(Sawyer and Keeler-Wolfe 1995, Skinner and Pavlik 1994). It is the most impacted

not only because of air pollution but also because of its location in the most favorable 

low-elevation private lands planned for development. By contrast, much chaparral

and coniferous forest is protected in public lands. However, all three vegetation types 

are subject to air pollution, both ozone and nitrogen deposition that originate from

automobile emissions. Up to 45 kg N/ha/yr are deposited in San Bernardino mixed

coniferous forests (Fenn and others 1998) and 30 kg/ha/yr in chaparral in the San 

Dimas Experimental Forest (Bytnerowicz and others 1987, Riggan and others 1985).

All three vegetation types have been impacted by ozone as well, with historical 

monthly averages as high as 350 ppb (parts per billion) in 1975 (fig. 1). 

However, ozone levels have declined with increasing regulation and pollution 

control; summertime peak concentrations above 150 ppb occur infrequently. In this 

analysis we hypothesize that ozone and N deposition have affected vegetation of

southern California historically and currently, but CSS, exotic annual grassland,

chaparral, and coniferous forest have different sensitivities to the two forms of air 

pollutants. 

To help elucidate the impacts of N on CSS, we present the preliminary results of 

an N fertilization experiment in CSS vegetation. Anthropogenically deposited nitrate, 

the dominant form in Riversidean CSS (Padgett and others 1999), is formed from 

nitrogen oxides, which co-occur with ozone. The concentrations of ozone and nitrogen oxides increase simultaneously along rural to urban gradients of air pollution. To

separate the confounding effects of ozone and N, a fertilization experiment was done 

in a rural area with relatively clean air, at Lake Skinner in the Western Riverside

County Multispecies Reserve. The objective of this experiment was to evaluate the 

80 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution and CSS Change—Allen and others 

long term effects of N on CSS plant species composition in a controlled field setting,

and we present results from the first 8 years of fertilization. 

0

0.05

0.1

0.15

0.2

0.25

0.3

0.35

0.4

JAN FEB MAR APR MAY JUN JUL AUG SEP OCT NOV DEC

Month

1975

1995

Figure 1—Mean monthly high concentrations of ozone in 1975 and 1995 in Riverside

(daily hourly peak values averaged over the month). Different analytical methods

were used in 1975 that detected total oxidants, but are corrected to give 

approximations of ozone (AQMD). 

Comparative Air Pollution Effects in CSS, Chaparral, and 

Coniferous Forest 

Ozone 

Pinus ponderosa and Pinus jeffreyi develop visible ozone damage at daily average 

levels of 50-60 ppb (Miller and others 1997). The ozone sensitivity of pines was 

known early from the studies of Miller and others (1963) when levels of ozone were 

extremely high in the Los Angeles Basin and surrounding mountains. The high

concentrations of past decades (fig. 1) contributed to conifer mortality in the 

mountains, in part coupled with drought, bark beetle and other stressors (Miller and

others 1997). The dead trees from the 1960s and 1970s were often harvested, 

eliminating the evidence of ozone-induced mortality (Arbaugh and others 1999,

Miller 1992). At current reduced levels of ozone, tree mortality does not seem to be a 

short-term response (Arbaugh and others 1999), but ozone is nevertheless affecting 

the physiology of trees in the field (Grulke and others 1998, Takemoto and others 

2000, Temple and Miller 1994). For instance, at Camp Paivika in the San Bernardino 

Mountains, ponderosa pine retains 1-2 years of needle growth, rather than the typical 

5 years (Grulke and Balduman 1999, Miller and others 1997). The ground at this site 

is littered with a deep cover of undecomposed needle litter. 

The few available studies of ozone effects on native CSS and chaparral shrubs 

indicate that visible injury occurs at levels of ozone 2 to 10 times higher than for

pines. For instance, seedlings of Adenostoma fasciculatum, Ceanothus leucodermis, 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 81 
Air Pollution and CSS Change—Allen and others 

Arctostaphylos glauca, and Quercus dumosa experienced visible damage at the

relatively high levels of 100 to 500 ppb ozone (Stolte 1982). Well-watered seedlings 

of Salvia mellifera and Eriogonum cinereum, common CSS shrubs, experienced leaf 

drop and reduced flowering after fumigation in the greenhouse with 100-200 ppb 

ozone (Westman 1990). One interesting observation that Westman (1990) reported

was that the exotic grass Bromus madritensis ssp. rubens had an ecotype less

sensitive to ozone in areas of the Los Angeles basin with high air pollution in the

early 1980s. 

We have observed mortality of CSS shrubs in the field, especially in urban areas

(Allen and others 1998, 2000), but the cause of death is not clear and has not been

quantified across the landscape. CSS shrubs are summer-deciduous and are 

physiologically active mainly in winter and spring (fig. 2), when mean daily high 

ozone levels are below 60 ppb (fig. 1). Thus they may avoid extensive ozone injury

by dropping leaves before high ozone exposures occur. Westman (1990) observed

that CSS shrubs experienced considerable ozone damage in the Santa Monica 

Mountains, but his work was done in the mid 1980s when springtime, growing

season ozone levels were much higher than today. Maximum hourly summer high 

ozone levels were 180-390 ppb ozone in the Santa Monica Mountains in 1986

(Westman 1990). Nevertheless, we should not discount the long-term, low level

effects of current springtime mean daily high exposures of 50-60 ppb, although shortterm chamber studies have shown no acute effects at these levels (Preston 1986). 

-70

-60

-50

-40

-30

-20

-10

0

JAN FEB MAR APR MAY JUN JUL AUG SEP OCT NOV DEC

Month

Adenostoma

Artemisia

Pinus

Figure 2—Xylem water potentials of three dominant species of coniferous forest

(Pinus ponderosa), chaparral (Adenostoma fasciculatum) and CSS (Artemisia 

californica). Soil moisture is available during the summer in the coniferous forest

where deep-rooted trees tap fractured rocks. Artemisia and other species of CSS are

partially or wholly deciduous in summer, while Adenostoma and other chaparral

shrubs have small schlerophyllous leaves that close their stomates in response to dry 

soil. Data from Poole and Miller (1975), Coyne and Bingham (1982). 

82 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution and CSS Change—Allen and others 

Chaparral may also be less sensitive to present-day levels of ozone than

ponderosa pine and Jeffrey pine. The sclerophyllous leaves of the chaparral dominant 

Adenostoma fasciculatum close their stomates in response to summer drought (Hanes 

1965, Poole and Miller 1975) and so are also unlikely to absorb a great deal of 

ambient ozone in summer at present day levels. By contrast, pines have deep roots

that seek water in rock fractures (Hubbert and others 2001a) and may continue to

photosynthesize and absorb ozone all summer long (Coyne and Bingham 1982, 

Hubbert and others 2001b). Photosynthesis is lower in summer than spring resulting 

in lower rates of ozone uptake (Temple 1996), but considerable amounts of ozone 

may still be absorbed under high summertime ambient concentrations. The 

phenological differences among these three vegetation types may explain why pine is

so sensitive to the high summertime levels of ozone. In 1975 even the springtime 

ozone levels were as high as 150 ppb (fig. 1). Had regulations not been enforced to 

reduce ozone to present day levels, coniferous tree mortality would have continued. 

The ozone damage symptoms that Westman (1990) observed in CSS shrubs in the 

mid-1980s would be more widespread, with unknown consequences on mortality. 

Further reductions in ozone are needed to improve the health of pine forests in 

locations that currently receive high exposures such as the western San Bernardino 

Mountains. 

Nitrogen 

Most of the N deposition in southern California occurs as dryfall in the form of 

HNO3/NO3

-

 originating from automobile exhaust, and about 10-20 percent arrives as 

NH3/NH4

+

 from agricultural origins (Fenn and others 1998, Padgett and others 1999) 

or more depending on proximity to local dairy farms (Fenn and others 2000). The 

proportion of agricultural input is expected to decline as agriculture moves from the 

south coast and Inland Empire counties to the Central Valley and the deserts. While 

neither form of N input is desirable, NH3/NH4

+

 acidifies the soil more than 

HNO3/NO3

-

, as documented in the Netherlands where up to 90 kg ha-1 yr-1 of 

NH3/NH4

+ are deposited annually (Bobbink and Willems 1987). CSS soils 

accumulate up to 87 μg/g of extractable N during summer near Riverside. These soils

are not more acid than soils in reserves that have less air pollution, such as Lake

Skinner with a maximum of about 20 μg/g extractable soil N in the dry season 

(Padgett and others 1999). Nitrate has been increasing in soils downwind of urban 

areas over the past 40 years, with different effects on native and invasive exotic 

vegetation, as described below. 

While historic air concentration data for HNO3/NO3

- are not available as for

ozone, NOX emissions have been calculated since 1938 (fig. 3). These emissions

reached a high of nearly 3,000 T daily (dy-1) in 1972 in the Los Angeles Basin and 

declined to present day levels of 1,300 T dy-1, except for a rise in 1999 that may be

related to La Niña stagnant air conditions. NOX is converted to NO3

- in the 

atmosphere, more so under hot and sunny summer conditions. Most of the N 

deposition occurs as dryfall during the summer (Padgett and others 1999), although

areas with foggy seasons, as occur in some pine forests in the San Bernardino 

Mountains, may have high N deposition at that time of year (Fenn and others 2000).

Summer-deposited nitrate accumulates on soil and leaf surfaces and becomes

available for plant uptake during the following winter rainy season when it is leached 

into the soil. Because HNO3 is sufficiently acidic to cause leaf lesions at high 

concentrations, dryfall of HNO3 actually has two separate effects, one to fertilize the 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 83 
Air Pollution and CSS Change—Allen and others 

soil and a second to damage the leaf. Some direct uptake of leaf-deposited N occurs 

also. These different effects of N are considered separately.

Figure 3—Emissions of NOX in the South Coast Air Basin, 1937-2000 (Alexis and 

others 2000).

Epicuticular leaf lesions, visible using the electron microscope, may be induced

by deposition of atmospheric HNO3 in California black oak and ponderosa pine 

(Bytnerowicz and others 1998, 1999). Lesions may be accompanied by structural 

changes of epicuticular waxes around the stomates. HNO3 concentrations of 50 ppb

for 12 hours were sufficient to cause damage to pine, but oak damage was observed 

at the higher dosage of 200 ppb. Cuticular damage was observed at ambient levels of 

30 ppb HNO3 in pine and oak under longer exposures (Parry 2001), although longterm impacts on plant growth still need to be evaluated. The latter study showed the 

importance of transcuticular uptake of HNO3, indicating that even physiologically 

inactive plants with closed stomates may be subject to damage from HNO3

deposition. CSS shrubs drop most of their leaves in summer and may be largely

unaffected by leaf-deposited HNO3/NO3

-

. Chaparral shrubs may be protected from 

extensive nitric acid damage because most of these species have sclerophyllous 

leaves with thick cuticles, but this is a hypothesis to be tested. 

While HNO3 may cause damage to the leaf surface, it also acts as a fertilizer. As 

much as one-half of the leaf-deposited N is taken up directly by the stomates into the 

mesophyll (Bytnerowicz and others 1999). The N that remains on leaf and other plant 

surfaces may enter the soil via throughfall and stem flow and become available for

uptake by roots. In vegetation with a low leaf area, such as Riversidean CSS, some of 

the N will deposit directly on the soil, also increasing the N available to plants. We 

have done a series of N fertilization experiments to determine whether exotic grasses 

and forbs have a greater growth response to soil NO3

-

 than native CSS plants. In

greenhouse and field studies, 15NO3

- was taken up at greater rates by individual

Bromus madritensis than Artemisia californica (Yoshida 1999, Yoshida and Allen 

84 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution and CSS Change—Allen and others 

2001). Three exotic species (Bromus madritensis, Avena fatua, Hirschfeldia incana) 

had luxury consumption levels of N in their tissues and increased biomass as well 

(Padgett and Allen 1999). Three native shrub species (Artemisia californica,

Eriogonum fasciculatum, Encelia farinosa) had continued growth response with 

increased soil N up to 80 μg/g, but without luxury uptake, that is, with constant levels

of tissue N. These experiments were based on comparisons of individual shrubs with 

individual weeds, but in the field the weeds are prolific seeders and individual shrub 

seeds or seedlings are overwhelmed by competition from many individual weedy 

plants (Eliason and Allen 1997). Additional long-term experiments showed that 

Artemisia and Encelia suffer mortality if soil extractable N is maintained at 30-50 

μg/g for 6 or more months (Allen and others 1998; Padgett, unpublished data). These 

concentrations probably only occur in surface soils during the dry season in the field, 

so we do not know whether shrubs may be dying from high levels of N in N-polluted 

sites in the field. However, exposure to high levels of fertilizer may shorten the 

lifespan of native plants, as has been shown in botanic gardens (Keator 1994). The 

annual grasses escape long-term nutrient stress by having a short lifespan with high 

seed production. 

Ongoing N fertilizer studies in the San Bernardino Mountains show that

ponderosa pine is responding to N additions of 50 kg/ha with increased diameter 

growth even in sites that currently receive high N deposition of 45 kg/ha/yr (Fenn 

and Poth 2001). Levels of excess N are so high that NO3

- in streamwater is the 

highest in North America from wildland watersheds (Fenn and Poth 1999). Even so,

the available soil N is apparently still limiting during the spring and summer when 

plants are physiologically active. N fertilization of chaparral caused a small increase

in branch length of Adenostoma fasciculatum and Ceanothus greggii, indicating N 

deficiency (McMaster and others 1982). This research was performed at the Sky 

Oaks Field Station, east of Mount Palomar, in an area that is relatively free of air

pollution even today. Plants of chaparral have lower CO2 uptake and lower growth

rates than CSS shrubs (Hanes 1988, Mooney 1988) and would not be expected to

respond to nutrient additions with rapid growth. In contrast, fertilized CSS shrubs

increased their growth and took up N at equivalent or only slightly lower rates than 

annual weeds, as discussed above (Padgett and Allen 1999, Yoshida and Allen 2001). 

The loss of shrub cover in CSS is surprising considering the shrubs respond so well 

to N, so we initiated a field N fertilization experiment to study the interactions of 

exotic annuals with native shrubs. 

Nitrogen Fertilization Experiment in CSS 

The effects of N deposition on vegetation can be determined by examining the 

cumulative effects of long-term changes along an N deposition gradient and also by 

conducting N fertilizer experiments in an area that receives low N deposition. 

Changes in the N gradient along the Riverside-Perris Plain (described in Padgett and 

others 1999) are confounded by multidimensional changes as for any natural

gradient. The covariance of ozone with nitrogen oxides is the primary concern, but

local differences in soil type and fire history also occur, even though sites have been

chosen to be as similar as possible. We hypothesized that N deposition or fertilization

induces an increase in annual grasses and weeds and a decrease in shrub cover. The 

opportunity to begin these studies arose after the November 1993 fires that burned 

about 4000 ha of the Western Riverside Multispecies Reserve. We initiated N 

fertilizer plots on the north shore of Lake Skinner in the Reserve, where the edge of 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 85 
Air Pollution and CSS Change—Allen and others 

the fire had burned, and included burned and unburned sites. The unburned CSS 

vegetation showed very little measurable growth response to N fertilizer, so data 

from these plots will not be shown here, but the burned site had several notable 

responses to N fertilization. 

Fertilization began in winter/spring 1994 with the application of NH4NO3 twice 

each growing season at 30 kg N ha-1 for a total of 60 kg ha-1 yr-1. This rate was chosen 

to double the highest known deposition in shrublands at the San Dimas Experimental 

Forest in the San Gabriel Mountains of 30 kg ha-1 yr-1 (Riggan and others 1985). The 

fertilizer was applied to 10 plots of 5x5-m, which were interspersed in a block design 

with 10 unfertilized controls. During the first several years, when the plots were 

dominated by herbaceous vegetation, percent cover data were collected in small

0.25x0.5-m quadrats. Cover was estimated to the nearest 1 percent by species within 

a gridded quadrat frame. When shrubs began to get larger, their percent cover was 

measured using 20 m of line transects within each 5x5-m plot, and individual shrub 

species were counted in the plots to estimate their density. Forb data were collected

in April each year when the forbs were at maximum cover, and shrub data were

collected in June or July when they had achieved their maximum cover for the year. 

To calculate grass biomass, a double-sampling technique was used where grasses in 

additional 0.25x0.5-m quadrats were clipped following percent cover estimates. A 

regression equation was used to calculate the biomass of the unclipped plots. 

Exotic grass cover was significantly higher in the fertilized than the unfertilized

plots during most years between 1994 and 2001 by 10 to 18 percent, becoming 

similar in the dry spring of 1999 when cover of all species declined greatly (fig. 4).

The species composition of exotic grasses varied somewhat each year, but included

Bromus madritensis, B. diandrus, B. hordeaceous, Avena fatua, Vulpia myuros, and

Schismus barbatus. For exotic grasses we also harvested biomass, which was

considerably higher in fertilized than unfertilized plots during wet years (fig. 5). The

biomass data were a more sensitive indicator of grass response because percent cover 

did not detect productivity changes that are related to height. 

Exotic forb cover was not significantly different in fertilized and unfertilized 

plots in any year (fig. 4). The most abundant exotic forbs were Erodium cicutarium

and Hypochaeris glabra. The native forbs responded to N with increased cover 

during the first year after the fire, but this trend reversed by 1998 (P = 0.06) when

they had decreased cover with N. The species of native forbs changed considerably 

over time, starting with about 26 species of native forbs immediately after the fire 

and 12 native species by 2001. These are the typical fire-following annuals as 

reported in other local studies (Carrington and Keeley 1999). There was low cover of 

both native and exotic forbs during the dry springs of 1996 and 1999. The shrubs 

followed the opposite pattern from the exotic grasses, having significantly lower

cover following fertilization in 1996 and 1997, although no longer significantly lower 

after 1998. Shrub cover also declined during the dry spring of 1999 and did not

recover. Shrub cover was very low during the first two years following fire and was 

not estimated until the third growing season, 1996 (see fig. 4 below and on the 

following page). 

86 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution and CSS Change—Allen and others 

A. EXOTIC GRASSES

0

20

40

60

80

100

120

140

1994 1995 1996 1997 1998 1999 2000 2001

 +N

 -N

*

*

*

*

*

Figure 4—Percent cover of vegetation in plots with and without 60 kg/ha/yr N

fertilizer following the 1993 fire at Lake Skinner. Vegetation was analyzed by life form 

as A. exotic grasses, B. exotic forbs, C. native forbs, and D. native shrubs. 

*=significantly different at P=0.05. 

B. EXOTIC FORBS

0

5

10

15

20

25

1994 1995 1996 1997 1998 1999 2000 2001

 +N

 -N

C. NATIVE FORBS

0

10

20

30

40

50

1994 1995 1996 1997 1998 1999 2000 2001

 +N

 -N

*

*

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 87 
Air Pollution and CSS Change—Allen and others 

Figure 4, continued 

D. SHRUBS

0

20

40

60

80

100

120

1996 1997 1998 1999 2000 2001

 +N

 -N *

*

*

* *

*

*

0

20

40

60

80

100

120

140

160

1994 1995 1996 1997 1998 1999 2000 2001

 +N

 -N

Figure 5—Biomass (g/m2) of exotic annual grasses calculated from double-sampling

a subsample of clipped plots and using a correlation applied to permanent sample 

plots. *=significantly different at P=0.05. 

88 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution and CSS Change—Allen and others 

Figure 6—Density of dominant shrubs by species in N fertilized and unfertilized plots. 

*=significantly different at P=0.05, +=P<0.10. ARCA=Artemisia californica, 

MAFA=Malacothamnus fasciculata, ACMI=Acourtia microstachys, KEAN=Keckiella 

anthirhinoides, LOSC=Lotus scoparius, ERFA=Eriogonum fasciculatum. 

The shrubs were dominated by six species (fig. 6) that each responded differently to N fertilizer. The density of Malacothamnus fasciculatus was greatly increased by N fertilization, while that of Lotus scoparius was greatly reduced. The 

only other shrub that responded significantly was Artemisia californica, which had

increased density with N fertilization. However, Artemisia had a relatively low density compared to the two short-lived colonizers Malacothamnus and Lotus. Lotus

mortality was highest between 1998 and 1999, accounting for the large drop in shrub 

cover between those years. Over time the remaining longer-lived shrubs (Artemisia,

Keckiella, and Eriogonum) that dominate mature CSS will fill in, most likely by

growing larger. Their percent cover was not different between 1999 and 2001, all

relatively dry years, but their future rate of growth may be determined by the level of 

N fertility. 

The mechanisms by which N affected growth of the plant species in this stand

cannot be determined directly from these field data, but they include a combination 

of individual plant responses to N and yearly climatic conditions and competitive 

interactions among groups of species that respond differentially to N. The reduction

in Lotus with N is expected, because legumes typically respond to N by cessation of 

N fixation. This makes them less competitive with fast growing neighbors (see 

Munoz and Weaver 1999). In another study Artemisia showed increased growth with 

N fertilizer in the field both in monoculture and in mixture with grasses (Allen and 

others 1998). However, it had greatly reduced growth with grass competition in this 

study, and the N response was too small to overcome the competitive interaction with 

grasses. The stand studied at Lake Skinner had a relatively high cover of exotic 

grasses even without N fertilization (fig. 4). The nearby unburned stands have a

patchy understory of exotic grasses, suggesting that exotics were present before the

1993 fire. The exotics may have been introduced by domestic grazing animals, which

are responsible for dispersing seeds of many species (Malo and Suarez 1995). These

lands had been grazed by cattle up to about 20 years prior to the experiment. CSS

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 89 
Air Pollution and CSS Change—Allen and others 

shrubland in western Riverside County has been widely invaded by exotic grasses 

and forbs (Minnich and Dezzani 1998), so a site without exotics probably no longer

exits. The most important effect of N deposition is likely to facilitate exotic grass 

dominance. Drought interacted with N fertilizer in our experiment, so that there was 

no measurable response to N in dry years with low productivity. In addition, the 

timing of rain may also give an advantage to native forbs, as occurred in the

relatively dry 2001 when most rainfall occurred during February. Because the rainfall 

dynamics have resulted in patterns of plant growth and interactions with N fertilizer, 

we plan to continue fertilization and observations of these plots for additional years. 

Conclusions 

The physiological effects of ozone and nitrogen discussed above are summarized in 

figure 7, which shows that CSS, chaparral, coniferous forest, and annual grassland 

respond differently to the combined effects of these major pollutants. Ponderosa pine

forest has negative growth responses to ozone and leaf-deposited nitrate, but the 

pines still respond to N fertilization of the soil with increased growth. CSS, on the 

other hand, may be sensitive to elevated soil N as Artemisia and Encelia experienced

mortality in the greenhouse under high soil N (Allen and others 1998). Alternatively, 

CSS avoids summertime high levels of ozone and leaf-deposited HNO3/NO3

-

 because

leaves are deciduous. Chaparral growth response to N has been measured only in an 

N-limited situation, and it is uncertain how these shrubs respond to high levels of N

deposition. It is likely that chaparral would not respond to N fertilizer as well as CSS

shrubs, as they are comparatively slow growing (Hanes 1988, Mooney 1988). Based 

on these slower growth rates, we would not expect large changes in productivity in

N-impacted chaparral, although controlled studies still need to be done. CSS shrubs, 

on the other hand, continued to grow at high rates with unlimited N, water, and other 

nutrients, but died after 6-9 months (Allen and others 1998, Padgett unpublished data, 

Padgett and Allen 1999, Yoshida and Allen 2001).

 Vegetation type O3 NO3 NO3 

soil leaf dep.

 Coastal sage scrub 0 0 

 Chaparral 0 0? 0 

 Pine forest

 Annual grassland 0 0 

Figure 7—Responses of dominant species of four vegetation types to air pollution

stresses. Large and small up and down-oriented arrows represent large or small 

responses to pollutants; 0=no or minimal response. Response arrows are based on

physiological responses of individual dominant species (see text for references). 

90 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution and CSS Change—Allen and others 

The overall “winner” in this race against pollutants is the exotic annual grassland, which is replacing CSS but not chaparral or pine forest. The annual grasses 

have a high response to N fertilization, and they escape ozone and HNO3/NO3

-

 leaf

damage by summer dormancy. The summary responses of figure 7 show individual 

dominant plant growth responses and do not indicate community or ecosystem level

changes and feedbacks that may occur. For instance, during competitive interactions

between Artemisia and Bromus, the grass took up N at greater rates (Yoshida 1999).

N deposition also changed the species composition of mycorrhizal fungi in CSS 

along a nitrogen deposition gradient in the Riverside-Perris Plain (EgertonWarburton and Allen 2000). Loss of certain mycorrhizal fungal species in high N 

soils caused a reduced growth response by host plant species, indicating the fungi are 

no longer effective mutualists (Sigüenza 2000). Many other changes are occurring, 

such as increases in N mineralization and nitrification rates and increased litter buildup in N-impacted sites in the mountains (Fenn and others 1996). The forest may be 

relatively buffered from the effects of elevated N by having a longer lifespan, higher

biomass, and soil organic matter to absorb the N. In addition, the mountains 

experience yearly flushing of accumulated N in snow meltwater and precipitation 

run-off (Fenn and Poth 1999). There have been some vegetation changes, such as an

increase in Pteridium aquilinum in the understory at high N sites in the San Bernardino Mountains, but no known changes in the overstory that can be attributed to N. 

The reasons for stability of chaparral after N fertilization are not as clear but

may also be related to the open watersheds that experience high levels of N flushing

(Fenn and Poth 1999). Chaparral shrubs are long-lived, and many resprout from 

lignotubers following fire. They have slow growth rates compared to CSS and may

not respond to N fertilization with rapid growth. Finally, many stands of chaparral 

accumulate undecomposed litter except on the steepest slopes, which may enable 

immobilization of deposited N. Chaparral stands seem to be stable across the

landscape (Minnich and Chou 1997), having been affected neither by ozone die-off in 

the past like pines nor by possible N impacts as in CSS. However, ecosystems may

appear healthy until they reach a threshold of N saturation (Aber 1992). A better 

understanding of chaparral responses to N is needed before we can be certain that it

will remain healthy with continued N accumulation. The dramatic type conversion of

CSS to annual grassland should be enough evidence that air pollution levels, 

especially N deposition, need to be reduced to improve the health of the vegetation. 

Acknowledgments 

We thank Eric Focht for assistance with data analysis. The research was funded by 

the National Science Foundation, the USDA Competitive Grants Program, and the 

Metropolitan Water District. 

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Air Pollution and CSS Change—Allen and others 

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USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 95 

Fire as a Threat to Biodiversity in Fire-Type 

Shrublands1

Jon E. Keeley2

Abstract 

Chaparral and coastal sage scrub communities have a disproportionately high number of rare

and endangered plants and thus are of particular conservation concern. Unnaturally high fire 

frequency has been a leading cause of degradation of chaparral and coastal sage scrub 

ecosystems. Although these shrublands are fire-adapted, below a certain threshold of fire 

frequency, resilience is inversely related to the fire return interval: this threshold is 3-5 years 

in coastal sage scrub and 10-20 years in chaparral, with the higher values more typical of

interior sites. High fire frequency depletes the native flora and increases the proportion of 

non-native herbaceous species. Resilience to different fire regimes varies across growth 

forms, and thus it is of particular significance that the growth form distribution of rare species

is significantly different from the proportions of growth forms in these communities. 

Key words: chaparral, endangered species, high fire frequency, resilience, sage scrub

Introduction 

The California landscape has been altered in many ways, with the potential for profound impacts on biodiversity and ecosystem functioning (Keeley and Swift 1995).

Other than direct development, one of the more important changes in shrubland ecosystems has been the anthropogenic alteration of the natural fire regime. Despite a 

long-standing policy of fire suppression, the primary impact has been a dramatic acceleration of fire occurrence (Keeley and Fotheringham 2002, Keeley and others 

1999, Moritz 1997). Although species in these shrublands are “fire adapted,” they are 

not adapted to all fire regimes, and one can distinguish species differences and broad 

growth form differences in resilience to increased fire occurrence. This is of particular concern because both chaparral and coastal sage scrub communities have a disproportionately high number of rare and endangered plants. 

Shrubland Biodiversity

The California Native Plant Society (CNPS) list of rare and endangered plant species

(Skinner and Pavlick 1994) places chaparral first in number of taxa that are of 

concern (table 1). Although coastal sage scrub ranks fifth, both shrub communities 

have a much higher number of rare taxa than expected based on their area occupied. 

If all the rare and endangered taxa in the top five habitats were distributed randomly,

 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 U.S. Geological Survey, Sequoia National Park, Three Rivers, CA 93271; tel 559-565-3170 (email: 

jon_keeley@usgs.gov) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 97 
Fire as Biodiversity Threat in Shrublands—Keeley

Table 1—Top-ranking habitats of California’s rare plants according to all CNPS lists (data

from Skinner and Pavlik 1994). These observed values are contrasted with the values

expected based upon the amount of land area occupied by each habitat (data from Jones and

Stokes 1987). ____________________________________________________________________________________ 

Rank Habitat Taxa ____________________________________________________________________________________ 

 Observed Expected 

1 Chaparral 516 432

2 Lower coniferous forests 359 294

3 Cismontane woodland 311 362

4 Valley/foothill grassland 247 431

5 Coastal scrub 211 132

(p < 0.001; Π2 = 164.2 >>> Π2

0.999[3] = 16.3) ____________________________________________________________________________________ 

one would expect chaparral to top the list based just on areal extent of this vegetation, 

but chaparral contains 18 percent more species than expected. Coastal sage scrub 

contains nearly 40 percent more taxa than predicted by its areal extent. 

In both chaparral and coastal sage scrub, herbaceous perennials, typically

geophytes, top the list of growth forms that are rare or endangered (Skinner and

Pavlick 1994). If we contrast these numbers just for southern California chaparral and 

coastal sage vegetation with those expected, based on extensive surveys of postfire 

peak of diversity (Keeley 1998a), we see that rarity is not randomly distributed across 

growth forms (table 2). Annuals are very under-represented. If rarity were randomly

distributed with respect to growth form we would expect three times as many annuals 

as observed in the lists. On the other hand, there are three times as many herbaceous

perennials as expected and double the number of shrub species. 

Resilience to different fire regimes varies across growth forms (Keeley 2000), 

and thus it is of interest to consider rarity in terms of fire specialization. Fire is a 

frequent ecological factor and has played an obvious evolutionary role in these 

communities. Most non-woody plant species reach their greatest population sizes,

and thus greatest potential for spread, in the postfire environment. However, not all

species or even all growth forms are equally specialized towards fire. Annuals 

comprise a rich diversity of species that range from extreme specialization—strictly

fire-stimulated germination (Keeley and Fotheringham 1998)—to generalized

opportunistic species that occupy many types of disturbance. Cursory examination of 

the rare annuals listed in Skinner and Pavlick (1994) indicates they are not highly 

restricted to burned sites. Herbaceous perennials, particularly geophytes, in these 

shrublands lack obvious specialization to fire, and their life cycle is not substantively 

different from that of geophytes in other less fire-prone habitats (Rundel 1996). At 

the time of fire most geophytes are dormant and nearly all have transient seed banks 

that do not require specialized fire cues (Keeley1991). Thus, most rare herbaceous 

species do not appear to be highly specialized for postfire recruitment. 

This pattern is not evident in the rare shrubs, where a large proportion are 

species of Ceanothus and Arctostaphylos that have seedling recruitment strictly tied 

to postfire conditions (Keeley 1998b). Thus, it is to be expected that the direct effects 

of alterations in fire regime would affect some components of the rare plant flora 

more than others. 

98 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Fire as Biodiversity Threat in Shrublands—Keeley

Table 2—Growth form distribution of CNPS rare and endangered taxa in southern

California chaparral and coastal sage scrub (data from Skinner and Pavlik 1994). Expected

values based on distribution of growth forms following fire in 90 0.1 ha sites in southern

California (Keeley 1998a) ____________________________________________________________________________________ 

Annual Herbaceous Suffrutescent Subshrub/

perennial shrub/tree ____________________________________________________________________________________ 

Observed 43 79 17 63 

Expected 131 27 16 28 

(p < 0.001; Π2 = 202.5 >>> Π2

0.999[2] = 13.8) ____________________________________________________________________________________ 

This is well illustrated by the extreme event of back-to-back wildfires studied by

Zedler and others (1983). At four Adenostoma fasciculatum dominated sites, postfire 

frequency of this shrub increased following a fire in mature vegetation but decreased

dramatically following the second fire. The impact was most profound on the nonresprouting shrub Ceanothus oliganthus, which was present from seed in threefourths of the plots after the first fire but nearly absent after the second fire. In

contrast, these repeat fires had little impact on the herbaceous perennials Calochortus 

weedii and Dichelostemma pulchella. 

Fire suppression has been frequently cited as a major threat to fire type rare 

species, for example, the rare locoweed Astragalus brauntonii that is restricted to 

sites around the Los Angeles Basin (Skinner and Pavlick 1994). This idea is a logical 

extension of the well-documented threat of fire suppression in many western

coniferous forests (SNEP 1996). However, fire records show that in southern 

California fire suppression has not effectively excluded fire (Conard and Weise 1998, 

Keeley and others 1999). Indeed, urban mountain ranges such as the Santa Monica 

Mountains in Los Angeles and Ventura Counties have a fire rotation interval of less 

than 15 years, and this is likely many times shorter than the natural regime (Keeley

2002a). Lack of fire is not likely a threat to the persistence of these shrublands or rare 

species within them. A far greater threat in many areas, particularly the Los Angeles 

Basin, is habitat degradation due to increased fire frequency. 

Fire-Induced Habitat Degradation 

Unnaturally high wildfire frequency has long been a leading cause of degradation of

chaparral and coastal sage scrub ecosystems, second only to land development. While 

these shrublands are fire-adapted, below a certain threshold of fire frequency

resilience is inversely related to the fire return interval. This threshold is probably

about 3 years in coastal sage scrub (but longer for interior sage scrub sites) and 10-20

years in chaparral (Keeley 2000). Generally speaking, as fire frequency increases,

herbaceous vegetation is favored over woody growth forms (Sauer 1975, Wells 

1962). Numerous studies have shown that unnaturally high fire frequency depletes 

the native flora and increases the proportion of non-native herbaceous species 

(Haidinger and Keeley 1993, Zedler and others 1983). In California it is quite likely 

that, except on fine-textured argillaceous soils, grasslands are degraded shrubland 

sites (Cooper 1922, Keeley 1990, 1993; Wells 1962), and even grasslands on certain 

argillaceous soils may have been dominated by the shallow rooted summerdeciduous coastal sage scrub (Kirkpatrick and Hutchinson 1980, Wells 1962). 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 99 
Fire as Biodiversity Threat in Shrublands—Keeley

0.01

1.34

2.67

4.00

L

o

g

D

e

n

sity (#

h

/

a)

Pre-fire 1990 1994

89 fire 93 fire

C. megacarpus

C. spinosus

Figure 1—Effect of repeat fires on the populations of an obligate-seeding shrub,

Ceanothus megacarpus, and a facultatively-seeding shrub, C. spinosus, in the Santa

Monica Mountains. The first fire occurred in mature chaparral in 1989, the second in

1993 (from Keeley 2000). 

The impact of frequent fires on the native shrub populations is well illustrated

by the back-to-back wildfires that Zedler and others (1983) discussed, above. This 

extreme event resulted in the near extirpation of one obligate-seeding shrub

population, and this can occur with even longer intervals between fires. Figure 1

shows the extirpation of a non-sprouting shrub with fires four years apart, whereas a 

resprouting congener survived this high fire frequency, albeit at a lower density. 

As native shrub cover is reduced due to high fire frequency, there is typically a 

type conversion to an herbaceous community dominated by non-native species. This 

is illustrated in the study by Haidinger and Keeley (1993) showing vegetation 

changes on adjacent sites experiencing different fire regimes. For example, two fires

in six years reduced shrub populations of Adenostoma fasciculatum and Salvia

mellifera but favored the suffrutescent Lotus scoparius (fig. 2). Three fires in six 

years were detrimental to most natives but conducive to the spread of non-native 

invasives such as Brassica nigra and species of Bromus. The results of this 

chronosequence study are corroborated by similar observations on a single site over

time (fig. 3). In this study a mature coastal sage scrub site burned in 1993, and over 

the following three growing seasons the exotic cover declined as the native shrubs 

recovered. Following a second fire in 1996 the exotics exploded; within two years 

they dominated the site. 

Such type conversions of shrublands to grasslands are not always the result of 

wildfires but have long been a goal of prescription burning programs (CDF 1978, 

Sampson 1944). Sometimes these are done for “range improvement,” to increase deer

and livestock range, but other times for reducing fire hazard. These herbaceous

(grassland) associations generate far less intense fires than shrub associations, and 

these grassland fires are more safely suppressed than chaparral fires (Keeley 2002a). 

100 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Fire as Biodiversity Threat in Shrublands—Keeley

2.6

2.8

3.0

3.2

L

o

g

D

e

n

sity (#

h/ a)

Adenostoma fasciculatum

resprouts

3.0

3.5

4.0

4.5

L

o

g

D

e

n

sity (#

h/ a)

A. fasciculatum

seedlings

3.0

3.6

4.2

4.8

L

o

g

D

e

n

sity (#/h

a)

Salvia mellifera

seedlings

3.0

3.5

4.0

4.5

5.0

5.5

Lotus scoparius

2.50

2.96

3.42

3.88

4.34

4.80

Brassica nigra

3.0

3.5

4.0

4.5

Bromus spp

123 1 2 3

Number of fires in 6 years

Figure 2—First growing season shrub and herb density in adjacent stands of mixed 

chaparral and coastal sage scrub burned once, twice, and three times in 6 years 

(redrawn from Haidinger and Keeley 1993). 

Type conversion not only alters fire intensity, but also it often leads to increased

fire frequency (Keeley 2002a). This results from several factors. These herbaceous

species dry rapidly during the late spring and thus greatly expand the seasonal window of opportunity for fire. In addition, they constitute fine fuels that ignite readily 

and spread fire both horizontally through the stand and vertically into the shrubs with

little wind. As the extent of herbaceous cover increases, it sets the stage for repeat

fires in a self feed-back process where more fires thin the shrub overstory and 

increase the presence of a persistent herbaceous layer. Over time, repeating this 

process will type-convert shrublands to annual grasslands dominated by non-native 

species, which in turn alters the fire regime by increasing further expansion of

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 101 
Fire as Biodiversity Threat in Shrublands—Keeley

the seasonal window for fire because non-native species dry earlier in the spring than 

most native herbaceous species (Keeley and Fotheringham 2003). 

93 94 95 96 97 98 99

0

20

40

60

80

100

C

o

v

er b

y

N

o

n-n

at vi e

s

%( )

Fire Fire

Figure 3—Percentage ground surface covered by non-native species on a coastal 

sage scrub site in the Santa Monica Mountains burned at 21 years of age in 1993 

and reburned three years later in 1996 (data from Keeley and Fotheringham,

unpublished data). 

Sites differ greatly in their propensity for repeat fires. In general, interior sites are

far more vulnerable to frequent fires than coastal sites due to the slower rate of shrub 

recovery (fig. 4). In another study of 90 sites burned during the same week in 1993,

total herbaceous cover was positively correlated with parameters such distance from 

the coast, soil phosphorous, organic matter, and sand content (table 3). Postfire

herbaceous cover was negatively correlated with stand age prior to the fire and

estimated annual insolation. In short, herbaceous cover sufficient to carry a repeat

fire is most likely when fires occur in young stands on fertile sites in the interior. 

History of Type Conversion 

Cooper (1922) believed there was abundant evidence, based upon his observations in 

the Coast Ranges, to say that burning by Indians accounted for a shift from woody 

vegetation to grasslands. He contended that this process of type conversion continued 

with the colonization by Europeans. Bauer (1930) believed the same applied to the

grasslands of the Tehachapi Mountains farther south and cited examples of relictual 

patches of shrubland as evidence. 

102 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Fire as Biodiversity Threat in Shrublands—Keeley

10

20

30

40

50

60

70

80

C

o

v

er %( G

S

C)

a) Coastal sage scrub shrubs

herbs

-1 0 1 2 3 4 5 6

Years Postfire

0

10

20

30

40

50

60

70

80

90

C

o

v

er %( G

S

C)

b) Interior sage scrub herbs

shrubs

Prefire 

Figure 4—Foliar cover before and after fire at coastal and interior sage scrub sites 

(redrawn from O’Leary and Westman 1988). 

Today the coastal ranges south of San Francisco have 25 percent of their 

landscape dominated by alien grasses, and there is reason to believe this derives from 

shrubland type conversion beginning with Native Americans (Keeley 2002b). Brown 

and Show (1944) recounted the history of rural land use in California, which included

the use of fire to convert “useless” brush to “more productive” grasslands. This type 

conversion process accelerated in the latter part of the 19th century with increasing 

competition for suitable grazing land. During the latter part of the last century such 

type conversion was officially sanctioned by the issuing of brush burning permits for 

“range improvement” by the California Division of Forestry (CDF 1978).

Throughout the 20th century, type conversion of shrublands occurred from other types 

of disturbance as well (Stylinksi and Allen 1999). 

Presently, we have relatively limited information on the extent of such type 

conversions. However, there are reasons to believe this was done on a massive scale.

One is the extensive distribution of non-native grasslands in the Coast Ranges of

central California and the lack of any obvious correlation with environmental 

parameters such as soil type or slope exposure (Wells 1962). Another is the 

widespread distribution of grasslands in the foothills east of San Francisco and the

demonstration of their rapid conversion to woody vegetation upon cessation of 

grazing (McBride and Heady 1968). Finally, quantitative measures of type

conversion in southern California coastal sage scrub indicate that over the past 60

years more than half of the vegetation has been partially or completely type

converted to grassland (Minnich and Dezzani 1998). 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 103 
Fire as Biodiversity Threat in Shrublands—Keeley

Table 3—Stepwise multiple regression of postfire herbaceous cover vs. environmental

variables at 90 sites of coastal sage scrub and chaparral (from Keeley and Fotheringham

unpublished data). 

_____________________________________________ 

r

2

 = 0.42 P < 0.001 

2-tail P

Positively correlated: 

Distance from coast 0.000 

Percentage sand 0.002 

Soil phosphorus level 0.003 

Soil organic matter 0.009 

Negatively correlated: 

Prefire stand age 0.002 

Annual solar insolation 0.033

_____________________________________________ 

Conclusions 

Besides human development, the greatest threat to biodiversity in Mediterraneanclimate shrublands of California is changes in the natural fire regime. Fire

suppression alters the fire regime by increasing the fire return interval and potentially

threatens the persistence of species with fire-dependent recruitment. However, this

impact is offset by the abnormally high number of fire occurrences due to humancaused ignitions. Consequently, the fire return intervals in southern California 

shrublands are higher than was historically the case, and the greatest threat to the

persistence of these vegetation types is type conversion to herbaceous species more

resilient to and more conducive to frequent fires. In general, non-native grasses and 

forbs are most favored by the current fire regime of frequent fires. 

In general the majority of rare herbaceous species show no specialized response

to fire. However, areas heavily affected by human-induced acceleration of fire return 

intervals may not be suitable sites for these natives because of the intense

competition from non-native invasives that are favored by high fire frequency.

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pubs/es.html

Stylinski, C.D.; Allen, E.D. 1999. Lack of native species recovery following severe exotic

disturbance in southern Californian shrublands. Journal of Applied Ecology 36: 544-554.

Wells, P.V. 1962. Vegetation in relation to geological substratum and fire in the San Luis

Obispo quadrangle, California. Ecological Monographs 32: 79-103. 

Zedler, P.H.; Gautier, C.R.; McMaster, G.S. 1983. Vegetation change in response to 

extreme events: The effect of a short interval between fires in California chaparral 

and coastal scrub. Ecology 64: 809-818. 

106 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Recovery Research for the Endangered 

Pacific Pocket Mouse: An Overview of 

Collaborative Studies1

Wayne D. Spencer2

Abstract 

The critically endangered Pacific pocket mouse (Perognathus longimembris pacificus), feared 

extinct for over 20 years, was “rediscovered” in 1993 and is now documented at four sites in

Orange and San Diego Counties, California. Only one of these sites is considered large 

enough to be potentially self-sustaining without active intervention. In 1998, I gathered a team

of biologists to initiate several research tasks in support of recovery planning for the species.

The PPM Studies Team quickly determined that species recovery would require active translocations or reintroductions to establish new populations, but that we knew too little about the 

biology of P. l. pacificus and the availability of translocation receiver sites to design such a 

program. Recovery research from 1998 to 2000 therefore focused on (1) a systematic search 

for potential translocation receiver sites; (2) laboratory and field studies on non-listed, surrogate subspecies (P. l. longimembris and P. l. bangsi) to gain biological insights and perfect

study methods; (3) studies on the historic and extant genetic diversity of P. l. pacificus; and 

(4) experimental habitat manipulations to increase P. l. pacificus populations. Using existing 

geographic information system (GIS) data, we identified sites throughout the historic range

that might have appropriate soils and vegetation to support translocated P. l. pacificus. Reconnaissance surveys of habitat value were completed in all large areas of potential habitat

identified by the model. Those sites having the highest habitat potential are being studied with 

more detailed and quantitative field analyses. The surrogate studies helped us design individual marking and monitoring methods and will be used to test translocation methods before

applying them to P. l. pacificus. Genetic results suggest that P. l. pacificus populations were 

naturally fairly isolated from one another prior to modern development, that genetic diversity

will continue to erode in the small populations that remain, and that individuals from extant

populations could probably be mixed if maximizing genetic diversity in any newly established

populations is an important recovery goal. Local populations should be increased in situ before they can supply donor animals for translocations. Experimental habitat management 

(shrub thinning) at one occupied site yielded a short-term, positive, behavioral response of

mice to thinned habitat plots. However, the overall population seems to be in decline, and 

long-term population responses to habitat manipulations are not yet evident. The approach of

the PPM Study Team has been to proceed cautiously and scientifically to obtain critical information and to design a translocation program, but we are prepared to recommend swift

action to prevent extinction despite “insufficient data.” At this point, political and economic

obstacles to species recovery seem larger than obstacles presented by scientific uncertainty. 

Key words: endangered species, Pacific pocket mouse, Perognathus longimembris pacificus,

recovery, reintroduction, translocation

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 Senior Conservation Biologist, Conservation Biology Institute, 815 Madison Avenue, San Diego, CA 

92116 (e-mail: wdspencer@consbio.org) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 107 
Pacific Pocket Mouse Recovery Research—Spencer 

Introduction 

Recovery planning for endangered species requires a pragmatic use of science that 

recognizes we cannot always expect significant sample sizes or fully informed

decisions. In recovery research, the goal of sustaining and increasing populations 

must sometimes overrule clean experimental design, and decisions must often be 

made based on untested hypotheses, lest we observe extinction while we await more

data. The overall guiding principal should be to proceed cautiously with good

science, but be prepared to act swiftly and decisively to avoid extinction, using

whatever imperfect information you have. 

Recovery planning for the Pacific little pocket mouse (Perognathus 

longimembris pacificus, commonly referred to as PPM) offers a useful example of 

this approach. This paper briefly summarizes background information on the species,

its history of decline, and initial studies in support of species recovery. I then review 

methods and preliminary results of some recent research tasks coordinated under the 

informal umbrella of the “PPM Studies Program.” Finally, I review current status of 

these studies and summarize future directions for the program. 

Background 

Perognathus longimembris pacificus is the smallest subspecies of the little pocket

mouse (Perognathus longimembris), a nocturnal, burrow-dwelling, mostly

granivorous, heteromyid rodent species restricted to arid southwestern North 

America (Hall 1981, Williams 1986, Williams and others 1993). It is unusual as the 

only subspecies or species of Perognathus found on the Pacific coast, rather than in

arid inland deserts and grasslands. Historically, this subspecies occurred on fine,

sandy soils within about 4 to 6 km of the Pacific coast of southern California from

near the Mexican border to El Segundo in Los Angeles County (Grinnell 1933,

Meserve 1976, Patten and others 1998, von Bloeker 1931). Since the 1930s, most 

historic habitat has been removed or fragmented by urban development and

agriculture, with the last known population being extirpated by development in 

Newport Beach during the early 1970s (Patten and others 1998). After approximately

20 years, during which time the species was not detected, a small population 

(tentatively estimated at 25-36 individuals) was rediscovered on the Dana Point

Headlands in 1993 (Brylski 1993). 

The subspecies was emergency listed as endangered by the U.S. Fish and 

Wildlife Service (USFWS) in 1994 due to immediate threats to this tiny, remnant 

population (USFWS 1998). Subsequently, three additional population sites have been

discovered on Marine Corps Base Camp Pendleton, San Diego County (Ogden 1995, 

1997; MBA 1997). Despite extensive survey efforts throughout the range, these four 

sites (fig. 1) represent the full extent of the current known subspecies range (USFWS

1998, Spencer and others 2000a, 2000b).

Criteria for species recovery include permanent protection of ten independently

viable populations (USFWS 1998). Currently, three of the four occupied sites (Dana 

Point, San Mateo North, and San Mateo South) are less than 12 ha each in size, and 

their populations are thought to number under 50 individuals each (Spencer and 

others 2000a, 2000b). Such small sizes predispose these populations to extirpation by 

stochastic events, catastrophes, inbreeding depression, or other factors (Noss and 

Csuti 1997). At present these three sites cannot confidently be called viable. The 

108 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

fourth and largest site (Oscar One/Edson Range) approaches 900 ha in total area (a 

mosaic of occupied, suitable patches and generally unoccupied, unsuitable matrix) 

and supports hundreds of individuals (Pavelka personal communication, Dodd and

Montgomery 1998, Spencer and others 2000a). Although the Oscar One/Edson 

Range site is also vulnerable to extirpation, it is considered sufficiently large to be

sustainable over at least the near term, barring catastrophe. 

Initial Studies and Founding of the PPM Studies Program

Since the listing of the species, various studies were initiated to monitor P. l. 

pacificus populations, improve field study methods, evaluate potential project impacts, and improve understanding of the species’ biology. The USFWS has been

studying the P. l. pacificus population, monitoring methods, and microhabitat selection at the Oscar One/Edson Range site since 1997, and S. Dodd and S. Montgomery 

Figure 1—Historic (open circles) and extant (closed circles) locations occupied by P.

estimates of population size are yet available. 

l. pacificus. For historic locations, parentheses indicate the number of specimens in

museum collections and the years of their collection. For extant populations, the 

numbers of individuals captured in recent years are shown. No capture-recapture 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 109 
Pacific Pocket Mouse Recovery Research—Spencer 

have performed yearly population monitoring at Dana Point since 1997. Dodd and

Montgomery (1998) also investigated effects of trap spacing and trapping duration on

WS and California Department of 

Fish

ific experts, and performed very limited pilot 

stud

(198

monitoring effectiveness at the Oscar One site. 

In 1997, the Foothill/Eastern Transportation Corridor Agencies entered into a

Memorandum of Agreement (MOA) with the USF

and Game to fund several research tasks specifically called for by the species’

Recovery Plan (USFWS 1998). In response, I assembled a team of biologists (including C. Schaefer, S. Montgomery, and S. Dodd), which in July 1998 initiated four 

specific research tasks called for by the MOA: (1) evaluate the feasibility of translocating P. l. pacificus to establish new populations, (2) determine dispersal characteristics of P. l. pacificus, (3) evaluate effects of Argentine ants on P. l. pacificus, and

(4) develop a P. l. pacificus translocation protocol. We also assembled a scientific

peer review committee to review progress reports and study designs throughout the 

process. Participants on the committee have included Drs. James Patton, William 

Longland, Jay Diffendorfer, Howard Shellhammer, Nicholas Holler, Michael

Wooten, and Richard Reading. 

During 1998, the study team reviewed existing literature pertaining to the four

initial research tasks, interviewed scient

ies. Based on this information, we concluded (Spencer and others 2000a) that (1) 

a translocation or reintroduction program is probably necessary to achieve species

recovery in the long term, but we knew too little to attempt such a program in the

short term; (2) knowledge of P. l. pacificus dispersal characteristics would be useful, 

but a study specifically designed to answer dispersal questions would be prohibitively

expensive, technologically impractical, and might do more harm than good to extant 

populations; (3) Argentine ants are unlikely to be a major limiting factor on P. l. 

pacificus due to differing habitat affinities (Argentine ants require moist ground conditions and are generally associated with irrigated landscapes in southern California 

[Suarez and others 1998, T. Case personal communication], whereas P. l. pacificus

require well-drained and open vegetation communities); and finally, (4) we felt we 

could not design a translocation program until critical information gaps were filled. 

Table 1 summarizes information on P. l. pacificus available in 1998 relative to

the information needed to design a translocation program, according to Nielsen

8) and USFWS (1998). The information gaps evident in table 1 were used to 

define several research tasks deemed of highest immediate priority to further 

recovery: (1) a systematic review of potential translocation receiver sites based on 

habitat suitability and land uses; (2) a controlled laboratory study using a non-listed 

subspecies of P. longimembris to derive a preferred individual marking technique; (3) 

an intensive, multiyear study of a non-listed subspecies to test field techniques, 

before using them on P. l. pacificus, to gain basic biological information on the

species and to test potential translocation methods; and (4) a population genetics 

study to establish baseline measures of genetic diversity within and between extant

and historic P. l. pacificus populations. Once results of these studies make it possible 

to perform more intensive field studies on P. l. pacificus that would do more good

than harm, we recommended (5) an intensive multiyear study of P. l. pacificus to 

obtain essential demographic and other natural history data. Finally, (6) we also 

recommended experimental habitat management, particularly controlled burning or

shrub thinning, to improve habitat quality and P. l. pacificus population sizes at one 

or more occupied sites. 

110 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

Table 1—Translocation feasibility considerations for P. l. pacificus as of 1998 based on

USFWS (1998) and Nielso

cient information exists to as

n (1988). A “yes” under “sufficient information?” indicates suffisess whether translocations are necessary or feasible, but does

not imply that information is sufficient to design a translocation program. 

Consideration

Sufficient 

information? Notes and comments

Reason for 

translocation 

Yes Recovery criteria probably cannot be met without a

translocation/reintroduction program. 

Status of wild 

populations

Yes Severely in danger of extinction. Most potential habitat has 

been surveyed: four known occupied locations, representing

three or four isolated populations. All remaining populations

No 

information available on related subspecies. 

density of 

wild populations

orth=37;

Movements and

distribution 

No/Yes

.

o

itment No 

ecies. 

uspected

ions s. 

ability

aining No potential habitat areas remain,

anslocation

receiver sites (this study). 

areas

Translocation No udy.

ue Yes r sites are relatively simple to

and No Systematic review of potential receiver sites is needed to

define number and relative distribution of desirable

rt of

translocated 

No 

11 unique captures on Oscar One (USFWS 1998) plus 20 unique individuals on Edson

gomery and Dodd 1998 unpublished data). Additional captures represent an unknown 

f additional individuals because they were not marked to identify recaptures.

small. 

Insufficient information on habitat requirements,

demographics, space-use, dispersal patterns, and so forth.

Sparse

Biology

Size and No Rough estimates (individuals captured within a single

season) available for four sites: Oscar One/Edson

Range=1311

; San Mateo South=20; San Mateo N

Dana Point=21 to 36. No capture/recapture population

estimates. 

Little available information on movements of P. l. pacificus

or related subspecies. Geographic distribution fairly well 

established

Male: female rati

Annual recru

Yes Dana Point: 1.1:1 to 1.4:1; San Mateo North: 0.8:1. 

No information on P. l. pacificus; limited information on

related subsp

Mortality rate and 

causes 

No No information on rates; some causes known or s

from limited observations. 

Health and Little information

condition of 

populat

No besides external observation of wildcaught specimens in hand. Most individuals appear healthy

and free of external parasite

Genetic vari

and integrity

Status of rem

No No information. To be determined by this study. 

Although it is known that few

habitat systematic review needed to identify potential tr

Depredation rates 

in target

No Limited information on presence or abundance of predators 

in potential receiver sites. 

To be determined by this st

strategy

Capture techniq

and technology

Number

Capture methods in dono

define. 

location of release 

sites 

determine adequacy (this study). Recovery criteria will be

used to

sites. 

Post-release To be determined by this study. 

suppo

animals 

1 Includes 1

Range (Mont

number o

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 111 
Pacific Pocket Mouse Recovery Research—Spencer 

The remainder of this paper focuses on methods and preliminary results of those 

primary research tasks conducted to date, focusing on the translocation receiver site

study, laboratory-marking study, field marking study, population genetics study, and

shrub thinning study.

eceiver Site Study

The search for sites to receive translocated populations uses a coarse-filter/fine-filter 

first developed a simple habitat evaluation model based on existing 

stem (GIS) data layers to identify, at a coarse scale, large

inations of soil and vegetation to 

nding to overestimate potential habitat value for P. l. pacificus to 

mini

m of the coast (measured 

from

izon texture, soil depth, 

and

Methods

Translocation R

approach. We

geographic information sy

areas near the coast having appropriate comb

possibly support translocated P. l. pacificus. S. J. Montgomery and S. Dodd then 

surveyed these areas in the field to better rank habitat conditions using a quick 

evaluation method developed for this purpose (coarse filter). Those sites having the 

highest rankings based on the field reconnaissance will be subject to more intensive, 

quantitative studies (fine filter). These efforts are restricted to coastal Orange and San 

Diego Counties, because soils maps are not available for Los Angeles County, and 

remaining native coastal areas there seem generally unsuitable for establishing new

P. l. pacificus populations due to inappropriate soils, habitat degradation, abundant 

exotic species, and other factors (W. Spencer, personal observations; P. Behrends, S. 

Montgomery, and M. Pavelka, personal communications; Ogden and Dames and 

Moore 1999). 

The GIS habitat model was used to screen out areas of Orange and San Diego 

Counties that are clearly unsuitable for P. l. pacificus reintroduction and to identify

areas deserving field evaluation. The model was therefore designed to be

conservative, te

mize chances of missing a suitable site. This conservative approach recognizes 

that available digital information for soils and vegetation is coarse in scale relative to 

P. l. pacificus habitat selection and may have inaccuracies. 

The model ranks P. l. pacificus habitat potential based on three criteria: distance 

from the Pacific coast, soil type, and vegetation type. Nearly all historic observations 

of P. l. pacificus are within 4 km of the Pacific coast, with a few as far as 6 km, so 

inland limits of the study area were conservatively set at 8 k

 the main shoreline or the eastern shore of larger bays). 

W. Spencer and M. Pavelka—in collaboration with Dr. Robert Graham,

Professor of Soil Mineralogy and Pedology at U.C. Riverside—ranked all soil types

mapped within this study area (USDA 1973, 1978) as having very high, high, low, or 

no potential to support P. l. pacificus based on surface hor

other factors (for example, presence of a hardpan or high gravel or cobble

content). Care was taken to consider all physical factors likely to affect burrowing or 

foraging abilities of P. l. pacificus, as well as the potential for higher quality soils to

exist as inclusions within lower-ranked soil polygons. The highest-ranking soils were 

classified based on results of detailed soils analyses performed by the USFWS at 

occupied versus unoccupied microhabitats at the Oscar One P. l. pacificus site 

(Pavelka, Winchell, and Graham, unpublished data). Those results suggest that P. l.

pacificus are very discriminating of soil characteristics at a finer scale (measured on 

20 x 20-m grids) than that of existing soils maps, and that they strongly select for

112 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

deep (>30 cm), fine-grained, loamy sands (not sandy loams) having less than about 7 

percent by volume of clay and less than 4 percent by weight of gravel and rock. 

Vegetation types were ranked as having high, low, or no potential to support P. 

l. pacificus using regional GIS vegetation maps (SANDAG and Ogden 1997 for San 

Diego County; Orange County Planning and Development Services Department 1993 

for Orange County). P. l. pacificus historically have been associated with open

coas

coast. As long as the soils are highly suitable to 

supp

ion, soil, and

othe

vegetation suitability, all coastal scrub, grassland, and dune vegetation 

ommunities were ranked high; all other natural vegetation communities were ranked low; all

tal scrub and grassland habitats having a diversity of annual herbs, so these

vegetation communities were ranked high. P. l. pacificus probably cannot persist in 

denser shrub communities, wetlands, or woodlands. However, the regional vegetation

maps are coarse in scale, and some upland areas mapped as generally unsuitable

natural vegetation may contain small areas of open scrub or grassland. Therefore all

natural vegetation types were considered to have at least a low potential to support P. 

l. pacificus. Areas mapped as agriculture, developed, disturbed, rock, or open water 

were ranked as having no potential. 

The soil and vegetation rankings were combined to create overall habitat

rankings within 8 km of the coast (table 2) based on the following logic: Evidence to

date suggests that soils are the primary determinant of habitat quality, at least within 

areas of natural vegetation near the

ort P. l. pacificus, low suitability vegetation (e.g., chaparral) may reduce P. l. 

pacificus habitat potential somewhat, but vegetation management (e.g., burning or 

thinning) might be used to increase habitat quality. Therefore, the presence of low

quality vegetation on high quality soils only slightly reduces the overall habitat

potential to support P. l. pacificus. However, if the soils are low quality for P. l.

pacificus, the vegetation community type or vegetation management will probably 

have little effect on P. l. pacificus habitat potential; consequently, the overall rating 

reflects the soil potential (low). If either soils or vegetation are inappropriate (e.g., 

clay soils or agricultural fields), the site has no P. l. pacificus potential. 

Once maps of the overall habitat potential rankings were created, relatively large

(> 20 ha) blocks of moderate to very high habitat value were outlined as field 

evaluation sites. S. J. Montgomery and S. Dodd then evaluated these sites in the field 

during 1999-2000 using a rapid evaluation data sheet to record vegetat

r site characteristics. The 5-10 sites receiving the highest scores during these

coarse-level field evaluations are currently subject to more intensive, quantitative 

(fine-filter) study to determine their potential to support translocated P. l. pacificus

populations. 

Table 2—Habitat potential to support P. l. pacificus based on soils and vegetation. See text 

for how soils were ranked (a complete list of soil rankings is available from W. Spencer upon

request). For

c

developed, disturbed, agricultural, rock, and open water areas were ranked none.

 Soil suitability

Vegetation

suitability Very high High Low None 

High very high high low none 

L gh moder ow none 

none none none none 

ow hi ate l

None

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 113 
Pacific Pocket Mouse Recovery Research—Spencer 

Lab ry Marking Study 

marking methods o ptive popul of a 

bris: (1) subcutaneous passive integrated trans-

 draw. Marking was done with the assistance and oversight

of a

ts or behaviors. The examination schedule was every other day for 

the f

ield Marking Study 

ld surrogate study during 1999-2000 to investigate longth effects in the wild, collect basic natural history and 

d B based in part 

on t

orato

We tested four permanent, individual

non-listed subspecies, P. l. longimem

n a ca ation

ponder (PIT) tags, (2) toe clipping, (3) tiny, custom-designed ear tags, and (4) colorcoded surgical sutures. P. l. longimembris was chosen as a logical surrogate for P. l.

pacificus due to close phylogenetic and body-size relationships and its non-sensitive

status. S. Dodd captured 36 P. l. longimembris in Sherman live traps near Yucca 

Valley, San Bernardino County, California—28 during June 1999 and 8 more during 

May 2000. The animals were housed individually in 19 x 21 x 32-cm plastic cages 

supplied with natural desert soils in a room having a skylight (for natural day/night

cycles) and an air filter. 

Captive animals were assigned to one of the four marking techniques or to a 

control group by random

veterinarian (Dr. Anthony Michael). Six animals were marked in one ear with

custom-designed aluminum ear tags bearing 3-digit numbers etched into their surface. Six animals were marked with a 2-digit toe clip code (one digit on a front paw

and one on a rear paw) using the standard mammalogical method (Baumgartner

1940, Melchior and Iwen 1965). Toes were clipped distal to the ultimate toe joint,

only far enough back to remove the nail and the cells producing nail growth, using

either a baby nail clipper or surgical scissors. Six animals were marked with two

color-coded monofilament sutures (size 4-0 monofilament nylon surgical sutures, one

black, one blue) stitched through the loose skin on the nape of the neck, tied with a 

triple surgeon’s knot, and cut to leave 1.0 to 1.5 cm of loose thread projecting from 

the knot. Seven animals were tagged subcutaneously between the scapulae with PIT 

tags that measure 11.5x2 mm. Six tags were injected using a 12-gauge hypodermic 

needle, and one tag was inserted by hand via a surgical incision across the nape after

shaving a patch of hair. Nine animals served as unmarked controls. They were 

removed from cages, weighed, and examined according to the same schedule as the

marked animals. 

After marking, the test and control groups were regularly weighed and examined for

adverse health effec

irst two weeks, then weekly for ten more weeks, and at least monthly thereafter. 

F

The study team initiated a fie

term mark retention and heal

demographic data, and to test various translocation techniques (for example, use of

enclosure fences, artificial burrow design). We established two permanent study grids 

near the village of Snow Creek in the Coachella Valley, Riverside County, California 

(San Bernardino Base and Meridian; USGS 7.5’ White Water Quadrangle, Section 21

of Township 3 South, Range 3 East). Habitat on the site is relatively homogeneous 

creosote- (Larrea tridentata) dominated desert scrub on a north-sloping bajada,

which is densely occupied by the non-listed subspecies P. l. bangsi. 

Two trapping grids (Grids A and B) were established approximately 80 m apart, 

with inter-trap spacing initially set at 5 m for Grid A and 7 m for Gri

he trap-spacing study for P. l. pacificus performed by Dodd and Montgomery

(1998). Grid A consisted of 14 x 14 trap stations (196 total stations) and Grid B of 10 

x 10 trap stations (100 total stations). The area trapped by each grid was thus 0.49 ha

114 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

(70 x 70 m), assuming a 1⁄2-trap-space perimeter strip around each grid. Each station

was set with two traps to reduce trap saturation by non-target species (especially

kangaroo rats, which often enter traps early in the evening). Thus, Grid A contained

392 traps at 196 stations and Grid B contained 200 traps at 100 stations, for a total of 

592 traps per night. 

The grids were trapped for eight consecutive nights from 27 July to 3 August 

1999 to mark residents and search for an asymptote in capture of new individuals. 

We

ce), location, recapture 

statu

pacing. We began marking

indiv

Population Genetics

i at UC Berkeley analyzed geographic genetic structure

species of little pocket mice in southern California (P.

also trapped the grids for two nights on 1 and 2 September 1999 to observe 

health and mark retention on previously marked individuals, and to see if the mice

were still active above ground so late in summer (seasonal timing and influences for 

entering and leaving aestivation/hibernation are poorly understood in P.

longimembris). All captured individuals were marked by a unique, 2-digit toe-clip

combination. A subset of the mice was also marked by attaching a uniquely

numbered ear tag to the left ear or a PIT tag inserted subcutaneously using the 

hypodermic needle technique. Toe clipping provided a redundant mark for animals 

also marked with ear or PIT tags, in case these tags were lost. 

We recorded the following data for all captures: sex, weight, reproductive 

condition, age (young of the year or adult, by pelage appearan

s, and notes on animal condition and condition of any previously applied marks.

Reproductive condition was based on the position of the testes for males and

condition of the vagina and size of mammae for females. 

Trapping was continued during 2000—for five nights in April and four nights in

July—with both grids trapped at the 5-m (14 x 14) trap s

iduals with either toe clipping or a PIT tag, but discontinued ear tagging due to

high loss rates for ear tags.

James Patton and Andrea Swe

within and among the five sub

l. bangsi, brevinasus, internationalis, longimembris, and pacificus) by examining

haplotype variation in their mitochondrial cytochrome b gene (Swei and others

2003). The 111 total P. longimembris samples they analyzed included 48 P. l.

pacificus specimens collected in 1995-2000 (by pulling hairs or clipping toe tips from 

live-caught individuals) and 12 P. l. pacificus museum skins collected in the 1930s

(from 2 x 2-mm skin snips). Full 810-base pair (bp) sequences were obtained from 

the live-caught specimens, and 430-bp fragments were amplified from the museum

specimens. To date the following live-caught P. l. pacificus samples have been 

analyzed: Dana Point (N=27), San Mateo South (N=10), San Mateo North (N=5), 

and Oscar One (N=6). The historic museum specimens were collected from El 

Segundo (N=1), Oceanside (N=6), and Tijuana Estuary (N=5). Swei and others 

(2003) applied a variety of analytic and statistical measures to describe phylogenetic 

and population genetic structure within and among the five subspecies and to infer

population genetic processes affecting P. l. pacificus populations historically and 

today (see Swei and others 2003, and citations therein, for complete laboratory and 

analytical methods). 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 115 
Pacific Pocket Mouse Recovery Research—Spencer 

Shrub Thinning Experiment

irical evidence suggests that vegetation has

nged in 4 x 10 arrays with 3-m spacing) were 

estab

thinning, shrub canopies on the experimental plots covered a mean of 

abou

n for seven consecutive nights, once during spring (late Aprilearly

Results 

tion Receiver Site Study

tat suitability model for the northern 

A variety of theoretical and emp

become too dense at some of the occupied P. l. pacificus sites for optimal habitat 

condition, and that thinning the vegetation by fire or other means might increase 

habitat value and P. l. pacificus population sizes. Experimental habitat thinning was 

therefore initiated within the Dana Point P. l. pacificus reserve during 1999. It was 

not possible to obtain permission for a controlled burn at the site, so shrub cover was 

thinned by hand on experimental plots. 

Eleven trapping grids (40 traps arra

lished and trapped during the spring and summer of 1999 (pre-thinning) and

2000 (post-thinning): three control grids and eight experimental grids. The three 

control grids were established in known, occupied P. l. pacificus habitat having

relatively low shrub cover; the eight experimental grids were established in areas of

dense shrub cover not known to harbor P. l. pacificus but having appropriate soils. 

The experimental grids overlapped four roughly rectangular shrub-thinning plots of 

about 30 x 60 m each, such that half of each of the experimental grids was in an area 

to be thinned and half in an area not to be thinned (hereafter, thinned and unthinned 

subgrids). 

Before

t 83 percent (range 75-87 percent), measured at 21-60 cm aboveground using

point-intercept transects. During January 2000, shrub cover was thinned on the four 

plots to a mean of 32 percent (range 26-36 percent). Thinning was done by removing

individual shrubs at ground level, using chainsaws and pruning tools to avoid soil 

disturbance. Concentrations of woody duff and litter were also lightly raked to

uncover mineral soils. 

Trap lines were ru

 May) and once during summer (mid August). Captured P. l. pacificus were

individually marked by clipping hair and using indelible colored markers (during

1999) or by toe clipping (during 2000). Use by P. l. pacificus of the control, thinned,

and unthinned grids was compared between the pre-thinning (1999) and postthinning (2000) periods using Chi-square. Spring and summer trapping sessions were

pooled within each year, and the sample frequencies were calculated as the

cumulative sum of all unique individuals captured on a grid. Thus, a recapture of an 

individual on the same grid was not counted as an independent observation, but a 

recapture on another grid was, because individuals could choose to move from one

grid (for example, a control grid) to another (for example, a thinned grid) during the 

study.

Transloca

Figures 2 and 3 illustrate output of the GIS habi

and southern halves of the study area, respectively. The maps also indicate general

areas investigated in the field for their potential to support translocated P. l. pacificus. 

These maps surely overestimate the actual extent of high and very high quality 

habitat, due to the conservative model assumptions and the gross nature of the digital 

map information relative to the scale at which P. l. pacificus select habitats. Although

the model correctly identified all extant P. l. pacificus population sites as having very

116 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

high habitat quality, it missed some areas of apparently high suitability due to

inaccuracies in the soils data. Nevertheless, the maps proved useful in focusing field 

efforts in areas with the highest likelihood of having suitable soils and vegetation to 

support translocated P. l. pacificus. Reconnaissance surveys indicate that six or seven

sites may have sufficient acreage of fine, sandy soils and appropriate vegetation to 

support translocated P. l. pacificus populations. 

Areas in Orange County and Northern San Diego County having potential 

potential due to inaccuracies in model input data. 

Figure 2—

to support translocated P. l. pacificus populations, based on the GIS habitat suitability 

model. Boxes indicate general areas assessed in the field. Note that these maps 

greatly overestimate actual habitat suitability in many areas due to conservative 

model assumptions and coarse data input scales. They may also miss areas of 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 117 
Pacific Pocket Mouse Recovery Research—Spencer 

Figure 3—Areas in San Diego County south of Camp Pendleton having potential to

port translocated P. l. pacificus populations, based on the GIS habitat suitability 

Laboratory Marking Study 

tested, toe clipping seemed clearly the least

for uniquely identifying animals over their life

sup

model. Boxes indicate general areas assessed in the field. Note that these maps 

greatly overestimate actual habitat suitability in many areas due to conservative 

model assumptions and coarse data input scales. They may also miss areas of 

potential due to inaccuracies in model input data. 

Of the various marking techniques

harmful and possibly the most reliable

span. The suture technique and ear tags resulted in high rates of tag loss (100 percent 

and 50 percent, respectively) and other complications, and they are not recommended 

118 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

for further testing. PIT tags seem promising and are recommended for further testing 

on surrogates. However, the puncture wound for tag insertion is large for such small 

animals, and once inserted, PIT tags sometimes migrate under the skin or cause 

rumpling of the skin and hair. PIT tags would offer substantial benefits over toe

clipping if they could be made smaller to reduce potential adverse health effects; they 

are not recommended for use on P. l. pacificus in their currently available

dimensions. 

Field Marking Study 

of male and female P. l. bangsi on the two grids in 

corded on grids using 5-m trap spacing (Grid A during 

his reflects a large reproductive output in 2000, with 126 of 162 aged individu

ble 3—P. l. bangsi captures at the Snow Creek study grids during 1999 and 2000. 

Table 3 summarizes captures

1999 and 2000. Sex ratios re

both years and Grid B during 2000) did not differ significantly from each other nor 

from 1:1 (χ2

=0 to 0.5; df=1; P> 0.25 for all comparisons). However, during 1999, 

when Grid B used 7-m inter-trap spacing, few females were captured and the

recorded sex ratio strongly favored males (24 males:9 females=2.67:1). This ratio

almost differed significantly from that recorded on the same grid using the 5-m 

spacing in 2000 (χ2

=3.79; df=1; P=0.051), and it did differ significantly from an even 

ratio (χ2=6.82; df=1; P< 0.01). These results support a hypothesis that the 7-m 

spacing may have under-sampled females on Grid B during 1999 (compare the low

number of females captured in this treatment relative to all others in table 3). 

Although we recognize that this conclusion is weak due to a lack of grid replication, 

reduced female captures at greater trap spacing would be expected if females move

less extensively than males, which is true in most pocket mice (review in Jones

1993). 

Captures on both grids more than doubled between summer 1999 and summer 

2000. T

als (78 percent) identified as young of the year during summer 2000. In contrast, 

we identified all individuals captured during April 2000, and all but one individual

captured during July-August 1999, as over-wintered adults. The reproductive season 

seemed near a peak during April 2000, with most individuals showing signs of sexual 

activity, including descended testes in males or perforate vaginas in females. By JulyAugust of both years, few individuals showed signs of sexual activity, and we were 

able to distinguish three age classes: adult, subadult, and juvenile. 

Ta

Sex 

Trap Session Male Female Total

Grid A 

 Summer 1999 22 22 44 

 Spring 2000 1 2 3

(4 no ed)

 Summer 1999 24 9

g 2000 1 1

(2 no ed)

5 0 5

 Summer 2000 53 40 97 

t sex

Grid B 1 33

 Sprin 7 5 32

 Summer 2000 37 32 71 

t sex 1 Traps on Grid B used 7-m inter-trap spacing in summer 1999; otherwise, all traps had 5-m inter-trap 

spacing. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 119 
Pacific Pocket Mouse Recovery Research—Spencer 

When both years are pooled, 248 individuals were toe clipped, 34 were PIT 

tagged, and 50 were ear tagged. All ear tags were applied in 1999, be rate

sion, at least a third of the ear-tagged individuals exhibited various levels of 

ear i

animals rarely showed inflammation or redness of the wound. 

Nev

ere too low 

(ther

all P. l. pacificus haplotypes analyzed to date are

unique to that subspecies, and all are closely linked genealogically. P. l. pacificus has 

apparently been diverging from its geographically closest relatives (P. l. brevinasus, 

) for some time, although none of these 

ther subspecies or populations. Of the total 

gene

cause a high

of tag loss caused us to discontinue this method in 2000. By the end of the first 1999 

trap ses

nflammation and redness, ripping around the tag insertion hole, or drooping ears.

One individual had lost its ear tag when it was recaptured 7 days after marking. At 

least seven more ear tags were lost by the second trapping session (4-5 weeks after

tagging). 

Toe clipping was the fastest and easiest marking technique under field conditions and seemed to have little or no adverse effects on animal health. Clipped toes 

exhibited no or little bleeding (one or a few drops) that ceased within minutes; and 

recaptured

ertheless, toe clipping is not without problems, as evidenced by about 25 problem 

incidences (out of 248 toe-clipped individuals) due to natural causes or human error: 

(1) ten cases of naturally missing, deformed, or injured toes, which would make toe

clipping potentially confusing; (2) six or seven cases where field workers apparently 

read the wrong code on a previously toe-clipped animal; (3) three or four cases where 

the toe tip grew back, apparently due to not clipping far enough back; (4) and four

cases where two individuals were mistakenly given the same toe clip code. In 

addition, the number of unique toe-clip codes is finite (depending upon the number of

digits clipped), which poses problems for population monitoring in a fairly long-lived 

species (up to at least 6 years, with about 30 percent over-winter survival; Bailey

1939, Behrends personal communication, Chew and Butterworth 1964). 

The hypodermic insertion PIT-tagging procedure worked well in the field with

practice. Most mice seemed to experience minor inflammation, scabbing, and sensitivity to touch at the insertion wound. Tags appeared to migrate under the skin, which

exhibited bulging on some individuals. Unfortunately, recapture rates w

e were no over-winter recaptures of PIT-tagged individuals) to ascertain health 

effects of PIT tagging in the field. 

Population Genetics

Swei and others (2003) found that

bangsi, internationalis, and longimembris

subspecies is reciprocally monophyletic (i.e., no subspecies has a common ancestor 

not shared with any other subspecies). 

Pooling all extant populations of P. l. pacificus, the subspecies exhibits similar 

levels of nucleotide and haplotype diversity as other southern California subspecies 

of P. longimembris, although the sample from Dana Point has markedly low haplotype diversity in comparison to all o

tic variation among samples of P. l. pacificus, about 39.1 percent is variance 

among individuals within populations, 54.4 percent is between populations within a

time period, and 6.5 percent (which is not significantly different from zero) is between modern and historic samples. Thus, the pattern of haplotype differentiation 

suggests that historical populations (in the 1930s) were already largely isolated from 

one another, and that current patterns of genetic diversity are due largely to natural 

isolation of these populations prior to the modern urbanization of coastal California.

120 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

Indeed, the genetic signatures measured by Swei and others (2003) suggest that 

although populations of P. l. pacificus were somewhat isolated in recent evolutionary

time, they do not yet reflect a loss of genetic diversity due to habitat losses in modern 

times. Swei and others (2003) concluded that 40 to 70 years of habitat reduction at

locations like Dana Point are insufficient to override the population’s deeper genetic 

history, but that we can confidently predict future losses of genetic diversity unless

current population declines are reversed. 

Measures of gene flow between populations of P. l. pacificus are quite low, both 

in historical and current times. This stands in sharp contrast to higher levels of gene 

flow measured across greater geographic distances in the Mojave Desert for P. l. 

longimembris. Measures of gene flow between the Dana Point population and all

othe

sary to meet the recovery goal (USFWS 1998) of maintaining the full 

rema

espond quickly and positively to the shrub thinning by

redistributing themselves into the newly thinned habitat areas. During 1999 (prewere captured, yielding 12 independent grid

ptured on two grids)—9 observations (9

r P. longimembris populations are especially and uniformly low. The Dana Point 

population also has the lowest haplotype diversity of any sample of P. longimembris

(only 9 unique haplotypes recovered from 27 individuals). Together, these results 

suggest that the Dana Point population has been relatively small, as well as isolated, 

for some time. Slightly higher gene flow measures and haplotype diversity were recorded among the three sites on Camp Pendleton (San Mateo North and South and

Oscar One). 

Swei and others (2003) concluded that, given current levels of fragmentation

and low population sizes, genetic diversity will continue to erode within P. l. 

pacificus populations. This suggests that mixing of individuals from different sites 

may be neces

ining genetic variability of the subspecies. Although the haplotypes of extant P. 

l. pacificus populations are largely apportioned among populations, they are all 

genealogically closely related, and genetic variability seems to be governed by

neutral (rather than selective) genetic processes. It is possible that there are selective 

differences among these populations that are more likely to be expressed in nuclear

than mitochondrial genes, but this seems unlikely given the small geographic range 

of the subspecies and similarities in ecological conditions at occupied sites. Thus it 

appears that mixing individuals from different P. l. pacificus populations during

translocations should not have any deleterious genetic effects (for example,

“outbreeding depression”). 

Shrub Thinning Experiment

P. l. pacificus seemed to r

thinning), 11 individual P. l. pacificus

observations (one individual was ca

individuals) were on the known occupied control grids, 2 were on subgrids to be 

thinned (including 1 individual that also used a control grid), and 1 on a subgrid not

to be thinned. In contrast, during the year 2000 (post-thinning), 10 individual mice

were captured for a total of 18 unique observations—4 on control grids, 11 on

thinned subgrids, and 3 on unthinned subgrids. This shift in the distribution of

captures is statistically significant (P=0.026), largely due to the great increase in 

captures on the thinned subgrids. During the year 2000, 6 of 10 individuals were 

captured only on thinned subgrids (5 were captured on only one subgrid, but 1 male 

was captured using three different thinned subgrids). Of the 4 individuals captured on 

control grids, 3 also were captured on thinned subgrids. No individuals were captured 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 121 
Pacific Pocket Mouse Recovery Research—Spencer 

only on unthinned subgrids—all three individuals captured on unthinned subgrids

were also captured using the adjacent thinned subgrid. In fact, all captures on 

unthinned subgrids were less than 6 m (two trap distances) from the thinned area that

these individuals also used. Moreover, all capture locations on unthinned subgrids 

were connected to the thinned plots via natural inter-shrub openings. 

Thus, the pattern of captures before and after the thinning suggests a strong shift 

in distribution to the newly thinned habitat areas or their edges. This reinforces observations that P. l. pacificus, like other species of Perognathus, are most abundant in

fairly open habitats, although they may prefer to forage in microhabitats in or near

shru

n of population decline at Dana

Poin

l collaborative process followed by the PPM Studies Team—with 

frequent re-evaluation of priorities, methods, and results—has thus far proved effecthe most urgent questions bearing on species recovery. We benean active peer input and review process that helped keep the 

al to designing a P. l. pacificus translocation 

prog

At this

poin

bs (Brown and Lieberman 1973, Harris 1984, Meserve 1976, Price 1978,

Reichman and Price 1993). Although P. l. pacificus may forage under shrubs near

openings, they do not seem to frequent interior portions of closed-canopy shrub

cover, especially when other options are available. 

Unfortunately, intensive trapping at the Dana Point site during May-June 2001 

captured only four P. l. pacificus (2 female, 2 male; USFWS unpublished data). This 

sample is too small to reveal any patterns relative to the shrub-thinning experiment. 

Moreover, these results reinforce a general patter

t since the population’s discovery in 1993. Whether or not shrub thinning may 

positively influence population size at this or any other site is, therefore, not yet 

determined. 

Discussion 

The informa

tive in addressing

fited greatly from

program focused and scientifically valid while allowing for flexible, pragmatic 

responses to new information. We have proceeded with a realization that statistical

certainty, while desirable, isn’t always possible or necessary when prudent actions

must be taken to recover rare species. 

During 1998, the study team concluded that P. l. pacificus translocations were

necessary to achieve species recovery, but that existing information was insufficient 

to design a translocation program. We therefore refocused our research to answer

those questions considered most critic

ram. We now feel we have gained sufficient information to begin designing the 

program as one essential recovery tool. Please note that our goal is translocations to 

establish new P. l. pacificus populations, and thereby to help rescue the species from 

the brink of extinction. No one should interpret our intent as translocation to mitigate 

proposed take over of P. l. pacificus or their habitat by development projects. 

In the future, we hope to comprehensively update our evaluation of translocation 

feasibility, design a framework translocation program based on the updated information, and design experiments to test translocation methods (for example, use of

release enclosures and artificial burrows) prior to implementing the program.

t in the process, I am confidant that continued research and common sense could 

sufficiently answer the most pressing biological questions bearing on species

recovery. Less certain, however, are methods of overcoming political and economic

obstacles. For example, subsequent to the drafting of this document, the U.S. Marine 

122 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Pocket Mouse Recovery Research—Spencer 

Corps denied requests from the PPM Studies Team to continue P. l. pacificus studies 

on Camp Pendleton, due apparently to concerns that the results might constrain 

military training and readiness. This has postponed indefinitely our ability to answer 

remaining questions and further species recovery. Such non-scientific obstacles must 

now take center stage if we are truly to recover populations of this charismatic little 

mouse. 

Acknowledgments 

Although I prepared this document and am responsible for any inaccuracies it 

contains, numerous individuals contributed to these studies in a uniquely collaboraon Corridor Agencies (TCA) provided much of the 

y Laura Coley Eisenberg, and later by Ann Johnsy lives of two little pocket mice. Journal of Mammalogy 20:

325-328. 

940. Trapping, handling and marking fox squirrels. Journal of 

ement 4: 444-450. 

dents in sand dune habitats. Ecology 54: 788-797. 

tive effort. The Transportati

funding, with guidance provided b

ton and Macie Cleary-Milan. Christina Schaefer served as project manager for the

TCA-funded projects and added valuable review and comment throughout. The California Department of Fish and Game, U.S. Fish and Wildlife Service, and Marine 

Corps Base Camp Pendleton provided additional funding. Shana Dodd coordinated

field studies for the Snow Creek surrogate study, the translocation receiver site reconnaissance study, and the Dana Point shrub-thinning study. She also collected P. l.

longimembris for the laboratory surrogate study. Stephen J. Montgomery, Rebecca 

Wacklor, and Hal Holland assisted with field data collection. Steve Montgomery also 

housed the laboratory population, which was maintained and monitored by Kris

Beaver, Hal Holland, and Michael Wellik. Dr. James Patton and Andrea Swei performed the genetics analyses. Dr. Robert Graham provided expertise in soil science 

and assisted in ranking soils for the habitat evaluation model. Dr. Anthony Michael, a

veterinarian, volunteered his expertise, time, and materials to assist with the laboratory marking experiment. Wendy Barto and Debbie Turner performed the GIS work

with direction from Pat Atchison. GIS data were used with the permission of the San

Diego Association of Governments and County of Orange, with special thanks to 

Gary Meideros. Comments from and discussions with the PPM scientific peer committee (Drs. James Patton, William Longland, Jay Diffendorfer, Howard Shellhammer, Nicholas Holler, Michael Wooten, and Richard Reading) have been invaluable

in guiding our studies. Mark Pavelka, Will Miller, Clark Winchell, and Loren Hays

of the U.S. Fish and Wildlife Service provided data, helpful discussion, and advice on

all phases of this work. David Lawhead, William Tippets, and John Gustafson of the

California Department of Fish and Game also provided guidance and monetary

support for the fieldwork and genetic analyses. William Berry and Deborah Bieber

provided coordination with Marine Corps Base Camp Pendleton. Comments by

editor Barbara Kus and reviewers Lee McClenaghan and James Patton greatly

improved this manuscript. 

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USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 125 

Herbaceous Responses to Livestock 

Grazing in Californian Oak Woodlands: 

A Review for Habitat Improvement and 

Conservation Potential1

 

Barbara Allen-Diaz and Randall D. Jackson2

Abstract 

Livestock grazing effects on herbaceous vegetation diversity and ecosystem function are 

summarized from peer-reviewed literature. Ecosystem responses are linked to grazing 

management via residual dry matter (RDM), which is alternatively called litter or mulch.

Residual dry matter is the dead herbaceous aboveground biomass remaining on a site at the

time of autumn germination in annual grassland and oak woodlands. RDM is known to play

an important role in nutrient cycling, soil erosion prevention, and herbaceous species

production and composition the following year. Minimum RDM guidelines are used to ensure

forage sustainability and protect ecosystem resources. Recent work has indicated that

maximum RDM standards may be necessary to meet wildlife habitat requirements at both site

(for example, burrowing owls) and landscape (for example, kit fox) scales and may be useful 

for fire-fuels management. Livestock grazing can be a useful conservation tool because 

managers can control the type, abundance, and spatial and temporal distribution of grazers to

meet desired goals. Furthermore, political and economic considerations are rendering 

prescribed fire, hand mowing, and large native herbivore reintroduction increasingly difficult

tools to apply. We review herbaceous vegetation responses to grazing management so that

they might be translated to advantages and disadvantages for specific ecosystem and habitat

conservation goals on a case-by-case basis. 

Key words: annual grassland, biodiversity, conservation, grazing, habitat improvement,

livestock, oak woodlands 

Introduction 

Oak woodlands and savannas occupy 3 million ha in California (Bolsinger 1988,

Griffin 1977). Seven series and 57 subseries have been described within the oak

community type (Allen and others 1991). Oak types occupy 52 of California’s 58

counties and are widely distributed west of the Sierra Nevada (fig. 1). Oak savannas 

and woodlands generally occur from 60 to 700 masl (meters above sea level),

between annual grassland at lower elevations and mixed conifer forests at higher 

elevations (Barbour and Major 1988). The climate is Mediterranean, with 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 Professor and Visiting Assistant Researcher, Ecosystem Sciences, University of California, Berkeley, 

CA 94720-3110 (email: ballen@nature.berkeley.edu) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 127 
Oak Woodland Grazing—Allen-Diaz and Jackson 

Valley oak (Quercus lobata) Black oak (Q. kelloggii) Coast live oak (Q. agrifolia) Interior live oak (Q. wislizenii)

Blue oak (Q. douglasii)

Engelmann oak (Q. engelmanii)

Oregon oak (Q. garryana) Canyon live oak (Q. chrysolepis)

Figure 1—Distribution of eight native oak trees in California (modified from Pavlik

and others 1991). 

precipitation occurring primarily as rain between October and May. Summer drought

results in 3 to 5 months of water deficit (Major 1977). 

Dominant trees in the oak woodland include blue oak (Quercus douglasii), 

valley oak (Q. lobata), interior live oak (Q. wislizenii), coast live oak (Q. agrifolia), 

Oregon white oak (Q. garryana), and Engelmann oak (Q. engelmannii). These 

species occur in monospecific to mixed stands (Allen-Diaz and others 1999). Ninety

percent of California’s oak woodlands are privately owned (Ewing and others 1988).

Over 300 species of vertebrates live in oak habitats, including ~80 mammals, 160 

birds, 25 amphibians, and 40 reptiles (Standiford and Tinnin 1996). Livestock 

grazing is the primary agricultural use of oak woodlands, but other economic 

products include firewood, wildlife, water, and recreation (McClaran and Bartolome 

1985, Standiford and Tinnin 1996). 

Annual grassland is the major oak understory vegetation, although shrubs and

perennial grasses may be an important component in some areas. Native perennial 

grasses are scattered throughout California (Bartolome and Gemmill 1981, Beetle

1947) and found interspersed within a matrix of annual grasses, forbs, and legumes 

that are mostly native to the Mediterranean region (Jackson 1985). 

Here we review the scientific basis for livestock grazing management on the

herbaceous component of California’s oak woodlands. The status of this ecosystem 

over the last 25 years has been addressed in a series of symposia (Pillsbury and others 

1997, Plumb 1979, Plumb and Pillsbury 1986, Standiford 1990, Standiford and others 

2002), the proceedings of which contain numerous studies of grazing and other

management effects on the regeneration of oaks, in addition to many other topics. We 

do not attempt to review or duplicate this literature. Instead, we emphasize the 

herbaceous plant communities of oak woodlands and their responses to management

so that they might be translated to advantages and disadvantages for specific 

ecosystem and habitat conservation goals on a case-by-case basis. As much as

possible we rely on literature specific to oak woodlands, but with their annual 

128 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Oak Woodland Grazing—Allen-Diaz and Jackson 

grassland understory, some references to open grasslands are applicable and cited. 

Finally, we discuss two approaches for managing and understanding complex

systems—Residual Dry Matter guidelines and State-Transition models—that have

gained wide use in the field of rangeland resource science as rational approaches to 

ecosystem management. 

Pre-European-settlement Herbaceous Flora 

Dominant herbaceous vegetation in California prior to European contact is unknown. 

Many believe native perennial grasses, particularly the bunchgrass Nassella pulchra, 

once enjoyed a more widespread distribution (Clements 1934, Heady 1977). Hamilton (1998) has argued rather convincingly against overuse of this paradigm, citing 

over-extrapolation of Clements’s climax community concept and the dogma that has

derived from it. He suggested that native annuals were once dominant, especially in

drier parts of the grassland. Holstein (2001) suggested the rhizomatous perennial

grass Leymus triticoides dominated the pre-agricultural Central Valley floor. 

However, his analysis relied on the relict method for which he and others criticize 

Clements, along with observational accounts of European explorers and settlers. 

We know of only two studies providing physical evidence of pre-European 

settlement composition. Bartolome and others (1986) found greater abundance of 

distinctive opal phytoliths, silica bodies that are resistant to decay with shapes 

specific to certain taxonomic groups, at soil depths corresponding to greater than 150

years past. The shapes these phytoliths took were specific to those found in perennial

grasses, indicating their greater abundance in the past at that particular site (Jepson 

Prairie). 

The second paper offering some physical evidence of pre-European flora is 

Mensing and Byrne (1998). They examined preserved pollen in sediment cores from 

the Santa Barbara channel and determined that the presence of the exotic annual

Erodium cicutarium—now ubiquitous in much of California—pre-dated European

settlement and livestock introduction. They show patterns suggesting it invaded from 

Baja California prior to the Mission Period in California’s history.

Controls on Herbaceous Community Structure 

Factors at many spatial and temporal scales interact to control herbaceous

productivity in California’s annual grasslands, although these factors may be 

modulated by the presence of oak canopy. In annual grasslands, peak standing crop 

(aboveground biomass at late spring seed set) varies interannually as a function of the 

timing and amount of precipitation, temperature (Bentley and Talbot 1948, George 

and others 1988, Heady 1958, Talbot and others 1939), and edaphic and topographic 

characteristics (Jackson and others 1990). A typical yearly production curve for the 

annual grassland understory (fig. 2) includes the onset of autumn germination with 

the first rains over 2.5 cm occurring within a 1-week period. Slow winter growth 

progresses as temperatures decline, followed by rapid spring growth as soil 

temperatures increase concurrent with adequate soil moisture. Peak standing crop of 

the herbaceous vegetation generally occurs between 1 April and 15 May, followed by

the death of the annual plants. Standing dead biomass slowly decomposes as summer 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 129 
Oak Woodland Grazing—Allen-Diaz and Jackson 

drought slows microbial activity, until the ensuing autumn rains stimulate 

decomposition concurrent with annual plant germination (Jackson and others 1988). 

50

100

150

200

250

Autumn Winter Spring Summer

Season

Bio

mass (

g·m-2)

Plant death

Live

Dead

Figure 2—Typical herbaceous production for the University of California’s Sierra

Foothill Research & Extension Center (SFREC), a blue oak woodland site NE of 

Sacramento. 

Herbaceous understory production is primarily controlled by interannual 

weather variability (Connor and Willoughby 1997, Frost and others 1991); however, 

several workers have demonstrated that tree canopy cover exerts a proximate but 

variable influence on peak standing biomass depending on regional location, tree 

density, and tree type (Bartolome and others 1994, Frost and McDougald 1989,

Ratliff and others 1991). Relative to open grasslands, canopy cover inhibits 

herbaceous production in areas of California receiving >50 cm annual precipitation 

(McClaran and Bartolome 1989), suggesting light reduction as a mechanism. The 

inverse relationship generally holds for drier portions of the state, where canopy

cover serves to attenuate drought stress. From a livestock production perspective,

Frost and others (1997) found that an increase in tissue and litter quality per unit

biomass under oak canopy in drier regions more than compensated for reduced 

herbage mass. Nutrient concentrations differed as a result of species composition 

differences rather than some change in individual plant nutrient use efficiencies. 

Species composition differences between oak understory and open grassland 

have been demonstrated by several authors (Borchert and others 1991, Jackson and

others 1990, Marañón and Bartolome 1993, Saenz and Sawyer 1986). Some species 

appear to be strongly controlled by this dichotomy, but generalizations for 

California’s oak woodlands are tenuous. A given species may be strongly associated

with canopy cover at one site but with open grassland at another. For instance,

Nassella pulchra is thought to be an open grassland species, but we have observed it 

scattered under relatively high oak canopy cover in the Sierran foothills (B. AllenDiaz, unpublished data). Alternatively, Cynosurus echinatus is rare in open grassland 

but common beneath oak canopy throughout the state. Rice and Nagy (2000) sought 

the mechanism for the spatial separation of Bromus diandrus (found under canopy)

and Bromus hordeaceus (found in the open), showing that interspecific competition 

was important in the high resource oak canopy soils, but only B. hordeaceus could 

tolerate the harsher conditions of open grassland; there was little evidence of 

competition between these species in the open. Working in blue oak woodlands of 

the Sierran foothills, Shlisky (2001) found that aspect, measured at the 10- to 100-m2

scale, was a more important determinant of species composition than canopy cover.

130 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Oak Woodland Grazing—Allen-Diaz and Jackson 

Hence, species composition in the herbaceous understory is a complicated mix of 

time, site, abiotic, and biotic interactions; generalizations are tenuous. 

Plant species diversity (measured as species richness) was shown to be highest

at intermediate herbaceous biomass levels (Heady and others 1992, Marañón and

Bartolome 1994) following the model of Grime (1979) and discussed by Marañón 

and Garcia (1997) and Garcia and others (1993). Marañón and Bartolome (1994)

listed optimum biomass levels for maximum species diversity on California annual 

grassland/oak woodlands at 35 to 57 g m-2. They also showed that maintaining an oak 

overstory component provides for maximal landscape diversity owing to different 

plant assemblages in canopy and the open. 

Grazing Effects on Herbaceous Vegetation 

Grazing animals have been part of California grassland, savanna, and woodland 

ecosystems for thousands of years (Edwards 1992). Before European contact and the 

establishment of widespread cattle and sheep grazing (Burcham 1957), large herds of 

pronghorn antelope, tule elk, mule deer, and rodents grazed California grasslands, 

savannas, and wetlands (Edwards 1996). Many grassland species have evolved habits 

(for example, prostrate growth) and structures (for example, basal meristems) to 

avoid or tolerate aboveground tissue loss (Briske 1991). 

Grazing animals may influence plant composition and community structure both 

directly via physical means and indirectly via biogeochemical and biotic feedbacks.

McBride and Heady (1968), for example, found that trampling by cattle prevented 

the encroachment of native coyote bush (Baccharis pilularis) into the grasslands of 

the Berkeley hills. Some evidence exists that high grazing levels promote invasion of

undesirable plant species via biological feedbacks (D'Antonio and others 1999); that 

is, grazers alter the competitive balance of natural communities by interrupting

nutrient cycles, altering soil organic matter quality, and creating gaps for 

colonization, in effect creating an alternate ecosystem more conducive to the invader.

However, this model assumes equilibrium-type conditions where biotic interactions 

play a significant role in shaping the plant community (Wiens 1984). California’s

Mediterranean climate, with its high interannual rainfall variability, does not lend 

itself to equilibrium conditions where competitive interactions among plants are 

allowed sufficient time to affect community structure. From one year to the next, 

community structure is primarily determined by abiotic constraints associated with 

autumn germination (Bartolome 1979, Evans and Young 1989). 

Evidence for the well-known Intermediate Disturbance Hypothesis (IDH) 

(Connell 1978, Sousa 1979) is inferred from Marañón and Bartolome (1994) and 

Bartolome and others (1980), who showed diversity maximized at intermediate levels 

of aboveground biomass. If intermediate levels of grazing disturbance leave intermediate levels of biomass, these results should apply. However, these results were from 

coast live oak (non-deciduous) woodlands on the Coast Range (Marañón and

Bartolome 1994) and from various open grassland sites (Bartolome and others 1980). 

The scientific basis for IDH as a mechanism in deciduous blue oak woodlands is 

lacking. Fehmi and Bartolome (2002) showed that a possible tradeoff between

livestock and rodent herbivory exists in these grasslands. In their study, rodents 

appeared to preferentially locate burrows (thereby disturbing vegetation) in areas 

where livestock were excluded, that is, high cover sites. Hence, any response of 

diversity to livestock grazing regime changes may be cancelled by rodent activity.

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 131 
Oak Woodland Grazing—Allen-Diaz and Jackson 

Similar results were noted by Dyer and Rice (1997). Keeley (2002) found slightly

reduced herbaceous species richness in livestock-excluded areas compared to grazed

areas of blue oak woodlands of the southern Sierran foothills. Nonetheless, these 

zones maintained greater plant diversity than adjacent mixed-conifer zones at higher 

elevations irrespective of grazing management, but the oak woodlands were more 

highly invaded. 

Spring-fed wetlands and riparian areas are often the only sources of water in oak

woodlands and are especially heavily utilized by grazing animals. However, light to

moderate, autumn/winter grazing had little effect on Sierra Nevada foothill spring-fed 

vegetation after 6 years of treatment (Allen-Diaz and Jackson 2000). Continued 

monitoring of these systems under experimental treatments has shown that by years 7 

through 10, moderate grazing reduced herbaceous cover, light grazing had minimal 

effect, and grazing removal significantly increased cover (Jackson 2002). These increases in cover resulted in an undesirable accumulation of plant litter that suppressed subsequent plant productivity. Studies examining grazing effects on 

vegetation structure in riparian systems other than spring-fed wetlands are few. High

intensity grazing can negatively affect water quality, plant biodiversity, productivity,

wildlife habitat, wildlife species biodiversity, and nutrient cycling in riparian areas in

regions with continental-type climates (Belsky and others 1999, Clary 1995, 

Fleischner 1994, Kauffman and Krueger 1984, Kauffman and others 1983).

However, extrapolation of these results to Mediterranean-type regions should be

made very cautiously (Gasith and Resh 1999, Larsen and others 1998). Effects of

moderate to light grazing on ecosystems of these regions tend to be overwhelmed by 

larger scale environmental fluctuations (Allen-Diaz and Jackson 2000, Allen-Diaz 

and others 1998). 

Many studies have sought to determine management techniques for enhancing 

native perennial grasses, namely grazing, burning, and grazing removal. If target

perennial grasses are not present on a site, these treatments will not be effective.

However, if present on a site, native perennial grasses have been shown to respond 

favorably to disturbance treatments, but the particular combination of treatments 

depends on geography and possibly weather patterns. Bartolome and Gemmill (1981) 

rejected the notion that N. pulchra represents relictual dominance and hypothesized

that it likely is a disturbance-adapted species that finds refuge in spaces where light is 

less limiting than belowground resources. Their hypotheses were largely confirmed

by experimental findings of Dyer and Rice (1999). Dyer (2002) found that N. pulchra

seed from grazed and/or burned individuals germinated and survived at higher rates 

than seed from undisturbed plants. He invoked maternal provisioning as the

mechanism by which this advantage is conferred. 

Working in Coast Range grassland, Bartolome and others (unpublished data) 

found a positive response of Nassella under post-treatment grazing removal, but not 

during the grazing removal treatment, indicating some residual effect of their 

seasonal grazing treatments that was magnified by removal of the disturbance. This

aligned with Dyer’s and Rice’s (1997) work in Valley grassland showing that N. 

pulchra abundance increased only when released from diffuse competition with 

neighboring annual grasses and forbs. 

The studies cited above illustrate the importance of rangeland management

based upon scientifically produced information that is applicable to appropriate sites 

and scales (Bartolome 1989). Range managers control grazing animal type, 

abundance, spatial distribution, and temporal distribution (season of use). Each of

132 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Oak Woodland Grazing—Allen-Diaz and Jackson 

these singular or interacting factors can produce disparate effects on various 

ecosystems and ecosystem components. Therefore, grazing efficacy should not be 

assessed as an “either-or” proposition (in the sense used by Belsky and others 1999). 

Grazing Effects at the Ecosystem Level 

Grazing can affect ecosystem processes as well as ecosystem structure, though this

topic has received considerably less attention in oak woodland understory than open 

annual grasslands. In general, grazing accelerates carbon and nutrient cycling by

effectively bypassing the microbial decomposition pathway. Livestock mineralize 

plant organic material and return it to the soil and atmosphere as feces, urine, and gas 

much more quickly than microbes. In perennial grasslands of the Midwest,

accelerated nutrient cycling is credited for stimulating net primary productivity

(Frank and Evans 1997, Frank and McNaughton 1993, Frank and others 1994). 

However, similar grazing effects on nutrient dynamics in California annual grassland 

were not evident (Dahlgren and others 1997, Davidson and others 1993). Nitrogen

quickly cycles within annual-dominated ecosystems, where plant species possess low 

nutrient use efficiencies and high litter qualities irrespective of herbivory (Davidson 

and others 1990, Jackson and others 1989, Schimel and others 1989). 

Dahlgren (1997) describes soils beneath oak canopy as “islands of fertility”

because of greater carbon, nitrogen, and phosphorous stocks compared to adjacent

open grasslands sites. The patchiness of oak woodland canopy may be enhanced by 

the ability of oaks to garner water and nutrients from beyond the canopy perimeter, 

from the open grassland spaces between them and their neighbors, and then 

preferentially returning leaf litter below the existing canopy, thereby redistributing

ecosystem resources. However, this process would constitute a positive feedback that 

would be unsustainable in the long-term, depressing herbaceous production in the 

open; this effect is not evident. An untested hypothesis is that herbivores provide a 

check on this effect by harvesting herbaceous resources from beneath the canopy and 

redistributing them relatively more homogeneously across the landscape. 

Camping and others (2002) found little effect of grazing on soil chemical 

properties in blue oak woodland of northern Sierran foothills. Dudley and others 

(2002), working at the San Joaquin Experimental Range, found higher soil bulk

densities in areas with moderate to heavy grazing intensities compared to historically 

ungrazed areas, indicating deleterious effects on soil physical properties. 

Spring-fed wetlands provide many important functions within oak woodland 

landscapes; livestock grazing modifies these functions. Much of the water exiting oak 

woodland watersheds passes through these highly productive zones at the terrestrialaquatic interface (Huang 1997). Wetland vegetation in these zones, typically cattails 

(Typha spp.), sedges, rushes, and perennial grasses, likely acts as a nutrient filter 

(primarily nitrate) as waters emerge at the soil surface. Jackson (2002) has 

determined with a paired-plot grazing removal experiment that nitrate concentrations

in surface waters where grazing was discontinued for only 2 years were as much as

five times greater than grazed counterparts. Furthermore, ungrazed concentrations far

exceeded the U.S. Environmental Protection Agency’s maximum standard for surface

waters of 714 μm (Fan and others 1987).

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 133 
Oak Woodland Grazing—Allen-Diaz and Jackson 

that while some species may respond to RDM manipulations in some years, overall 

plant community composition is relatively insensitive to these changes. Composition 

seems to be entrained by annual weather patterns producing nonequilibrium 

conditions that render overall community manipulation via livestock grazing futile. 

However, this is not to say that a single species such as a native perennial grass or

diversity (in the sense of Meyer and Schiffman 1999) cannot be managed for by

manipulating the type, intensity, or distribution of grazing (Hatch and others 1999). 

Table 1—Minimum residual dry matter guidelines (g·m-2) for livestock grazing management

in Californian annual grassland. 

Slope class

Precipitation regime 

(yearly total in cm) Lower or flat Average or gentle Upper or steep 

< 25 22 28 39

25-100 45 67 90

>100 84 112 140

RDM is now the preferred criterion for measuring grazing management 

effectiveness on California’s rangelands. Livestock operators estimate RDM using 

various methods including harvesting a known area, dry weight rank, and visual 

estimation (Harris and others 2002). Desired RDM levels are determined based on

combinations of slope and average total rainfall (table 1). Experiments are currently 

underway to produce RDM guidelines for oak woodlands where overstory litter

complicates the matter. The University of California Cooperative Extension 

publishes scorecards indicating minimum RDM levels that should sustain system

productivity in the face of fluctuating weather conditions. Guidelines like these 

should be developed for various conservation management objectives. 

State-transition Models for Adaptive Management 

The emphasis throughout this paper has been on uncertainty and hedged statements 

about environmental and management effects on grassland understory in oak 

woodlands. Unraveling of the complexity inherent in this system is occurring, albeit 

slowly. This is attributable to the plodding pace of the scientific method combined 

with the context-dependent nature of responses in Mediterranean ecosystems (Gasith

and Resh 1999). Experimental information is needed from applied treatment 

combinations under the range of climate patterns for each setting. 

Complex phenomena such as these beg for an organizing framework where results are catalogued and testable hypotheses generated. One such organizing tool, 

state-transition models, has been espoused for over a decade by rangeland ecologists 

working in the inherently variable and difficult-to-manage arid and semi-arid regions

of Australia and North America (Brown 1994, Filet 1994, Taylor and others 1994,

Westoby and others 1989, Whalley 1994). Applications vary, but all state-transition 

models are simple box-and-arrow diagrams of actual ecosystem states (defined by

combinations of vegetation, soils, fauna, and so forth) and transitions among states 

related to all possible factors—anthropogenic or otherwise (fig. 4). Observed and/or 

hypothetical transitions among states are catalogued, their causes to be tested and

 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 135 
Oak Woodland Grazing—Allen-Diaz and Jackson 

V

Ceanothus,

manazanita, and

poison oak

dominant

VI

Oak trees,

saplings, and

seelings with 

annual grassland

understory

VII

Oak trees,

saplings, and 

seedlings with

Purple

needlegrass

understory

IV

Purple

needlegrass 

dominant

T9

T3

T1

T11 T12

I

Medusahead

dominant

II

Filaree 

dominant

III

Soft chess,

ripgut brome,

and wild oats

dominant

T7

T8

T10

T4

T6

T2 T5

Catalog of states

State I. Medusahead forms nearly pure stands with heavy

litter that effectively excludes most other annual grasses

and forbs. High silica content is believed to slow

decomposition, ...

State II. Filaree dominates seedbank and the

aboveground standing crop. Medusahead density is

substantially reduced...

...

State VII. Blue oak and interior live oak savanna with a 

perennial grass understory.

Catalog of transitions

T1. Medusahead seedbank reduced by 50 to 90%. This

can be effected by several forms of disturbance...

T2. Filaree dominance is reduced as other species,

especially annual grasses, successfully invade and

colonize the patch...

...

T12. Yearlong continuous grazing and drought reduce

needlegrass vigor and survival.

Figure 4—Example of a state-transition model for a Sierran foothill blue oak 

woodland site. Models such as these could be developed for ecosystem conservation 

as a means of organizing information and generating testable hypotheses about 

management induced effects. Note that this model, adapted from George and others 

1992, has been simplified; it is shown here only for demonstration of state-transition 

model structure.

validated or rejected. State-transition models have since been applied to California’s 

oak woodlands (Huntsinger and Bartolome 1992) and annual grasslands in various 

forms (George and others 1992, Jackson and Bartolome 2002). However, these 

applications usually summarize data or hypotheses under relatively restrictive 

conditions; they seldom allow for filling in gaps in information. More useful models 

would exhaustively identify as many different weather scenarios and management

goals as possible, explicitly differentiating reliable knowledge from hypotheses and 

conjecture. Publication of this information would minimize redundancy and allow for 

refinement of research questions. 

State-transition models are especially useful where system response is complex 

and dynamic and therefore should gain wide use in conservation management. They

are most useful when data-driven by spatially and temporally explicit results acquired 

through monitoring and adaptive management. 

Livestock Grazing Alternatives 

Alternatives to livestock grazing for vegetation management include non-commercial 

grazing, prescribed fire, herbicidal application, and hand weeding. All are 

ecologically effective at some scale, yet all possess significant drawbacks. Fire has

great potential as a vegetation management tool as it is generally more effective than

grazing at removing aboveground biomass. Successful applications include control of 

yellow starthistle (DiTomaso and others 1999), barbed goatgrass (DiTomaso and

others 2001, Hopkinson and others 1999), and medusahead (Murphy and Lusk 1961).

However, it is ultimately cost-prohibitive to many managers and agencies because of 

the high costs associated with maintaining emergency personnel on site during “burn 

136 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Oak Woodland Grazing—Allen-Diaz and Jackson 

windows.” Burn events are prone to postponement and cancellation as ideal 

environmental conditions are sought to maximize the burn’s efficacy while 

minimizing air pollution. Finally, catastrophe may result from the inevitable escaped

prescribed burn. 

Potential side effects of herbicide treatment render it a politically as well as an 

ecologically unsound alternative over extensive areas. Furthermore, long-term

efficacy of such treatments tends to be poor (Allen-Diaz and Bartolome 1998).

Combinations of herbicidal spraying and livestock grazing are currently being tested

for efficacy of yellow starthistle control in northern California. Finally, hand-weeding and non-commercial grazing by waterfowl, goats, and the like for conservation 

management are effective, politically feasible measures, but at local scales only.

Regional, landscape, and even site scales are not manageable by these means. 

Future Research 

We do not propose livestock grazing as a panacea—a vegetation management tool

whose application will help meet every conservation goal. When managed 

improperly, livestock cause damage to riparian and upland systems and alter habitat,

affecting other species. On the other hand, there is a renewed interest in using 

livestock to achieve specific conservation objectives (Meyer and Schiffman 1999,

Pykala 2000, Stolzenburg 2000). This presumes that researchers and managers 

understand enough about the structure and function of the ecosystem to implement 

grazing management that will effect the desired response. 

Future studies must continue elucidating the relationships between a wide range 

of grazing intensities and ecosystem responses. Experiments must continue assessing

grazing gradients from none to light to moderate to heavy. Furthermore, research 

should focus on uncovering mechanisms controlling system dynamics in the hope

that general principles might evolve. Overgrazing, defined as grazing intensity which

will not sustain the desired state of an ecosystem, should not be compared to grazing 

removal; information from such studies does not advance our understanding but

serves to further polarize complex ecological, political, and management issues. 

Managing grazing for conservation objectives will require more than the simple

RDM scorecards for simultaneously maximizing herbage removal and regrowth used 

by livestock producers now. Optimal timing and distribution of grazing as well as the 

type of grazer and grazing intensity will differ for each site and/or species of interest, 

necessitating monitoring and adaptive management until ecosystem responses under

controlled situations can be assessed. Resource consultants are beginning to use 

remotely sensed RDM estimates in grazing management plans (Standiford and others 

1999). Linking geographic information systems to direct management is necessary so 

that these tools can enjoy wider application. Monitoring efforts should be aimed at

multiple spatial and temporal scales for state-transition model development that will

lead toward heuristic and even serendipitous knowledge and away from the dogma 

that surrounds much of rangeland ecology today.

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Management Tools 

Residual Dry Matter (RDM) Manipulation 

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Figure 3—Herbaceous production response to residual dry matter manipulations at 

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species composition component of the Bartolome and others (1980) data set indicates 

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144 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Impacts in the Mixed Conifer 

Forests of Southern California1

Patrick J. Temple, Andrzej Bytnerowicz, Mark E. Fenn, and 

Mark A. Poth2

Abstract 

Air pollution, principally in the form of photochemical ozone and deposition of nitrogen

compounds, has significantly affected mixed conifer forests in the mountains of southern

California. Foliar injury, premature needle abscission, crown thinning, and reduced growth 

and vigor have been well documented, particularly for ponderosa (Pinus ponderosa Laws.) 

and Jeffrey (P. jeffreyi Grev. and Balf.) pines on the western side of the pollution deposition 

gradient in the San Bernardino Mountains. Tree mortality of the more ozone-susceptible 

ponderosa and Jeffrey pines has led to alterations in stand composition, in favor of increased 

dominance by more ozone-resistant species such as incense cedar (Calocedrus decurrens

[Torr.] Florin), white fir (Abies concolor [Gord. & Glend.] Lindl.), and sugar pine (P. 

lambertiana Doug.). Increased rates of litter deposition, alterations in C/N ratios in litter and 

soil, and reductions in fine root biomass of trees have also altered the dynamics of

biogeochemical processing in stands impacted by ozone and excess N deposition. Research 

into the effects of atmospheric deposition across the mixed conifer forests of the San

Bernardino Mountains continues to provide insights into the complex interactions among

anthropogenic and natural stresses in a forest ecosystem. 

Key words: air pollution, forest health, nitrogen deposition, ozone, San Bernardino National 

Forest, water quality 

Introduction 

For the past 40+ years, mixed conifer forests in the mountains of southern California 

have been exposed to the highest levels of oxidant air pollution and the highest levels 

of nitrogen deposition of any ecosystem in North America (Miller and McBride 

1999, Takemoto and others 2000). These high concentrations of air pollutants arise as 

a consequence of a unique set of topographic, climatological, and social factors 

characteristic of the Los Angeles (LA) air basin. First, the Transverse Ranges,

including the Santa Monica, San Gabriel, and San Bernardino Mountains to the north, 

and the Peninsular Ranges, including the San Jacinto and Santa Rosa Mountains to

the east, are barriers to the westward flow of oceanic breezes, particularly during the

summer months. The stagnant air masses and strong radiational heating contribute to 

the formation of a subsidence inversion layer, typically beginning at about 500 m 

(Edinger 1973). The inversion layer limits vertical mixing, which traps air pollutants

at its base, leading to increased pollutant concentrations near ground level. Second, 

increased industrialization of the region during and after World War II, rapid 

population growth, and increased reliance on the automobile as the basic and

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 Atmospheric Deposition Program, Pacific Southwest Research Station, USDA Forest Service, 4955 

Canyon Crest Dr., Riverside, CA 92507 (email: ptemple@fs.fed.us) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 145 
Air Pollution Effects on Conifer Forests—Temple and others 

necessary form of transportation across the region contributed into the atmosphere

massive amounts of precursor pollutants in the form of hydrocarbons from gasoline 

and oxides of nitrogen from auto exhaust. Photochemical processes converted the 

precursor molecules to ozone (O3), peroxyacyl nitrates (PANs), aldehydes, nitric 

acid, and other highly oxidative compounds (Finlayson-Pitts and Pitts 1986). These 

photochemical air pollutants, principally O3, have had significant adverse effects on

the vegetation of the region, from agricultural crops near the coast to the mixed 

conifer forests of the San Bernardino Mountains. 

Historical Trends of Air Quality in the San Bernardino 

Mountains 

Tropospheric O3 forms by the action of ultraviolet radiation (UV) on oxides of

nitrogen, principally NO2. The photolysis of NO2 produces NO and singlet O, which

quickly reacts with molecular O2 to produce O3. This reaction is rate-limited by the

decomposition of O3 and its reaction with NO, so that in the absence of other air

pollutants, the concentration of O3 produced by the photolysis of NO2 is relatively

low (Finlayson-Pitts and Pitts 1986). However, in the presence of reactive 

hydrocarbons, such as those from gasoline and auto exhaust, a series of complex 

catalytic reactions reduces the concentration of NO, leaving very little NO to react

with O3 and permitting the concentration of O3 to greatly increase. The amount of 

surface ozone formed by these reactions depends upon time of day, amounts of 

precursor pollutants, and meteorological conditions. In the morning, large amounts of

precursors are produced by morning rush hour traffic, but the concentration of O3 is 

low. Ozone concentrations increase during the day as photochemical reactions

continue. Concentrations also increase along a gradient from west to east as onshore 

breezes transport the polluted air mass eastward. In addition, the height of the 

inversion layer also increases from west to east as elevations increase from the

coastal plain to the mountains and as temperatures increase from the coast to the 

interior. As a result, the highest O3 concentrations in the LA Basin are typically found

not in the valleys, but in mountain slopes facing LA at 1200 to 1500 m, near the base 

of the inversion layer. Afternoon heating of the mountain slopes also vents the highly

polluted layers up and over the mountain ridgeline by upslope flow (Edinger 1973).

The inversion layer may also serve as a reservoir for ozone in the evening. The lack 

of vertical mixing as the air mass cools and stabilizes reduces contact of ozone with 

scavenging processes such as deposition to ground surfaces and reaction with NO 

molecules. As a consequence, high concentrations of O3 persist in the evening and 

early morning hours in forested areas of the mountains, while O3 concentrations in 

the valleys typically decrease rapidly as the sun sets (Edinger and others 1972). Thus 

the mixed conifer forests on the western slopes of the San Bernardino Mountains 

(SBM) have been subjected not only to high concentrations of O3 during mid to late 

afternoon but in the early morning as well, when humidity is relatively high and

physiological activity of the trees is also at its peak. 

Monitoring oxidant concentrations in the SBM began in 1968 with the

placement of a Mast ozone meter at Rim Forest/Sky Forest (NRC 1977). Data from 

this instrument and correlated data from a monitoring station in San Bernardino

indicated a continuing upward trend from 1963 in both peak ozone concentrations 

and in numbers of hours of elevated O3 in the mountains. These trends culminated in

1974, when maximum hourly O3 concentrations at Sky Forest exceeded 0.33 parts 

146 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Effects on Conifer Forests—Temple and others 

per million (ppm) on 46 days. The highest hourly concentrations averaged 0.38 ppm, 

and the highest hourly concentration ever recorded was 0.60 ppm on June 28, 1974

(NRC 1977). Peak and mean O3 concentrations in the mountains remained relatively

constant throughout the 1970s, with growing season (May to September) means 

frequently exceeding 0.10 ppm for 10 or more hours daily (Miller and McBride 

1975). However, in response to concerted air pollution control activities, O3 levels 

have declined steadily since the 1980s (Miller and others 1997). Growing season 

ozone means in the late 1990s have declined to 0.05 to 0.06 ppm, and recent peak 

concentrations have rarely exceeded 0.22 ppm in the areas most severely impacted by

ozone (CARB 1999). 

A network of ozone analyzers located across the SBM was established in the 

1970s. Data from these instruments showed that O3 concentrations decreased along a

gradient from west to east and from south to north across the mountain range (Miller 

and others 1986). Highest concentrations were found at sites facing the LA Basin, 

such as Crestline, Lake Gregory, and Sky Forest. Ozone concentrations decreased

gradually until Barton Flats, and then they more rapidly approached baseline levels at 

Big Bear, Heart Bar, and Holcomb Valley. A more recent survey of air quality in the

SBM indicated that a similar trend continues today, although current O3 levels are 

significantly lower than those recorded in the 1970s (Watson and others 1999). 

It is more difficult to establish historical trends for nitrogen deposition in the 

SBM because systematic monitoring for total N deposition has been conducted only

relatively recently (Fenn and Poth 1999a). However, NOx emissions in the south 

coast (Los Angeles) air basin decreased by 13 percent from 1985 to 1995 and NO2

concentrations have decreased by 39 percent since 1980 (Alexis and others 1999), 

suggesting that N deposition has been decreasing since the 1980s. Deposition of

reduced N forms (NHx) in the western SBM is expected to decrease in the next 10 to 

15 years as dairy operations in the Chino/Norco area, the major source of reduced N

forms in the atmosphere (Luebs and others 1973), move north to central California. 

Measurements of N in throughfall, dry deposition, and fog demonstrate that the

gradient of N deposition is steeper than that of O3 (Fenn and Bytnerowicz 1993).

Total N deposition, based on throughfall data, is estimated to average 31 kg ha-1

 y

-1 in 

the western SBM and 5 kg ha-1

 y

-1 at Barton Flats, located at approximately the 

median of the O3 gradient (Fenn and others 2000). Spatial variability in N deposition 

is high due to heterogeneous canopy cover and varying exposure to air pollution.

Nitrogen deposition in microsites with high leaf area index is greater than plot-level 

average values. In the western SBM, it is estimated that total N deposition is greater 

than 50 kg ha-1

 y

-1 in sites with complete canopy cover and direct exposure to 

incoming pollutant masses. Nitrogen deposition in fog is also high in the western 

SBM, constituting about a third of total annual N deposition (Fenn and others 2000b),

due to extremely high concentrations of N compounds in fog and effective fog

interception by canopy surfaces. Wet and dry deposition of N pollutants is also much 

higher than previously expected in the SBM during the winter months (Fenn and

Bytnerowicz 1997), the season when, compared to summer, ozone concentrations are

extremely low. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 147 
Air Pollution Effects on Conifer Forests—Temple and others 

Effects of Ozone on the Mixed Conifer Forest 

Physiological Responses 

Ozone is a gaseous air pollutant with a molecular weight similar to that of carbon 

dioxide (CO2), and it follows the same diffusive pathway as CO2 from the air

surrounding a plant to intercellular air spaces. The primary mode of entry of gases 

into leaves is through the stomata, and the rate at which this gas exchange occurs is 

known as stomatal conductance. All the physiological and environmental factors that

control stomatal conductance, such as light intensity, temperature, relative humidity,

and soil moisture, also affect the rate at which O3 enters the leaf (Runeckles 1992).

Stomatal conductance is also determined by intrinsic genetic factors often correlated

with the rate of growth of the plant. In general, fast growing annual species and 

agricultural crops have higher rates of stomatal conductance than perennials, shrubs,

and trees. Plants with deciduous leaves have higher rates of conductance than those 

with evergreen foliage. Species adapted for growth under dry conditions have lower 

rates of stomatal conductance and, in general, are less affected by ozone than those

growing under favorable soil moisture conditions. These broad correlations among 

plant growth rates, stomatal conductance, and ozone response are generally useful in

predicting patterns of ozone effects across the landscape, but they cannot be used to 

predict individual species responses to ozone. 

Once inside the leaf, O3 diffuses through intercellular leaf spaces, following the 

diffusive pathways of CO2, to areas of highest cellular physiological activity. The

precise intercellular pathways and mechanisms by which O3 is absorbed into plant 

cells and the biochemical mechanisms of O3 toxicity to intercellular components are 

not well understood (Heath and Taylor 1997). Ozone, or the free radical by-products 

of its initial reactions at the cell surface, may react with components of the cell

membrane, altering active transport of ions such as Ca2+ and K+

 (Castillo and Heath 

1990). This change in membrane polarity will also affect the pH gradient across the 

membrane, leading to a chain of events that include specific enzyme activation and 

deactivation and altered gene transcription (Heath 1987). If the amount of O3

absorbed exceeds cellular detoxification and repair mechanisms, membrane function 

is irrevocably compromised, and normal biochemical activity ceases. One of the first 

indications of O3 toxicity at the cellular level is loss of the enzyme responsible for 

carbon fixation, RUBISCO (ribulose bisphosphate carboxylase/oxygenase; Pell and 

others 1994). This response is reflected at the whole-leaf level by a reduction in the

rate of photosynthesis of leaves exposed to O3. This reduction is reversible at first, 

but as the length of exposure or O3 concentrations increase, the loss of photosynthetic

capacity, accompanied by a loss of chlorophyll, becomes permanent. The first visible

signs of foliar O3 injury are chlorotic areas on the upper leaf surface, which reflect

the permanent loss of photosynthetic capacity in the chloroplasts of the underlying

cells. In broad-leaved plants these lesions usually are sharply defined by the network 

of leaf veins and veinlets, producing a chlorotic stipple over the leaf surface. In 

conifers, particularly in pines, the injury is more diffuse, producing a characteristic 

O3 injury pattern known as chlorotic mottle. As O3 injury accumulates, the leaf 

appears to age prematurely, leading to accelerated senescence and abscission of older 

leaves. The reduction in both rate of photosynthesis and amount of photosynthetic 

tissue in plants exposed to O3 can lead to significant reductions in total amounts of 

carbon fixed over the course of the growing season, with concomitant reductions in 

carbohydrates available for repair, growth, reproduction, and defense (Heath and 

Taylor 1997).

148 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Effects on Conifer Forests—Temple and others 

Ecological Responses 

The first indication that air pollution was having an effect on trees in the mixed

conifer forest of the SBM was the observation in 1953 of significant needle chlorosis, 

necrosis, and premature needle abscission on ponderosa pine (Pinus ponderosa

Laws.) in the vicinity of Crestline-Lake Gregory (Miller and McBride 1975).

Possible scale insect or pathogenic disease etiologies were investigated, but 

observations of similar injury symptoms on grapes growing in the San Gabriel Valley

suggested that an air pollutant might have been involved. Laboratory and field

exposures of ponderosa pine branches to high concentrations of O3 elicited injury

symptoms similar to those observed in the field, and the chlorophyll content of 

needles exposed to O3 was less than that of controls, as was also observed in the field 

(Miller and McBride 1975). Monitoring of pollutant concentrations in the Crestline 

area also revealed concentrations of oxidants similar to those that had elicited pine

needle injury in the lab, confirming the identification of O3 as the causal agent. 

Similar foliar injury symptoms were observed on a variety of plant species in the

SBM, including Jeffrey pine (P. jeffreyi Grev. & Balf.), California black oak 

(Quercus kelloggii Newb.), sycamore (Platanus racemosa Nutt.), big-leaf maple

(Acer macrophyllum Pursh), and several other trees, shrubs, and perennials (Richards

and others 1968). Among conifers native to the SBM, ponderosa and Jeffrey pines

were the most susceptible to O3 injury, followed in order of increasing resistance by

Coulter pine (P. coulteri D. Don), white fir (Abies concolor [G. &G.] Lindl.), bigcone Douglas-fir (Pseudotsuga macrocarpa [Vasey] Mayr), knobcone pine (P. 

attenuata Lemmon), incense cedar (Calocedrus decurrens [Torr.] Florin), sugar pine 

(P. lambertiana Doug.), lodgepole pine (P. contorta ssp. murrayana [Grev. & Balf.]

Critch.), and limber pine (P. flexilis James) (Miller and McBride 1975). 

By the late 1960s observations of the extent of oxidant injury to the mixed

conifer forest from Crestline to Lake Arrowhead showed that over 46,000 ac of forest 

had heavy foliar injury, 54,000 ac had moderate amounts of injury, and 61,000 ac had

light to no injury (Wert and others 1970). Estimates of tree mortality in the most 

heavily polluted areas of the forest suggested that over 25 percent of the pines might

have died as a result of severe O3 injury in combination with other causes (Miller and 

McBride 1975). Pines weakened by O3 were particularly susceptible to attack by pine

bark beetles (Dendroctonus spp.). Ozone-injured pines apparently emitted α-pinene,

a bark beetle attractant (Stark and others 1968). The weakened pines produced less 

sap, the primary defense against bark beetle emergence. Tree mortality of ponderosa 

pines in the heavily damaged areas of the SBM due to increased intensity of bark 

beetle attacks was estimated to be up to 24 percent during a 3-year period in the late 

1960s (Cobb and Stark 1970). Severely O3-injured pines also did not recover from

periods of drought stress as did uninjured trees (Miller and McBride 1975). Thus, 

pine mortality in the SBM was due to severe O3 injury acting in combination with 

other biotic and abiotic stressors. 

Historical Trends in Ozone Injury to the Mixed Conifer Forest 

In the early 1970s, a series of 18 permanent observation plots were established in the 

mixed conifer forest of the SBM along the air quality gradient from west to east (fig. 

1). At each plot, 50 ponderosa and/or Jeffrey pines >30 cm in diameter were

permanently tagged; stem diameter of each was measured, and the amount of O3

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 149 
Air Pollution Effects on Conifer Forests—Temple and others 

injury on the tree crowns was evaluated. Ozone injury was measured using a scale

called the FPM Score, which recorded the youngest annual whorl with visible O3

injury symptoms (FPM=Forest Pest Management; Miller and others 1996). In this 

scale, low numbers indicate higher O3 injury. All other trees on the plots >10 cm

DBH (diameter at breast height) were also measured and evaluated for O3 injury

(Taylor 1973). Elevation, slope, aspect, precipitation, available soil water, and 

estimated O3 dose were also recorded at each plot. Tree condition of pines was 

measured annually from 1973 until 1978 and at a sub-set of these plots at 4-year

intervals from 1978 until 1994. Measurements of crown conditions in 1974 clearly

showed a trend for decreased amounts of O3 injury from plots on the western edge of 

the forest closest to Los Angeles to those in the northeast. Remeasurements of those

same trees in 1988 showed general improvements in crown conditions at all plots

except Camp Paivika, the plot with the highest amount of O3 injury and the highest 

input of ambient O3 (Miller and Rechel 1999). Data from remeasurements of a 

smaller subset of plots in 1994 showed that pines at Camp Paivika showed no 

improvement in O3 injury scores over the 20-year period and that the condition of

trees at Camp Angeles and Heart Bar had only slightly improved (fig. 2). Despite the 

relatively high O3 injury scores and high rates of foliar senescence and abscission of 

ponderosa pines at Camp Paivika, many of these trees have greater relative growth 

rates and higher standing biomass than ponderosa pines growing in less polluted 

areas (Grulke and Balduman 2000). The high rates of N deposition at the western end

of the gradient appear to compensate for, or indeed override, the adverse effects of O3

in this area. 

Although the FPM injury score showed only slight improvement over the 20-

year period from the 1970s until the 1990s, other measures of tree health indicated

that severely injured pines were capable of recovering in response to lower ambient 

O3 in recent years (Miller and Rechel 1999). Measurements of the number of annual

whorls retained on a sample of over 200 ponderosa pines at Dogwood campground, 

near Blue Jay, showed that in 1973 most pines held only two to three annual whorls

(fig. 3). By 1992 many more pines retained four to five annual whorls (fig. 3), and 43

percent of all the pines showed improvements in crown condition as indicated by

increased needle retention (Miller and Rechel 1999). Increased needle retention has 

been reflected in increased tree growth in recent years in the same area (Arbaugh and

others 1999). Tree ring width data from increment borings of 166 ponderosa pines at 

Dogwood campground were converted to increments of basal area. 

Growth patterns of ponderosa pine from the 1880s to the present can generally 

be divided into three phases (fig. 4). Growth gradually increased from the 1880s until

about 1945, but over the next 30 years basal area incremental growth declined, 

reaching its lowest point in the mid-1970s. After 1975, tree growth increased rapidly,

reaching rates similar to those seen before the 30-year growth decline. Although the

post-war decline in tree growth and the subsequent recovery roughly corresponded to

changes in ambient O3 air quality over the same time interval, it is likely that longterm patterns of annual precipitation in the forest were the primary factor controlling 

tree growth. A century of annual precipitation data from Big Bear dam indicates that 

the period 1925 to 1949 was the wettest on record, 1950 to 1974 the driest, and 1975

to 1991 again was a wet period (Arbaugh and others 1999). However, analysis of tree 

ring patterns of individual trees suggested that some pines with significant amounts 

of O3 injury might have been less able to recover from the drought stress than healthy 

trees. A similar pattern of growth in response to O3 and drought stress also has been 

150 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Effects on Conifer Forests—Temple and others 

observed in bigcone Douglas-fir growing along the O3 gradient in the SBM (Peterson

and others 1995). 

Figure 1—Location of permanent forest plots in the San Bernardino National Forest, 

established in 1972 to monitor long-term effects of ozone on ponderosa and Jeffrey 

pines and other dominant tree species in the plots. (Adapted from Miller and Rechel 

1999) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 151 
Air Pollution Effects on Conifer Forests—Temple and others 

Figure 2—Changes in ozone injury scores from 1974 to 1994 for ponderosa and 

Jeffrey pines at three plot locations in the San Bernardino Mountains, based on a 

average of 50 trees per plot. Camp Paivika is in the high ozone area, Heart Bar is a

relatively clean site, and Camp Angeles is intermediate. (Adapted from Miller and

Rechel 1999) 

Figure 3—Average number of annual needle whorls retained by ponderosa pines 

(n=219) in an area of high ambient ozone in 1973 and in 1992. The higher number of 

annual whorls retained in 1992 is an indication of improved tree health. (From Miller 

and Rechel 1999) 

152 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Effects on Conifer Forests—Temple and others 

Figure 4—Average annual basal area increments of ponderosa pines (n=40 to 166,

based upon tree age) growing near Lake Arrowhead, San Bernardino Mountains.

Dashed lines are 95 percent confidence limits. (From Arbaugh and others 1999) 

Ozone and Drought 

The interaction between O3 and drought stress is complex and may involve both

synergistic and antagonistic reactions (Grulke 1999, Temple and Miller 1998,

Temple and others 1993). Trees exposed to O3 under conditions of drought stress will

generally show less O3 injury than those growing under favorable soil-water 

conditions because of reduced rates of gas exchange in the drought-stressed trees.

This interaction between O3 and drought has been observed both under experimental 

conditions (Temple and others 1992, 1993) and in the field (Temple and Miller 

1998). However, if the mechanism of stomatal closure is impaired because of O3

injury (Grulke 1999, Patterson and Rundel 1989, Torsethaugen and others 1999) the 

trees may be more vulnerable to drought stress because of slow or incomplete 

stomatal closure. Reduced root biomass in ponderosa pine trees chronically stressed 

by O3 may also predispose O3-injured trees to increased susceptibility to drought

stress (Grulke and others 1998). Severely O3-injured pines may also increase rates of

stomatal conductance on remaining foliage to compensate for the loss of older annual

whorls (Beyers and others 1992, Coyne and Bingham 1982). This increase in the rate 

of gas exchange may render these needles more susceptible to O3 injury than needles 

from uninjured trees. It is also possible that those individual pines that are

particularly susceptible to O3 may have intrinsically higher rates of stomatal

conductance than more resistant trees, although the evidence for this is currently 

tenuous (Coyne and Bingham 1982, Grulke 1999). 

Seasonal and annual patterns of precipitation will also alter tree responses to O3. 

In the SBM, where almost all soil water is contributed by winter precipitation, the 

total amount of available water in the upper soil column at the beginning of summer 

will control the physiological activity, length of growing season, and thus the total 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 153 
Air Pollution Effects on Conifer Forests—Temple and others 

amount of seasonal O3 flux for most trees in the mixed conifer forest. Significantly

greater amounts of O3 will be absorbed into leaves following a wet winter than 

during a dry year, even if ambient O3 concentrations are the same in both years 

(fig. 5) (Grulke 1999, Temple and Miller 1998). As an additional complication, the

morphology and physiology of pine needles are altered by the environmental 

conditions under which they develop. Leaves developing during periods of favorable

moisture conditions tend to be larger, longer, and have greater intrinsic rates of 

stomatal conductance than those produced during drought years. Those larger needles 

will develop greater amounts of O3 injury than the smaller ones, even when exposed 

to similar ambient O3 concentrations (Temple and Miller 1998). Ponderosa and 

Jeffrey pines will retain greater numbers of annual whorls of needles during periods 

of drought. But in the first wet year following a drought, the great increase in the

amount of O3 injury on older pine foliage leads to accelerated senescence and 

premature abscission of older age classes of needles. As a consequence, litter

accumulation in the high O3 areas of the mixed conifer forest increases greatly in the

year following a drought (Temple and Miller 1998).

Effects of Ozone on Species Composition 

The mixed conifer forest of today in the SBM differs greatly from the forest found by

the original European settlers. The original forests were subjected to frequent lowintensity fires that burned grasses, shrubs, and competing understory trees. The forest

landscape was dominated by large-diameter pines and other fire-resistant species in 

open, irregularly distributed stands. Following a century of intense logging of oldFigure 5—Mean daily ozone flux to 1-year-old needles of mature ponderosa and 

Jeffrey pines growing at Barton Flats, San Bernardino Mountains, from 1992 to 1994. 

Higher fluxes in 1992 reflected improved growing conditions in 1991-1992, following

several previous years of drought. (Adapted from Temple and Miller 1998) 

154 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Effects on Conifer Forests—Temple and others 

growth trees, stand-replacement fires, bark beetle infestations, salvage logging, strict 

fire suppression tactics, O3 damage, and nitrogen deposition, the forest has been

transformed. In the western, more mesic areas of SBM, the open growth, fireresistant pine-dominated forest has become a closed-canopy forest dominated by 

dense stands of incense cedar and white fir (Minnich 1999). While fire-suppression 

policies have played a major role in this transformation, the impacts of air pollution 

should not be minimized. Ponderosa and Jeffrey pines are significantly more 

susceptible to O3 than are incense cedar and white fir (Miller and McBride 1978), and 

they suffered far greater damage due to O3. Salvage logging removed over 25 percent 

of mature pines damaged by combinations of O3 and bark beetle infestations in the

1970s (Miller and McBride 1975). Measurements of basal diameter growth of all

trees >30 cm in diameter in plots along the O3 gradient from west to east showed that 

in high O3 areas ponderosa pine basal area growth was reduced relative to incense 

cedar and white fir (Miller and others 1997). Measurements of tree seedling mortality 

and establishment also showed increased success of incense cedar and white fir 

seedlings and saplings, relative to that of pines (McBride and Laven 1999). Ozone 

injury to pines also greatly increased the amount of leaf litter on the forest floor in

high O3 areas of the forest, because of accelerated senescence and abscission of O3-

injured pine needles (Arkley and others 1980). Rates of leaf litter accumulation and 

decomposition may also be significantly increased by the higher rates of N deposition 

at these plots on the western side of the SBM (Fenn 1991). The thick litter layer 

discourages the establishment of pine seedlings, which is an additional factor in the 

transformation of the mixed conifer forest from open-growth pine stands to a closed

canopy, dense incense cedar/white fir-dominated forest. 

The effects of O3 in altering species composition and community dynamics of 

forest understory species are less well known. Ozone injury symptoms have been

noted since the 1950s on understory species such as blue elderberry (Sambucus 

mexicana C. Presl.), wild grape (Vitis girdiana Munson), and several woody species

of riparian areas such as sycamore (Platanus racemosa Nutt.) and big-leaf maple 

(Acer macrophyllum Pursh) (Richards and others 1968). Foliar O3 injury symptoms 

also have been observed on a variety of annual and perennial herbaceous plants in the 

mixed conifer forest (Temple 1999). However, unlike the systematic study of O3

effects on injury and growth of overstory trees, no careful chronological sequence 

measurements of understory species frequency or density have been conducted along

the air pollution gradient in the SBM. Shrub and herbaceous plant cover was 

measured once along the pollution gradient in 1973 (Taylor 1973), but these 

observations were not sufficient to establish a causal relationship between ambient O3

concentrations and changes in community structure or composition. The same species

of herbaceous annuals and perennials and woody shrubs recorded in plots in high O3

areas in 1972 were also present in 1996 (Temple 1999). Some of these species, such

as blue elderberry and mugwort (Artemisia douglasiana Besser) are also highly

susceptible to O3 (Temple 1999). These observations, while hardly definitive, suggest

that O3 may have had only minor direct effects on understory plant community

composition in the SBM in comparison with other factors, such as habitat alteration, 

fire suppression, invasion by weedy plants, and N deposition (see below). 

Most of the rare, threatened, endangered or endemic plant species in the SBM 

are found to the north and east of high O3 areas, primarily in the Bear Valley and on 

the limestone deposits on the north slope of the mountains. However, at least 23 

species or subspecies of rare or threatened plants are found in western areas of the 

forest (Skinner and Pavlik 1994). Scattered observations on some of these species did 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 155 
Air Pollution Effects on Conifer Forests—Temple and others 

not reveal any foliar O3 injury in recent years (Temple 1999). However, no 

systematic surveys of O3 injury have been conducted on these plants, so no 

conclusions can be drawn regarding the impacts of O3 on threatened and endangered

plant species in the SBM. Epiphytic lichen communities in the San Bernardino

Mountains have been dramatically reduced both in species diversity and in numbers 

both historically and relative to comparable habitats with little air pollution (Nash 

and Sigal 1999). However, the role of O3 in this reduction, relative to other factors 

such as N deposition or other pollutants, has not been determined. 

Nitrogen Deposition in the Mixed Conifer Forest 

Nitrogen Cycling in Forest Ecosystems 

Nitrogen is the mineral nutrient most limiting to growth of forest trees. Most N in 

forest ecosystems is stored in the soil in the form of organic N, not readily 

assimilable by plant roots. Mineralization of organic N and nitrification convert N to 

forms readily available to plants. Because mineral N is generally present in limiting

amounts in forest ecosystems, these systems have evolved tightly-coupled processes 

to conserve N through the cycle of litter decomposition, mineralization, bacterial and

fungal associations, and internal N cycling (Chapin 1980). Conifers in particular are 

adapted to nutrient-poor soils, and they have evolved efficient mycorrhizal associations for nutrient uptake from the soil and internal retranslocation of N from 

senescent foliage prior to abscission (Chapin 1980). Because of this tight coupling of 

N deposition, mineralization, and uptake, cycling of N through forest ecosystems is 

generally highly efficient, and in a healthy ecosystem leaching of mineral N is low 

(Johnson 1992). Mixed conifer forests of California are no exception, and trees 

respond with increased growth to added N (Kiefer and Fenn 1997). Thus, increased

rates of deposition of atmospheric N may actually increase tree growth if other 

nutrients and particularly water are not also limiting. However, if N deposition from 

atmospheric sources is in excess, normally conservative processes of N cycling and N 

retention within the ecosystem can be disrupted, producing a condition referred to as 

N saturation (Aber and others 1995). 

Nitrogen Deposition in the San Bernardino Mountains 

Recent studies in the mixed conifer forest in the SBM have suggested that western

sections of the forest are currently experiencing N-saturated conditions (Fenn and 

Poth 1999a). Rates of N deposition, primarily in the form of dry deposited nitrate,

ammonium, and nitric acid and N deposition in fog, can exceed 30 kg ha-1

 y

-1 as an 

average near Crestline/Lake Gregory and points west (Fenn and Bytnerowicz 1993).

As mentioned earlier, total N deposition is estimated to be as high as 50 kg ha-1

 y

-1 in 

sites with complete canopy cover and direct exposure to polluted air masses (Fenn

and others 2000). In contrast, total N deposition on the eastern side of the SBM is 

generally <5 kg ha-1

 y

-1. Evidence for N-saturation in areas of high N deposition 

includes increased nitrate concentrations in stream water from contaminated

watersheds, elevated nitrate concentrations in soil solution and soil leachate,

increased volatilization of N from the soil, decreased C:N ratios in soil and foliage, 

increased N concentrations and N:P ratios in foliage, nitrate accumulation in foliage

of understory and overstory plants, and altered rates of litter decomposition (Fenn

and others 1996, Fenn and Poth 1999a, 1999b). 

156 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Effects on Conifer Forests—Temple and others 

Effects of Excess N Deposition 

The direct and indirect effects of excess N deposition on the mixed conifer forest are 

under investigation. Deposition of nitric acid on plant foliage can alter surface wax

structure and chemical composition (Bytnerowicz and others 1998b). In addition, 

nitric acid deposited on plant leaves may be a source for plant N, both through a

transcuticular pathway (Hanson and Garten 1992) and stomatal uptake (Bytnerowicz 

and others 1998a). Concentrations of N in plant foliage from contaminated areas of

the forest are significantly higher than those from areas of low N deposition (Fenn 

and others 1996). As a result of the combined effects of ozone stress (enhanced foliar 

senescence and abscission) and nitrogen (greater foliar and branch growth), litter 

production is increased in the western SBM. Litter with high N concentrations 

frequently exhibits higher decomposition rates in the short term (Fenn 1991, Fenn

and Dunn 1989), but numerous studies show greatly reduced decomposition rates and

greater humus accumulation in the later stages of decomposition when litter N 

content is high (Berg 2000, Fog 1988). Thus, N enrichment as a result of chronic N 

deposition favors litter accumulation on the forest floor by stimulating litter 

production and simultaneously causing long-term reductions in litter decomposition 

rates. The available evidence also suggests that C sequestration also may be greater

in these forests influenced by the combined effects of ozone and N, with greater

amounts of C stored in the forest floor and in woody biomass. 

As noted earlier, the thick litter layer in the western SBM can retard germination 

of pine seeds, contributing to the change in species composition of the forest. It has 

not been clearly established whether higher rates of N deposition have increased

growth of pines or other trees in the mixed conifer forest, but growth rates are higher 

in the western high pollution sites than in the less polluted sites to the east, 

notwithstanding the high O3 concentrations that co-occur with N deposition (Arbaugh 

and others 1999, Grulke and Balduman 1999, M.E. Fenn and M.A. Poth, unpublished 

data). However, other environmental factors such as elevation, precipitation, 

temperature, and soil characteristics also vary across this gradient, making it difficult 

to determine the most important factors causing higher growth rates in the high 

pollution sites. Interactions between O3, which can inhibit tree growth, and N, which

can enhance growth, are complex, particularly as both interact with seasonal and

annual variations in soil water availability. However, the higher growth rates in the 

western SBM suggest that elevated N deposition in areas such as Camp Paivika has 

acted to ameliorate the adverse effects of O3 on the growth of ponderosa pine in this 

area of the forest. The effects of excess N on N-fixation, mycorrhizal associations, N 

mineralization, insect and disease infestations, and other ecosystem processes need

further investigation (Fenn and Poth 1999a). 

Nitrogen enrichment of the mixed confer forest may also play a role in altering

species composition and competitive interactions among species (Takemoto and 

others 2000). Because most native forest species have adapted to an N-limited

environment, they may be at a competitive disadvantage to species able to utilize the 

extra N from atmospheric deposition to enhance growth and reproductive potential. 

The latter may be particularly true for alien weedy species of grasses that have

invaded the mixed conifer forest in recent years. Excess atmospheric deposition of N 

may have contributed to the recent rapid conversion of coastal sage shrub habitats to 

annual grasslands dominated by weedy annual species of Mediterranean origin (Allen 

and others 1998). However, there is as yet no direct evidence that excess N has 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 157 
Air Pollution Effects on Conifer Forests—Temple and others 

contributed to the invasion of the mixed conifer forest in the SBM by weedy species

or has altered competitive interactions among native species of annuals, perennials, 

or woody plants. Bracken fern (Pteridium aquilinum [L.] Kuhn), a nitrophilous 

species with a known capacity to accumulate nitrate in foliage (Fenn and others 

1996), is a dominant, vigorously growing understory species at Camp Paivika, an Nsaturated site in the SBM. However, it is not known whether long-term N deposition 

has favored bracken fern or whether bracken fern was already the dominant 

understory species prior to the occurrence of elevated N deposition. 

The clearest response of the mixed conifer forest in the SBM to excess N

deposition is the greatly increased concentrations of nitrate in stream water draining

areas with high N deposition. Elevated nitrate loss is observable at the watershed

scale and is the classic symptom of an N-saturated ecosystem (Fenn and others 

1996). Surprisingly, and notwithstanding the N enrichment of soil and foliage in the 

western SBM, the bole diameter growth of ponderosa pine and California black oak 

trees at Camp Paivika increased significantly after three years of N fertilization at 50 

or 150 kg ha-1 yr-1. It is not yet clear how the forest can export large quantities of

excess N in stream water and as gaseous N emissions from soil and yet the 

production of woody biomass is still N-limited. Earlier studies indicated that foliar

biomass production in ponderosa pine was not N-limited at Camp Paivika, but it was 

N-limited in low pollution sites in the SBM and San Gabriel Mountains (Kiefer and 

Fenn 1997). Further studies are underway to address this phenomenon. The 

interactive effects of high N deposition and ozone on conifer growth and carbon 

allocation are also under active investigation. Early results indicate that O3 injury and 

N deposition can interact to increase rates of foliage loss and reduce fine root growth 

of ponderosa pine (Grulke and Balduman 2000). The consequences of this interaction

in relation to drought stress and seasonal patterns of soil water availability are 

currently under investigation. 

In a 3-year study of 19 streams across the SBM, a strong relationship was found 

between N deposition rates and nitrate concentrations in stream water (Fenn and Poth 

1999b). Stream nitrate concentrations generally peaked in mid-winter, in response to

large volume winter storms, although extremely high peak summertime 

concentrations also occurred following summer thundershowers in some streams 

draining the San Gorgonio Wilderness. Asynchronies between periods of maximum 

availability of mineralized N and biological demand contribute to increased leakage 

of N from the system (Fenn and Poth 1999a, 1999b). This temporal asynchrony is a

function of the Mediterranean climate, characterized by major runoff periods in the 

winter, while plant N demand is expected to be greatest in the warm, dry spring and 

summer seasons. Another critical factor contributing to the highly open N cycling

nature of these N saturated systems is the high nitrifying potential of the soil. 

Laboratory and field studies of N mineralization and nitrification show that 

ammonium is rapidly and completely nitrified in soil. This means that all the 

mineralized N is rapidly converted to nitrate, a highly mobile form of N, for which 

there are no effective retention mechanisms in soil except for microbial and plant 

uptake. The possibility that high concentrations of nitrate in stream water may be

toxic to aquatic organisms during sensitive periods in their reproductive cycle will be

investigated in future studies (Fenn and others 2000a).

Nitrate concentrations in streams of the SBM and San Gabriel Mountains are the 

highest reported for wildland ecosystems in North America. Peak concentrations 

approaching 400 μEq L-1 (fig.6; Fenn and Poth 1999b, Riggan and others 1985) are 

158 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Air Pollution Effects on Conifer Forests—Temple and others 

not unusual during high runoff periods. In the more severely N saturated watersheds 

in the SBM, nitrate levels in streams and springs remain elevated (>70 μEq L-1) yearround. By comparison, in relatively undisturbed forested watersheds, peak nitrate

levels are typically 5-10 μEq L-1 or lower, and baseline values are beyond detectable 

limits. 

Figure 6—Nitrate concentrations in streamwater draining from areas of high nitrogen 

deposition (Devil Canyon) and low to moderate deposition (San Gorgonio 

Wilderness) in the San Bernardino Mountains, 1995 to 1997. (Adapted from Fenn

and Poth 1999) 

Summary

Montane forests in southern California have been subjected to high levels of ozone 

air pollution and deposition of nitrogen compounds for several decades. Severe O3

injury to ponderosa and Jeffrey pines and other forest tree species was observed

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 159 
Air Pollution Effects on Conifer Forests—Temple and others 

beginning in the late 1950s. By the early 1970s, over 25 percent of the pines growing

on the west side of the San Bernardino Mountains had died because of O3 damage in 

combination with bark beetles, drought, and other natural stressors. The selective 

removal of pines in association with fire-suppression policies has contributed to the 

change in species composition of the mixed conifer forest, from open, pinedominated stands to dense, closed stands dominated by incense cedar and white fir. 

Ambient O3 concentrations in the mountain forests are now significantly lower than 

in previous decades, and crown conditions and growth rates of pines have improved 

in recent years. However, O3 injury to pines and other species of native annuals, 

perennials, and woody plants of the mixed conifer forest continues to be observed on

slopes and ridge-tops facing Los Angeles, and O3 continues to have significant 

adverse impacts on these forests. 

The effects of excess N deposition on the mixed conifer forest have only 

recently been investigated. Evidence has accumulated that the westernmost areas in

the SBM are N-saturated, with consequences for biogeochemical cycling of N, tree

growth, C sequestration, and export of N from the forest ecosystem. Both N

deposition and ozone toxicity interact with seasonal and annual variations in 

precipitation and in soil water availability. The effects of excess N deposition in the 

mixed conifer forest on productivity, biogeochemical cycling, and soil and stream 

water hydrology are areas of continuing investigation. 

Acknowledgments 

The authors would like to acknowledge the extraordinary contributions of Dr. Paul R. 

Miller to the research on ecosystem effects of atmospheric deposition in the San

Bernardino Mountains. 

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164 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Sensitive Species of Snakes, Frogs, and 

Salamanders in Southern California Conifer 

Forest Areas: Status and Management1

Glenn R. Stewart

2

, Mark R. Jennings3, and Robert H. Goodman, Jr.4

 

Abstract 

At least 35 species of amphibians and reptiles occur regularly in the conifer forest areas of

southern California. Twelve of them have some or all of their populations identified as 

experiencing some degree of threat. Among the snakes, frogs, and salamanders that we 

believe need particular attention are the southern rubber boa (Charina bottae umbratica), San

Bernardino mountain kingsnake (Lampropeltis zonata parvirubra), San Diego mountain

kingsnake (L.z. pulchra), California red-legged frog (Rana aurora draytonii), mountain 

yellow-legged frog (R. muscosa), San Gabriel Mountain slender salamander (Batrachoseps 

gabrieli), yellow-blotched salamander (Ensatina eschscholtzii croceater), and large-blotched

salamander (E.e. klauberi). To varying degrees, these taxa face threats of habitat degradation 

and fragmentation, as well as a multitude of other impacts ranging from predation by alien 

species and human collectors to reduced genetic diversity and chance environmental

catastrophes. Except for the recently described San Gabriel Mountain slender salamander, all 

of these focus taxa are included on Federal and/or State lists of endangered, threatened, or

special concern species. Those not federally listed as Endangered or Threatened are listed as

Forest Service Region 5 Sensitive Species. All of these taxa also are the subjects of recent and 

ongoing phylogeographic studies, and they are of continuing interest to biologists studying the

evolutionary processes that shape modern species of terrestrial vertebrates. Current 

information on their taxonomy, distribution, habits and problems is briefly reviewed and 

management recommendations are made. Further research is needed to elucidate their 

biological status and needs and to provide the basis for appropriate management programs. 

Programs must be monitored to ensure that desired objectives are achieved. 

Key Words: collecting, degradation, endangered species, ensatina, extinction, extirpation,

fragmentation, fuel wood, habitat loss, mountain kingsnake, off-highway vehicle, red-legged

frog, riparian, rubber boa, salamander, special concern, threatened species, yellow-legged frog

Introduction 

Our area of concern in this paper is limited primarily to the conifer forest and 

included riparian habitats above 1,230 m in roughly the southern two-thirds of the

southern California mountains and foothills assessment region described by 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2 Professor of Zoology and Environmental Science, California State Polytechnic University, Pomona, 

California 91768 (e-mail: grstewart@csupomona.edu) 3

 Rana Resources, 39913 Sharon Avenue, Davis, California 95616 and Research Associate, Department 

of Herpetology, California Academy of Sciences, San Francisco, California 94118 (e-mail: 

RanaResources@aol.com) 

4

 Professor of Forestry and Natural History, Citrus College, Glendora, California 91741 (e-mail: 

rgoodman@citrus.cc.ca.us) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 165 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

Stephenson and Calcarone (1999). This elevation break generally corresponds to the

lower limit for conifer forest habitats at this latitude. The area includes the south 

coast ranges (Sierra Madre, San Rafael, Santa Ynez, and Topatopa Mountains and 

Pine Mountain) in Ventura County, the Mount Pinos area straddling the VenturaKern County line, the Piute and Tehachapi Mountains in east-central Kern County,

the transverse ranges (Santa Monica, San Gabriel, and San Bernardino Mountains) in

Los Angeles and San Bernardino Counties, the peninsular ranges (San Jacinto, Santa 

Rosa, and Laguna Mountains) in Riverside and San Diego Counties, the Santa Ana 

Mountains in eastern Orange and western Riverside Counties, and certain isolated

mountains in San Diego County. At least 35 species of amphibians and reptiles are 

known to occur regularly in the area: five salamanders, seven frogs and toads, one 

turtle, eight lizards, and 14 snakes (table 1). Many of these are wide-ranging species

that include the mountains within their geographic range, but others are more or less 

restricted to montane habitats. 

Mostly adjacent to the greater Los Angeles-San Diego metropolitan area, with a 

population of over 15 million people (Stephenson and Calcarone 1999), many of

these amphibian and reptile species are impacted negatively by human activities, and 

12 have some or all of their populations identified as experiencing some degree of

threat (table 1). We discuss here three taxa of snakes, two taxa of frogs, and three 

taxa of salamanders that we believe are in particular need of attention as sensitive 

species. Although the California red-legged frog generally occurs at lower elevations 

peripheral to conifer forest areas, and ranges farther north along the coast, it is 

included because of its critically threatened status and omission from other habitat 

area discussions in these proceedings. 

Southern Rubber Boa Charina bottae umbratica Klauber 

1943 

Description 

A small (total length 35-55 mm, Hoyer and Stewart 2000a), rather stout-bodied snake 

with smooth scales and a blunt head and tail (fig. 1; Brown 1997). Adults are light 

brown or tan in dorsal color with an unmarked yellow venter; juveniles are pale 

without a distinct margin between dorsal and ventral coloration (Hoyer and Stewart 

2000a). The subspecies may be further characterized as follows: frontal scale usually 

sub-rectangular with a straight or only slightly convex or angular posterior margin, 

generally 41 or fewer mid-dorsal scale rows, usually 196 or fewer full-sized ventral 

scales (Rodriguez-Robles and others 2001). 

166 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

Table 1−Amphibian and reptile species known to occur regularly above 1,230 m in conifer 

forest areas of southern California. (All names follow Crother 2000.)

Salamanders 

 Ensatina-Ensatina eschscholtzii (SC, FS)*

Black-bellied Slender Salamander-Batrachoseps nigriventris

Tehachapi Slender Salamander-Batrachoseps stebbinsi (ST, FS)

San Gabriel Mountain Slender Salamander-Batrachoseps gabrieli (FS)*

 Arboreal Salamander-Aneides lugubris 

Frogs and Toads 

 Pacific Treefrog-Pseudacris regilla

 California Treefrog-Pseudacris cadaverina 

 Red-legged Frog-Rana aurora (FT, SC)*

 Mountain Yellow-legged Frog-Rana muscosa (FE, SC)*

 American Bullfrog-Rana catesbeiana (introduced) 

 Western Toad-Bufo boreas

 Arroyo Toad-Bufo californicus (FE, SC) 

Turtles 

Pacific Pond Turtle-Clemmys marmorata (SC, FS) 

Lizards 

 Coast Horned Lizard-Phrynosoma coronatum (SC, FS) 

 Granite Spiny Lizard-Sceloporus orcutti 

Western Fence Lizard-Sceloporus occidentalis

Common Sagebrush Lizard-Sceloporus graciosus

 Common Side-blotched Lizard-Uta stansburiana

 Western Skink-Eumeces skiltonianus

 Gilbert's Skink-Eumeces gilberti

Southern Alligator Lizard-Elgaria multicarinata

Snakes 

 Western Threadsnake-Leptotyphlops humilis

 Rubber Boa-Charina bottae (ST, FS)*

 Ring-necked Snake-Diadophis punctatus (FS)

 Striped Racer-Masticophis lateralis

 Gophersnake-Pituophis catenifer

 Common Kingsnake-Lampropeltis getula

California Mountain Kingsnake-Lampropeltis zonata (SC, FS)* 

Western Terrestrial Gartersnake-Thamnophis elegans

 Two-striped Gartersnake-Thamnophis hammondii (SC, FS) 

Western Black-headed Snake-Tantilla planiceps

 Western Lyresnake-Trimorphodon biscutatus

 Nightsnake-Hypsiglena torquata

 Speckled Rattlesnake-Crotalus mitchellii

 Western Rattlesnake-Crotalus viridis 

FE–Federally Listed Endangered Species

FT–Federally Listed Threatened Species

ST–State Listed Threatened Species

SC–State Species of Special Concern 

FS–Forest Service Region 5 Sensitive Species 

*Species discussed in text 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 167 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

Figure 1−Adult female southern rubber boa from Heaps Peak, San Bernardino

Mountains, San Bernardino County, California. Photo by Robert H. Goodman, Jr. 

Taxonomy 

The infraspecific taxonomy of the rubber boa (Charina bottae) has been much

debated (Brown 1997, Stewart 1988), with various authors recognizing either three 

(Klauber 1943, Stebbins 1985), two (Stewart 1977), or no subspecies (Collins 1990,

Nussbaum and Hoyer 1974), or suggesting full species status for the southern form 

(Erwin 1974). While scalation patterns are highly variable within rubber boa 

populations, and there is considerable overlap among populations, Stewart (1977) 

noted fairly consistent differences between southern rubber boas (C.b. umbratica) 

and northern rubber boas (C.b. bottae). He recommended that these forms be retained 

as distinct subspecies and regarded populations in the Mount Pinos area and

Tehachapi Mountains as intergrades. Recent analyses of mitochondrial DNA by 

Rodriguez-Robles and others (2001) have shown that the southern rubber boa 

populations form a clade that separated from all other rubber boa lineages 

approximately 4.4 to 12.3 million years ago. Together with existing morphological 

and allozyme data (Weisman 1988), this suggests C.b. umbratica could be

recognized as a full species. Although the Mount Pinos area and Tehachapi 

Mountains populations cannot be identified reliably by morphological characters, 

they cluster with C.b. bottae in the mitochondrial DNA analyses (Rodriguez-Robles 

and others 2001). 

168 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

Distribution 

The southern rubber boa is restricted to the San Bernardino and San Jacinto

Mountains (fig. 2) at elevations between about 1,540 m and 2,460 m (Stewart 1988,

1991). It is known from over 40 localities in the San Bernardinos (Stewart 1991) and 

eight in the San Jacintos (Keasler 1982, Loe 1985). Twenty-six of the San Bernardino 

localities occur in a 16-km-long strip of habitat in the Lake Arrowhead area between

Twin Peaks on the west and Green Valley on the east (Stewart 1988, 1991). 

Habitat and Habits 

Typical habitat for this snake is mixed conifer-oak forest or woodland dominated by two or more of the following species: Jeffrey pine (Pinus jeffreyi), 

yellow pine (P. ponderosa), sugar pine (P. lambertiana), incense cedar (Calocedrus decurrens), white fir (Abies concolor), and black oak (Quercus kelloggii) (Stewart 1988). A relatively open canopy seems to be preferred, at least 

during the spring months (April and May), and rock outcrops evidently are 

important as hibernacula (Keasler 1982, Stewart 1988). In the Heaps Peak 

area near Lake Arrowhead, the boa occurs at rock outcrops in open areas

characterized by mixed grasses and bracken fern (Pteridium aquilinum) together with variable numbers of shrubs and small trees (Hoyer and Stewart 

2000a). Because it is semi-fossorial, primarily crepuscular or nocturnal, and

Figure 2–Historic and current distribution of the southern rubber boa 

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highly secretive, its seasonal activity and habitat use are difficult to determine. Boas 

are most easily found during the spring months at rock outcrops where scattered surface rocks and wood debris provide movable cover under which to search (Hoyer and 

Stewart 2000a, Keasler 1982, Stewart 1988,). The snakes essentially disappear during 

the summer months, though they may emerge for surface activity on humid nights 

(Alten and Keasler 1978) or after a rain (Stewart pers. observ.). It is likely that some 

individuals remain at the rock outcrops during the summer and retreat deeper into

crevices (Stewart pers. observ.), but others apparently move into cooler, moister forest and riparian habitats (Loe 1985, Stewart 1988). In all habitat types, rock outcrops 

and surface materials (such as rocks, logs, and a well-developed litter/duff layer) are

important habitat components because they provide cover and maintain soil moisture 

(Loe 1985). Lizards, particularly common sagebrush lizards (Sceloporus graciosus) 

and western fence lizards (S. occidentalis), often are associated with the outcrops 

where southern rubber boas are found. Hoyer and Stewart (2000b) demonstrated that 

both juvenile and adult boas readily consume lizard eggs, with adults also taking

nestling voles (Microtus spp.) and shrews (Sorex spp.). These snakes are viviparous;

Hoyer and Stewart (2000a) recorded a litter size of 2 to 5 in 12 females. 

Status and Threats 

The southern rubber boa is listed by the State of California as Threatened and is a 

Forest Service Region 5 Sensitive Species (Stephenson and Calcarone 1999). While 

an assessment of the boa's current status is hindered by its secretive behavior, it is

continuing to experience significant habitat loss and degradation. Most of the prime 

habitat in the 16-km strip near Lake Arrowhead is private land subject to residential 

and commercial development (Stewart 1988, 1991). Elsewhere in the San Bernardino 

and San Jacinto Mountains, substantial areas of known or potential habitat are on San 

Bernardino National Forest lands (Loe 1985, Stewart 1991). However, Forest 

Service-permitted activities, such as personal fuelwood harvesting, off-highway 

vehicle use, and commercial timber sales, increasingly degrade and fragment what 

once were large, contiguous tracts of rubber boa habitat (Loe 1985, Stewart 1991).

The combined factors of habitat degradation, fragmentation, and loss increase the 

likelihood of local extirpations and restrict gene flow, potentially leading to reduced 

genetic diversity and greater vulnerability to chance environmental catastrophes. 

Management Recommendations 

A Southern Rubber Boa Advisory Committee was established in 1980, and with its 

assistance the San Bernardino National Forest produced a comprehensive set of 

habitat management guidelines (Loe 1985). The guidelines seek to maintain habitat 

quality, minimize negative impacts on Forest Service lands, and coordinate 

management activities with county and city governments. We recommend continued

adherence to these guidelines, with monitoring to insure that objectives are being 

met. A better understanding of seasonal movements and habitat utilization, obtained 

through radio telemetry studies, would permit fine-tuning of management to the

needs of this snake. Special efforts should be made to establish protective 

management on a large (30 hectares or more) tract of private property containing boa

habitat in the Heaps Peak area (Stewart 1991) and several smaller (10 hectares or so) 

tracts elsewhere within the boa's range. This could be done through conservation

easements with willing landowners and/or by setting up a land-bank system that

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developers pay into to get mitigation credits for their habitat impacts. Ideally, these 

acquired properties should be adjacent to public lands to maximize the size of

contiguous habitat areas. Although rubber boa populations in the Mount Pinos area 

may now be assigned to the northern rubber boa (Rodriguez-Robles and others 2001), 

similar guidelines should be developed for these populations by the Los Padres 

National Forest. They are small and isolated (Alten and Keasler 1978, Hoyer and 

Stewart 2000b), and they provide opportunities for further study of evolutionary 

processes. 

San Bernardino Mountain Kingsnake Lampropeltis 

zonata parvirubra Zweifel 1952 

Description 

A medium-sized (total length 55-111 cm, Jennings and Hayes 1994a) snake with a 

tricolor pattern of black, white, and red rings around the body (fig. 3; Brown 1997).

Diagnosis of the different subspecies of the California mountain kingsnake 

(Lampropeltis zonata) is based almost entirely on color pattern characteristics. The 

basic color unit is a triad composed of a black ring, partly or completely divided by a

red band or ring, and separated from adjacent black rings by a pair of white rings 

(Rodrigues-Robles and others 1999, Zweifel 1952, 1974). The number of body triads 

in L.z. parvirubra typically is 37 or more, and usually fewer than 60 percent are 

completely divided by a red ring (Zweifel 1974). 

Figure 3–San Bernardino mountain kingsnake from Sawpit Canyon, San Gabriel

Mountains, Los Angeles County, California. Photo by Robert H. Goodman, Jr. 

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Taxonomy 

Identification of this taxon is problematic because of the wide range of variation and 

broad overlap of its color characteristics with other subspecies of L. zonata. 

Examining a large series of specimens representing all seven of the named

subspecies, Rodriguez-Robles and others (1999) concluded that color patterns cannot

be used to reliably differentiate among them. However, mitochondrial DNA analyses 

by these authors showed that the San Bernardino mountain kingsnake and San Diego 

mountain kingsnake (L.z. pulchra), together with two other subspecies from northern 

Baja California, Mexico, form a southern clade that may be distinct from the more

northern subspecies (Rodriguez-Robles and others 1999). Resolution of California

mountain kingsnake systematics must await the completion of additional 

morphological and genetic studies. In the meantime, we believe it is important to

continue to recognize the named infraspecific taxa of this mountain kingsnake to help 

ensure that the diversity of the herpetofauna in the southern California mountains is

adequately protected. 

Distribution 

The San Bernardino mountain kingsnake is restricted to the San Gabriel, San 

Bernardino, and San Jacinto Mountains (fig. 4) at elevations from about 370 m to

2,470 m (Jennings and Hayes 1994a). 

Figure 4–Historic and current distribution of the San Bernardino mountain kingsnake. 

Modified from Jennings and Hayes (1994a). 

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Habitat and Habits 

This snake tends to occur in well-illuminated rocky canyons in association with bigcone Douglas-fir (Pseudotsuga macrocarpa) and various chaparral species at lower

elevations and with black oak, incense cedar, Jeffrey pine, and yellow pine at higher 

elevations (Jennings and Hayes 1994a). Above 1,540 m in the San Bernardino and

San Jacinto Mountains, it commonly is found in rock outcrops together with the 

southern rubber boa (Hoyer and Stewart 2000a, Keasler 1982). Rock crevices and

outcrops likely serve as hibernacula and cover, and they also provide potential food

sources. Mountain kingsnakes are known to be primarily lizard eaters, most often 

taking common sagebrush lizards, western fence lizards, and western skinks 

(Eumeces skiltonianus) (Jennings and Hayes 1994a). While secretive and 

infrequently observed, mountain kingsnakes are more likely to move about on the 

surface during daylight hours of spring months than are southern rubber boas

(McGurty 1988, Stewart pers. observ.). Later in the season and at lower elevations, 

they may be active after dark (Jennings and Hayes 1994a). As in the case of rubber 

boas, rock outcrops and a variety of surface materials (rocks, logs, bark) probably are

important habitat components in all habitat types (McGurty 1988). Mountain

kingsnakes are oviparous and have been reported to produce clutches of 3-9 eggs 

(McGurty 1988). 

Status and Threats 

The San Bernardino mountain kingsnake is a State of California Species of Special 

Concern (Jennings and Hayes 1994a) and a Forest Service Region 5 Sensitive Species 

(Stephenson and Calcarone 1999). It is impacted by most of the same factors (habitat 

degradation, fragmentation, and loss) that affect the southern rubber boa. In addition,

it is highly prized by reptile fanciers and continues to be collected despite State laws

limiting collecting and preventing the sale of native reptiles in California (California 

Freshwater Sport Fishing Regulations 2002, Jennings and Hayes 1994a). While the 

impact of habitat alteration and collecting pressures is difficult to assess, it results 

from the ripping apart of decaying logs and stumps and removal of exfoliating pieces

of granite boulders, practices that are prohibited by State law (California Freshwater

Sport Fishing Regulations 2002). These microhabitats probably require decades to

millennia to develop, but only seconds to minutes to destroy. Their loss may lead to 

local extirpations and reduced population viability, especially at locations easily

accessed by roads. 

Management Recommendations 

Implementation of management recommendations for the southern rubber boa in the 

San Bernardino National Forest also benefits the San Bernardino mountain

kingsnake. In addition, particular attention should be paid to maintaining surface 

cover and minimizing human impacts in rocky draws and canyons. Similar

management guidelines should be developed for the Angeles National Forest. 

Because visitors to National Forest lands commonly confuse mountain kingsnakes

with venomous coral snakes (Jennings pers. observ.), which also have a tri-color

pattern but are not found in California, posters and other educational materials should 

be provided at convenient locations to counter misinformation about snake 

identification and biology. Again, radio telemetry studies would provide more 

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information on the activity and habitat use patterns of the San Bernardino mountain

kingsnake and allow for more precise management prescriptions. 

San Diego Mountain Kingsnake Lampropeltis zonata 

pulchra Zweifel 1952 

Description 

A medium-sized (total length 53-108 cm, Jennings and Hayes 1994a) snake with a 

tricolor pattern of black, white and red rings (fig. 5; Brown 1997) similar to the San

Bernardino mountain kingsnake. On the average, however, there are fewer body

triads and more of the triads have complete red rings (Rodriguez-Robles and others 

1999). Zweifel (1974) gave the number of triads as typically 36 or fewer, and the 

number of triads with complete red rings as usually 60 percent or more. 

Figure 5–San Diego mountain kingsnake from Boiling Springs, Cuyamaca Mountain, 

San Diego County, California. Photo by Robert H. Goodman, Jr. 

Taxonomy 

Taxonomic comments made for the San Bernardino mountain kingsnake also apply 

to this subspecies. 

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Distribution 

The San Diego mountain kingsnake occurs in the Santa Monica, Santa Ana, Santa 

Rosa, and Laguna Mountains. In San Diego County, it also occurs on the more 

isolated Palomar, Hot Springs, Cuyamaca and Corte Madera Mountains (fig. 6). It 

ranges in elevation from near sea level to about 1,800 m (Jennings and Hayes 1994a). 

Habitat and Habits 

Much of the information given for the San Bernardino mountain kingsnake is 

applicable to the San Diego subspecies. In the interior mountain ranges, it occurs 

primarily in associations of yellow, Jeffrey and Coulter pine (Pinus coulteri) and 

black oak. At lower elevations and in coastal areas, it usually is found in canyon

bottoms with western sycamore (Platanus racemosa), Fremont cottonwood (Populus 

fremontii), coast live oak (Quercus agrifolia) and willows (Salix spp.). Rock outcrops 

appear to be a particularly important habitat element for this snake in San Diego 

County (McGurty 1988).

Status and Threats 

The San Diego mountain kingsnake is a State of California Species of Special 

Concern (Jennings and Hayes 1994a) and a Forest Service Region 5 Sensitive Species 

Figure 6–Historic and current distribution of the San Diego mountain kingsnake.

Modified from Jennings and Hayes (1994a). 

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(Stephenson and Calcarone 1999). It is subject to the same anthropogenic impacts as 

the San Bernardino mountain kingsnake. Due to its tendency to display more red in

its color pattern, however, it may be even more highly prized by collectors. McGurty 

(1988) has noted a local population decline and examples of habitat destruction

(dismantling of rock outcrops and shredding of logs and stumps) in San Diego 

County. These activities continue despite California Department of Fish and Game 

and U. S. Forest Service regulations prohibiting such habitat alterations (California 

Freshwater Sport Fishing Regulations 2002, Jennings and Hayes 1994a). 

Management Recommendations 

All management recommendations made for the San Bernardino mountain kingsnake 

also apply to the San Diego subspecies and should be developed for the Cleveland 

National Forest. In addition, a program of systematic habitat monitoring and public 

education would be particularly important here to ensure that habitat impacts are 

minimized. 

California Red-legged Frog Rana aurora draytonii 

Baird and Girard 1852 

Description 

A large (adult snout-urostyle length 85-138 mm, Jennings and Hayes 1994a) frog

with a dorsal pattern of dark flecks and spots on a brown, gray, or reddish-brown

background color (fig. 7; Stebbins 1985). The distribution of red or red-orange pigment is highly variable, but usually restricted to the belly and undersurfaces of the 

thighs, legs, and feet (Jennings and Hayes 1994a). Frogs in southern California often 

have red only on the undersurfaces of the feet (Jennings pers. observ.). There are 

prominent dorsolateral folds that are yellow or orange in juveniles (Stebbins 1985).

The groin has a distinct black mottling on a white or yellow background. The iris is 

dark brown with iridophores on the upper and lower portions of the iris (Jennings and 

Hayes 1994a). Larvae (fig. 8) range from 14 to 80 mm long and have two or three 

upper and three or four lower rows of teeth (Stebbins 1985). Newly hatched tadpoles

generally are blackish in color, gradually changing to a brown background color with

darker marbling or spots after a week or two of growth (Storer 1925). 

Taxonomy 

Recent molecular phylogenetic analysis of red-legged frogs (Rana aurora) from 

across their range supports independent morphological (Hayes and Kremples 1986, 

Hayes and Miyamoto 1984), allozyme (Green 1985, Hayes and Miyamoto 1984), and

life history data (Hayes and Miyamoto 1984, Jennings and Hayes 1994a) suggesting 

that Rana aurora draytonii should be considered a full species (Bradley [sic] and

others 2000). Further, limited samples indicate that populations south of the Transverse Ranges form a clade distinct from those populations further north in coastal 

southern and central California (Bradley [sic] and others 2000). 

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Figure 7–Adult California red-legged frog from the Santa Rosa Plateau Preserve, 

Riverside County, California. Photo by Mark R. Jennings. 

Figure 8–Larval California red-legged frog from Corral Hollow, San Joaquin County,

California. Photo by Norman J. Scott, Jr.

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Distribution 

Historically, Rana aurora draytonii was found in Pacific slope drainages from the 

vicinity of central Mendocino County and Redding, Shasta County (Bradley [sic] and

others 2000, Storer 1925) south to the Santo Domingo River drainage in Baja California, Mexico (Linsdale 1932). Isolated populations occurred in four desert drainages in southern California (Jennings and Hayes 1994b). In the area covered by this 

paper, it was historically found in low to mid-elevations of much of the South Coast 

Ranges, Tehachapi Mountains, Greenhorn Mountains, Transverse Ranges, Peninsular 

Ranges, and South Coast Plain between near sea level (Gaviota Creek, Santa Barbara 

County) and 1,706 m (Crystal Lake in the San Gabriel Mountains, Los Angeles 

County) (fig. 9). Currently, red-legged frogs still are extant over much of the South 

Coast Ranges and in drainages north of the Santa Clara River in Los Angeles and 

Ventura Counties (Jennings and Hayes 1994a). However, south of the Santa Clara 

River, frog populations appear to have been extirpated, with only two small

populations still known in this area. All known populations in the Tehachapi and

Greenhorn Mountains now appear to be extinct (Jennings and Hayes 1994a). 

Habitat and Habits 

The California red-legged frog typically is found in or next to livestock ponds and

pools in perennial streams in southern California (Jennings pers. observ.). The most 

optimal habitat is characterized by dense, shrubby riparian vegetation associated with 

Figure 9—Historic and current distribution of the California red-legged frog in

southern California. Modified from Jennings and Hayes (1994b). 

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deep (>0.7 m), still, or slow-moving water (Hayes and Jennings 1988, Jennings

1988). The shrubby riparian vegetation that structurally seems to be most suitable for 

this frog is that provided by arroyo willow (Salix lasiolepis), although cattails (Typha

spp.) and bulrushes (Scirpus spp.) also can provide suitable habitat (Jennings 1988). 

Although California red-legged frogs are found in ephemeral streams and ponds, 

populations cannot be maintained where all surface water disappears (Jennings and 

Hayes 1994a). This frog is infrequent or absent in habitats where introduced aquatic 

predators such as green sunfish (Lepomis cyanellus), red-swamp crayfish (Procambarus clarkii), and bullfrogs (Rana catesbeiana) are present (Hayes and Jennings

1986, 1988), probably because its larval stages are susceptible to such predators 

(Jennings and Hayes 1994b). Subadult and adult California red-legged frogs often

include other anurans and small mammals in their diet (Hayes and Tennant 1986).

They are highly nocturnal and quite wary (Hayes and Tennant 1986, Storer 1925),

but juveniles tend to be much less wary and frequently are active diurnally (Hayes 

and Tennant 1986). This frog breeds from late November to April, and females attach

masses of 2,000 to 6,000 eggs to emergent vegetation in shallow water (Hayes and 

Jennings 1986, Jennings pers. observ., Storer 1925). Larvae require 4-5 months to 

metamorphose (Storer 1925) and apparently, under limited circumstances, will overwinter (Fellers and others 2001). 

Status and Threats 

The California red-legged frog is a State of California Species of Special Concern

(Jennings and Hayes 1994a) and was federally listed as a Threatened Species on 23

May 1996, based on a range reduction of approximately 70 percent and continued 

threats to remaining populations (U.S. Fish and Wildlife Service 1996). Frogs in

southern California have declined largely due to habitat loss and the introduction of

non-native aquatic predators such as green sunfish, red-swamp crayfish, and 

bullfrogs (Jennings and Hayes 1994b). It is possible that a pathogen also helped to 

eliminate frog populations in southern California during the 1970s (Fellers and others 

2001). Recent work suggests that nitrate/nitrite pollution (Marco and others 1999) 

and pesticide drift (Davidson and others 2001, 2002) also may be responsible for frog 

declines in California. Although California red-legged frogs still are present in some 

numbers in South Coast Range drainages north of the Santa Clara River, they are

virtually extinct south of this location (Jennings and Hayes 1994a). The population in

the Las Virgenes Creek drainage in extreme western Los Angeles County numbers 

fewer than 25 adults (R. Smith pers. comm.). The population at the Santa Rosa 

Ecological Preserve in western Riverside County was down to only two males in

2001 and no frogs were seen in 2002 (Jennings unpubl. data). Populations at both

locations are at risk due to surrounding development, non-native aquatic predators, 

and native predators such as the raccoon (Procyon lotor) (Jennings unpubl. data). 

Management Recommendations 

Surveys at several sites on the Cleveland National Forest have failed to reveal any 

California red-legged frog populations (D. Holland pers. comm.), but searches need 

to continue, focusing especially on private lands in southern California. Known 

populations need to be protected from threats such as urban development, flood

control projects, spread of known frog pathogens, increased raccoon populations, and 

introduced fishes and bullfrogs. Finally, the genetic diversity of extant populations 

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needs to be conserved, including breeding of captive populations, for possible 

reintroduction onto public lands with suitable remaining habitat. The largest pool of

frogs for this task appears to be confined to northern Baja California, Mexico. 

Mountain Yellow-legged Frog Rana muscosa Camp 1917 

Description 

A medium-sized (snout-urostyle length about 40-80 mm, Jennings and Hayes 1994a) 

frog with a variable dorsal pattern (fig. 10; Stebbins 1985). The markings usually

range from discrete dark spots that can be few and large, to smaller and more

numerous ones in a mixture of sizes and shapes, to irregular lichen-like patches, or a 

poorly defined reticulum (Zweifel 1955). Color also is highly variable, usually a

mixture of brown and yellow, but often with gray, red, or green-brown (Jennings

pers. observ.). The posterior half of the upper lip is weakly light-colored and

dorsolateral folds are present but normally not prominent (Stebbins 1985). The throat 

is white or yellow, sometimes with a mottling of dark pigment (Zweifel 1955). 

Undersurfaces of the legs and lower belly are pale yellow to sun yellow; the latter 

color usually present on the largest individuals (Jennings and Hayes 1994a). The iris

is gold with a horizontal, black counter-shading stripe (Jennings pers. observ.). 

Larvae (fig. 11) range in length from 10 to 65 mm long and have two to four upper 

and four lower rows of teeth (Stebbins 1985). Newly hatched tadpoles generally are 

Figure 10–Adult mountain yellow-legged frog from Vincent Gulch, San Gabriel

Mountains, Los Angeles County, California. Photo by Galen B. Rathbun. 

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blackish in color, gradually changing to a brassy background color with darker 

marbling and spots after a week or two of growth (Jennings pers. observ.). 

Taxonomy 

The mountain yellow-legged frog was originally described as two subspecies of Rana

boylii, one in the Sierra Nevada and the other in southern California (Camp 1917).

On the basis of morphological data, Zweifel (1955) separated these two forms from 

R. boylii and combined them as the distinct species R. muscosa. This arrangement 

was supported by subsequent genetic work (see Jennings and Hayes 1994a). Recent

mitochondrial DNA studies indicate that mountain yellow-legged frog populations in 

all southern California mountain ranges form a clade distinct from the populations in

the Sierra Nevada (Macey and others 2001). Further studies may suggest species

level recognition for these southern California frogs. 

Figure 11—Larval (2nd year) mountain yellow-legged frog from Little Rock Creek,

San Gabriel Mountains, Los Angeles County, California. Photo: Norman J. Scott, Jr. 

Distribution 

In southern California, the mountain yellow-legged frog historically was present in

most streams, watercourses, and natural lakes in the San Gabriel, San Bernardino, 

and San Jacinto Mountains, as well as on Mount Palomar in San Diego County (fig. 

12). At many locations prior to 1965, the frogs could be extremely abundant 

(Jennings and Hayes 1994b, Stewart pers. observ.). The elevation range was from 

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370 m at Eaton Canyon, Los Angeles County, to over 2,290 m near Bluff Lake, San 

Bernardino County (Jennings and Hayes 1994a, Zweifel 1955). Today, mountain 

yellow-legged frogs are largely restricted to small watercourses at elevations greater 

than 1,500 m (Jennings 1993, 1994). 

Habitat and Habits 

This species inhabits mid- to upper-elevation perennial streams, often in locations 

with bedrock pools (Jennings 1993, 1994; Jennings pers. observ., Mullally 1959). 

Adult frogs are diurnal and often observed sitting in shallow pools of water with their 

back exposed to direct sunlight. Although streams may be partly shaded by riparian 

vegetation such as white alder (Alnus rhombifolia), most streams that contain frog 

populations are relatively open (Jennings pers. observ.). While life history data for

southern California mountain yellow-legged frogs are very limited, the frogs appear 

to be strictly insectivorous and do not include vertebrates in their diet (Jennings and 

Hayes 1994a). In the late 1960s, Stewart (pers. observ.) found egg masses in April at

700 m elevation in Bear Creek in the San Gabriel River drainage. Oviposition 

probably occurs later at higher elevations. Egg masses typically are deposited in

shallow water and not attached to objects (Jennings and Hayes 1994a, Stewart pers. 

observ.). In the Sierra Nevada, larvae are known to overwinter and take between 1

and 2.5 years to metamorphose (Jennings and Hayes 1994a). Overwintering of larvae 

has been observed at all locations where this species still occurs in southern 

California (Jennings pers. observ.). 

Figure 12—Historic and current distribution of the mountain yellow-legged frog in

southern California. Modified from Jennings and Hayes (1994a). 

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Status and Threats 

The mountain yellow-legged frog is a State of California Species of Special Concern 

(Jennings and Hayes 1994a) and a Forest Service Region 5 Sensitive Species 

(Stephenson and Calcarone 1999). It is believed to have been extirpated from 

approximately 99 percent of its historic range in southern California (Jennings 1995a, 

Jennings and Hayes 1994a), and the southern California population was federally

listed as an endangered species on 2 July 2002 (U.S. Fish and Wildlife Service 

2002). In the San Gabriel Mountains, small populations of frogs exist only in the

upper reaches of Little Rock Creek, Big Rock Creek, Devil's Canyon, and the East 

Fork of the San Gabriel River (Jennings 1993, 1994, 1995b, 1998, 1999; Jennings

and Hayes 1994b). In the San Bernardino Mountains, a small population currently

exists in a short section of the lower portion of the East Fork of City Creek (Jennings

pers. observ., U.S. Fish and Wildlife Service 2002). In the San Jacinto Mountains, 

small populations of frogs exist only in the upper reaches of the North Fork of the

San Jacinto River, Hall Canyon, Dark Canyon, and Fuller Mill Creek (Jennings 1998, 

1999). Current estimates put the total number of adult frogs in the nine known 

populations at fewer than 200 individuals (Jennings 1995a). 

All of these populations are extremely vulnerable to extirpation via

recreation (U.S. Fish and Wildlife Service 1999), especially when such activities 

occur on a repeated basis year after year (U.S. Fish and Wildlife Service 2002). For 

example, the second author has witnessed the children of rock climbers chasing frogs

and adults disturbing sunning frogs in the Little Rock Creek drainage; adults hiking 

directly in and up the streams of Little Rock Creek and the North Fork of the San

Jacinto River, disturbing the larval and metamorphosing frogs present; a mother and

daughter photographing flowers in the North Fork of the San Jacinto River and 

disturbing a small population of metamorphosing frogs; and adults and children

swimming (and associated litter) in Dark Canyon, Mill Creek, and the East Fork of 

the San Gabriel River where larvae and subadult frogs were present. Such activities 

are not limited to areas near roads. For instance, one individual reported hiking 

directly up a stream containing a known frog population on Mount San Jacinto 

because he “just wanted to get away from the crowds.” 

Other recreational activities which negatively effect frog populations include 

fishing for trout (Oncorhynchus mykiss and Salmo trutta) at almost all known frog 

locations and camping at locations along the North Fork of the San Jacinto River and 

East Fork of the San Gabriel River where frog populations are still extant. All of 

these recreational activities greatly increase the chances of pathogen introduction (via

shoes, clothing, and fishing equipment) that could decimate the small, isolated populations. Continuing requests for water removal for ski operations in the uppermost 

reaches of Little Rock Creek also threaten the small frog population there through

potential dewatering of the stream (U.S. Fish and Wildlife Service 1999, 2002). 

Finally, the negative effects of chance catastrophic events, such as fires, 

droughts and floods, also greatly endanger these localized populations (U.S. Fish and 

Wildlife Service 2002). The 1968-69 floods in southern California decimated many

of the frog populations that formerly were abundant in the region (Jennings and 

Hayes 1994a, 1994b). 

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Management Recommendations 

There needs to be a long-term, comprehensive program to determine and monitor the 

number of adults, juveniles, and egg masses or larvae for each of the remaining 

populations. This will require careful yearly surveys. Otherwise, most of these 

populations could become extinct without us knowing it. Additionally, campgrounds 

in sensitive frog breeding sites need to be closed for a portion of the year or removed. 

Also, trails near sensitive frog habitats need to be relocated well away from riparian

areas. If possible, volunteers should be used to keep recreationists out of the most

sensitive frog breeding habitats throughout the summer. Trout stocking should be 

discontinued in frog habitat, and existing trout populations should be removed. 

Finally, a captive breeding population of frogs from each southern California 

mountain range needs to be started to assure that the genetic diversity of this species 

is preserved for potential reintroduction programs. 

San Gabriel Mountain Slender Salamander

Batrachoseps gabrieli Wake 1996 

Description 

A slender salamander of medium size (snout-vent length 40-50 mm, Wake 1996)

with a relatively broad, flattened head that is well demarcated from the neck (fig. 13; 

Wake 1996). It also has long limbs, large feet, a tapered tail, and 18 or 19 costal 

grooves between the front and hind limbs (Wake 1996). With a complete interlacing 

of melanophores, the ground color generally is black in this species throughout its 

range. Individuals from the western part of the range (San Gabriel Canyon, San

Gabriel Mountains, Los Angeles County) have a dorsum overlain with a dense layer 

of coppery iridophores that are more concentrated around the shoulders and pelvic 

region, forming two short, broad stripes. Salamanders from the central part of the 

range (San Antonio and Lytle Canyons, San Gabriel Mountains, Los Angeles and

San Bernardino Counties) usually have white-to-silvery iridophores in addition to the 

coppery iridophores. Specimens from the eastern part of the range (Devil and 

Waterman Canyons, San Bernardino Mountains, San Bernardino County) typically

are devoid of any coppery iridophores, but have the white-to-silvery iridophores.

Taxonomy 

Batrachoseps gabrieli is a recently described salamander known only from the 

cismontane side of the San Gabriel and San Bernardino Mountains in southern 

California (Goodman and others 1998, Hansen and others 2005, Wake 1996). Wake 

(1996) notes that this species is morphologically and genetically differentiated from 

all other Batrachoseps, with separation from other lineages occurring approximately

8-13 million years ago. 

Distribution 

The range (fig. 14) of B. gabrieli is poorly known. Originally, this species was known

from only two locations within the San Gabriel Mountains (Wake 1996), but it 

recently has been collected and/or observed at 11 other sites within the San Gabriel 

184 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

and San Bernardino Mountains (Goodman and others 1998, Goodman pers. observ.). 

Figure 13—Two adult San Gabriel Mountain slender salamanders. The animal on

the left is from Waterman Canyon, San Bernardino Mountains, San Bernardino

County, California and the animal on the right is from Alpine Canyon, San Gabriel

Mountains, Los Angeles County, California. Photo by Robert H. Goodman, Jr. 

The distribution apparently is discontinuous, with populations restricted to narrow 

swaths of suitable habitat in separate drainages, bounded on the west by San Gabriel 

Canyon and on the east by Waterman Canyon. All specimens have been collected in

either the Angeles or San Bernardino National Forests. The elevation range is from 

approximately 850 m to 2,380 m.

Habitat and Habits 

Habitat associations for the new populations of B. gabrieli (Goodman and others 

1998) are similar to those described by Wake (1996) for the type locality, although 

generally lower in elevation. Most individuals have been observed in and around

talus, under rotting logs, bark, downed branches, downed fern fronds and rocks 

(Hansen and others 2005). Vegetation components present within and around the

talus include canyon live oak (Quercus chrysolepis), bigcone Douglas-fir, Jeffrey

pine, yellow pine, and chaparral components such as ceanothus (Ceanothus spp.), 

manzanita (Arctostaphylos spp.), mountain mahogany (Cercocarpus betuloides) and 

chaparral yucca (Yucca whipplei). Surface activity appears to peak during southern 

California’s winter snow/rain season; individuals have been observed while snow 

was present (Goodman pers. observ., Wake 1996,). Three other plethodontid

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salamanders—the Monterey salamander (Ensatina eschscholtzii eschscholtzii), blackbellied slender salamander (Batrachoseps nigriventris), and arboreal salamander 

(Aneides lugubris)—are known to occur sympatrically with B. gabrieli (Goodman 

pers. observ., Wake 1996).

Figure 14—Current distribution of the San Gabriel Mountain slender salamander. 

Status and Threats 

The San Gabriel Mountain slender salamander currently has no State or Federal 

status, but it is a Forest Service Region 5 Sensitive Species (Stephenson and

Calcarone 1999). Fortunately, all populations of B. gabrieli occur within national 

forest boundaries, and because it occurs within and around talus slopes and seems to

be surface active only during the winter months, there appear to be no direct threats

to it at this time. 

Management Recommendations 

Since this salamander currently is known from only 13 locations within the central to 

eastern San Gabriel Mountains and western San Bernardino Mountains, we 

recommend that it be listed by the State of California as a Species of Special 

Concern. For management purposes, and to obtain a better estimate of its range,

further surveys should be conducted throughout the San Gabriel, San Bernardino, and

San Jacinto Mountains to locate new populations. In addition, management

guidelines should be developed to protect the talus habitat of this species. 

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Yellow-blotched Salamander Ensatina eschscholtzii 

croceater (Cope 1869) 

Description 

A medium-sized (snout-vent length 48-78 mm, Jennings and Hayes 1994a) 

salamander with large, irregular, pale, lemon-yellow to yellowish cream blotches on 

a deep blackish brown to black ground color (fig. 15; Stebbins 1985). As in other 

forms of Ensatina, there are 12-13 costal grooves and a noticeable constriction at the 

base of the tail (Stebbins 1949, 1985). 

Figure 15—Adult female yellow-blotched salamander from San Emigdio Mountain, 

Kern County, California. Photo by Robert H. Goodman, Jr. 

Taxonomy 

The ensatina (Ensatina eschscholtzii) has long been considered a classic example of a 

“ring species” (Highton 1998, Wake and Schneider 1998). Seven named subspecies 

are distributed around the Central Valley of California, apparently intergrading in

contact areas, but overlapping without interbreeding at the ends of the ring (E.e. 

eschscholtzii and E.e. klauberi) in southern California (Stebbins 1949). Behaving as 

closely related but distinct species, unblotched eschscholtzii and blotched klauberi

are known to hybridize occasionally (Brown 1974, Wake and Yanev 1986), and 

Highton (1998) advocated species-level recognition with hybrid zones for all the infraspecific taxa within Ensatina. However, Wake and Schneider (1998), while acUSDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 187 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

knowledging that the ensatina does not meet all of the criteria for a true ring species, 

argued against disrupting the long-stable taxonomy of this species complex until 

further morphological and genetic studies are completed. We follow that policy here. 

Distribution 

The current known range of the yellow-blotched salamander essentially is restricted

to east-central Kern and northern Ventura Counties, extending from the Piute 

Mountains in the Sequoia National Forest southwestward to the vicinity of Alamo 

Mountain in the Los Padres National Forest (fig. 16). Northwest of the Piute

Mountains, ensatina populations appear to be intergrades between E.e. croceater and 

E.e. platensis. Two hundred km east of Alamo Mountain, blotched ensatina collected 

in the San Bernardino Mountains look like intergrades between E.e. croceater and

E.e. klauberi (Stebbins 1949; fig. 17), with specimens from some localities looking 

more like one or the other of the two taxa (compare figs. 16 and 19). However, all are

genetically closer to the latter (Wake and Schneider 1998). Interestingly, blotched 

ensatina seem to be absent from the intervening San Gabriel Mountains, an area 

known as “Bob’s Gap” in reference to Robert Cyril Stebbins who did much of the 

early work with the species complex. E.e. croceater occurs from about 427 m to 

2,285 m in elevation (Jennings and Hayes 1994a). 

Figure 16—Historic and current distribution of the yellow-blotched salamander.

Intergrades indicated in the San Bernardino Mountains look more like E.e. croceater

than E.e. klauberi. Modified from Jennings and Hayes (1994a). 

188 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
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Habitat and Habits 

The yellow-blotched salamander occurs in a wide variety of vegetation associations 

including oak (Quercus douglasii and Q. kelloggii) woodlands, pine (Pinus jeffreyi

and P. ponderosa) and fir (Abies concolor) forests, and open chaparral (Jennings and 

Hayes 1994a, Stephenson and Calcarone 1999). It typically is found under downed

logs, leaf litter and duff, woody debris, and medium-to-large rocks (Goodman pers. 

observ., Jennings and Hayes 1994a). As with most other southern California 

salamanders, the habits of this subspecies are poorly known, although surface activity

generally peaks during the winter months. Gravid females have been observed in

April and May (Jennings and Hayes 1994a). 

Status and Threats 

This salamander is a State of California Species of Special Concern (Jennings and 

Hayes 1994a) and a Forest Service Region 5 Sensitive Species (Stephenson and

Calcarone 1999). Although it is more widespread and abundant than originally 

realized, there are concerns about increased development in the Tehachapi

Mountains, Cummings Valley, and Bear Valley areas and increased wood harvesting 

on the Tejon Ranch (Jennings and Hayes 1994a). These activities occur primarily in

oak woodlands and may have a direct impact on known and undiscovered salamander 

populations. 

Figure 17—Subadult intergrade between yellow-blotched and large-blotched

salamanders from Crystal Creek, San Bernardino Mountains, San Bernardino

County, California. Photo by Robert H. Goodman, Jr. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 189 
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Management Recommendations 

Systematic surveys and habitat assessments should be conducted within known

populations of the yellow-blotched salamander in order to gain a better understanding

of this taxon's natural history. Similar work might then be able to locate populations

in areas like the San Gabriel Mountains. If specimens are collected in “Bob’s Gap,”

they may yield significant information on the phylogeography of the ensatina species

complex (Wake and Schneider 1998). While E.e. croceater may be more directly

impacted by habitat degradation on private lands than by prevailing land use 

activities on public lands (Stephenson and Calcarone 1999), we believe that 

guidelines should be developed by the Sequoia and Los Padres National Forests to 

prevent over-harvesting of standing trees, downed logs, and wood debris. 

Large-blotched Salamander Ensatina eschscholtzii 

klauberi Dunn 1929 

Description 

A medium-sized (snout-vent length 45-82 mm, Jennings and Hayes 1994a) 

salamander with large, rectilinear, orange blotches (sometimes spots or bands) on a 

deep blackish-brown-to-black ground color (fig. 18; Stebbins 1985).

Figure 18—Subadult large-blotched salamander from near Julian, Cuyamaca

Mountain, San Diego County, California. Photo by Robert H. Goodman, Jr. 

190 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

Taxonomy 

The taxonomic comments made for the yellow-blotched salamander also apply to this

subspecies. 

Distribution 

The known range of the large-blotched salamander is from the San Jacinto 

Mountains, Riverside County to Cottonwood Creek, San Diego County (fig. 19). Its

distribution is discontinuous, and as, previously noted, possible intergrades between 

E.e. klauberi and E.e. croceater occur in the San Bernardino Mountains (Stebbins 

1949; fig. 17). Recently, E.e. klauberi has been observed in the Sierra San Pedro 

Martir, Baja California Norte, Mexico (Mahrdt and others 1998). The elevation range 

for this salamander is from about 518 m to 1,646 m (Jennings and Hayes 1994a). 

Figure 19—Historic and current distribution of the large-blotched salamander in

southern California. Intergrades indicated in the San Bernardino Mountains look more 

like E.e. klauberi than E.e. croceater. Modified from Jennings and Hayes (1994a). 

Habitat and Habits 

Vegetation associations of this subspecies are similar to those of the yellow-blotched 

salamander, but the oaks are more likely to be Quercus chrysolepis and Q. kelloggii, 

and the pines Pinus coulteri and P. ponderosa, along with incense cedar (Jennings

and Hayes 1994a, Stephenson and Calcarone 1999). Again, the salamanders typically

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Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

are found under various kinds of surface litter and rocks (Goodman pers. observ., 

Jennings and Hayes 1994a), and surface activity generally peaks when moisture is 

present. 

Status and Threats 

The large-blotched salamander is a State of California Species of Special Concern

(Jennings and Hayes 1994a) and a Forest Service Region 5 Sensitive Species 

(Stephenson and Calcarone 1999). Even though it also is considerably more common

than was formerly believed (Jennings and Hayes 1994a), threats from habitat 

destruction are increasing. As with the yellow-blotched salamander, the largeblotched salamander may be more impacted by habitat degradation resulting from 

development on private lands than land use activities on public lands (Stephenson

and Calcarone 1999). However, potential impacts to populations from mining exist in 

the Crystal Creek area of the San Bernardino Mountains (Goodman pers. observ.,

Jennings and Hayes 1994a).

Management Recommendations 

Recommendations discussed with respect to E.e. croceater are applicable here. 

According to D. Wake (pers. comm.), known populations of blotched salamanders 

from the San Bernardino Mountains are few and disjunct, and more surveys need to 

be done to locate additional populations for study. To maintain habitat quality for 

ensatinas, the Cleveland and San Bernardino National Forests should develop

guidelines to prevent over-harvesting of trees, logs, and wood debris. 

Conclusion 

As described in the foregoing accounts, all of these taxa face varying degrees of

threat—in some cases, severe threat—from a multitude of anthropogenic factors 

prevalent in southern California. Lest there be any doubt that these factors can result

in extinctions here, the case of the foothill yellow-legged frog (Rana boylii) must be 

mentioned. In the south coast area, this frog historically occurred below the conifer 

forest habitats in parts of the South Coast Ranges and San Gabriel Mountains from 

near sea level (Gaviota Creek, Santa Barbara County) to 1,219 m (Indian Creek in the

San Rafael Mountains, Santa Barbara County). It has not been observed in this area 

since before 1978 (H. DeLisle, M. Long, and S. Sweet pers. comm.; Jennings and 

Stewart pers. observ.), and all of these populations are now believed to be extinct

(Jennings 1995a, Jennings and Hayes 1994a). 

With their relatively low vagility, all of the taxa discussed in this paper face 

threats from habitat degradation, fragmentation, and loss. These may lead directly to 

extinction in the short term for those already reduced to very small, isolated

populations or to a more gradual extinction process involving local extirpations, 

restricted gene flow, reduced genetic diversity, and increased vulnerability to 

environmental changes. To the extent that global warming scenarios prove to be 

correct, long-term climate change may become a serious problem for amphibians,

and the taxa may be restricted to montane habitats that already are essentially insular

relicts. 

192 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Snakes, Frogs, and Salamanders—Stewart, Jennings, and Goodman 

As specialists in herpetology, we want to emphasize that the taxa discussed here

are the subjects of recent and ongoing phylogeographic studies using the most 

advanced techniques of phyletic and genetic analysis. Consequently, their taxonomy

is in a state of flux. More importantly, however, they are of continuing interest to 

biologists studying the evolutionary processes that shape modern species of terrestrial

vertebrates. 

To prevent further declines and losses in the herpetofauna of the southern

California mountains, additional research must be conducted to increase our

knowledge of the distribution, habitat needs, life history, and phylogenetic 

relationships of the taxa occurring there. Collaboration between the scientific 

community, the private sector, and public agencies at the Federal, State and local 

levels will be essential to obtaining the necessary data and incorporating them into

sound management programs. These programs must be monitored regularly to ensure 

that desired objectives are being met. Only then can we expect that the montane

components of southern California's rich biodiversity will be maintained for

continuing scientific study and public enjoyment. 

Acknowledgments 

We thank Harold F. DeLisle, Michael C. Long, and Samuel S. Sweet for providing 

unpublished field notes and other pertinent information. Galen B. Rathbun and 

Norman J. Scott, Jr. generously provided some of the photographs used in this paper. 

Timothy A. Shoemaker kindly assisted in formatting the manuscript. Portions of this 

work were funded by grants provided by the California Department Fish and Game, 

the U.S. Department of Agriculture Forest Service, and the U.S. Geological Survey. 

Finally, we thank the organizers of this symposium for inviting us to participate and 

the reviewers of our manuscript for their helpful comments. 

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USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 197 

The Spotted Owl in Southern California: 

Ecology and Special Concerns for 

Maintaining a Forest-Dwelling Species in a 

Human-Dominated Desert Landscape1

William S. LaHaye and R. J. Gutiérrez2

Abstract 

The California spotted owl is an uncommon forest-dwelling resident of southern California, 

found in most of the major mountain ranges in the region. We studied this species from 1987 

to 1998 in the San Bernardino Mountains and collected empirical and demographic evidence

that the population declined during this period. Numerous factors may affect the long-term 

viability of spotted owls in this region. Key factors include a naturally fragmented

distribution, loss and degradation of nesting habitat, loss and degradation of dispersal habitat,

severe drought, and (potentially) air pollution. Our insight into the persistence of this owl is 

limited because we do not know the cause of the population decline we observed in the San 

Bernardino Mountains. However, it is likely that the decline has resulted from a complex

interaction among weather, habitat changes, prey dynamics, and other factors. We provide a

regional update on this owl and discuss the problems that may be affecting the species in

southern California. 

Key words: California spotted owl, habitat degradation, population decline, Strix occidentalis

occidentalis

Introduction

The spotted owl (Strix occidentalis) ranges from southwestern British Columbia into 

Baja California Norte along the Pacific Coast. In addition, its distribution includes 

the southern Rocky Mountains from Utah and Colorado into the mountains of 

Mexico, where it reaches its southernmost occurrence near Mexico City (Gutiérrez

and others 1995). It primarily occurs in mountainous terrain where it typically

occupies closed-canopied forests (Gutiérrez and others 1995). However, the spotted 

owl also occupies steep, narrow canyons with relatively few trees in some portions of

the southwestern United States (Rinkevich and Gutiérrez 1996). 

The American Ornithologists’ Union (1957) recognizes three subspecies: The 

northern spotted owl (S. o. caurina) is found in the Pacific Northwest and 

northwestern California, the California spotted owl (S. o. occidentalis) occurs in most

California mountains other than those occupied by the northern subspecies, and the 

Mexican spotted owl (S. o. lucida) ranges from southern Utah and Colorado into 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 Research fellow, Department of Fisheries, Wildlife, and Conservation Biology, University of 

Minnesota, St. Paul, MN 55108; current address: P. O. Box 523, Big Bear City, CA 92314 (email: 

blahaye@gte.net); and Professor and Gordon Gullion Endowed Chair, Department of Fisheries, 

Wildlife, and Conservation Biology, University of Minnesota, St. Paul, MN 55108

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 199 
Spotted Owls in Southern California—LaHaye and Gutiérrez

Mexico (Gutiérrez and others 1995). The northern and Mexican subspecies are listed

as “threatened” under the Endangered Species Act (U.S. Fish and Wildlife Service 

1990, 1993) because of habitat loss and failure of existing regulatory mechanisms to

protect the owl. 

In southern California, forested habitat is restricted to the upper elevations of the 

higher mountain ranges and in canyon bottoms along streams and rivers. Thus, 

suitable spotted owl habitat is discontinuous and results in a series of semi-isolated

owl populations within the region (table 1; LaHaye and others 1994, Noon and

McKelvey 1992). The largest of these populations, located in the San Bernardino 

Mountains, declined between 1987 and 1998 (table 2; LaHaye and others 1999).

Table 1—The estimated number of territories within each subpopulation of the southern

California spotted owl metapopulation. Data from LaHaye and others (1994).

 

Estimated

number of

Subpopulation territories

Northern Monterey 50

Southern Monterey 30

Cerro Alto 3 

Southern Santa Lucia 16

Sierra Madre 12

San Rafael 40 

Santa Ynez 18

Mt. Pinos 15

Tecuya 10

Cobblestone 10

Tehachapi Mountains 12

Pelano 20

San Gabriel Mountains 95

San Bernardino Mountains 133 

Santa Ana Mountains 12

San Jacinto Mountains 20

Thomas Mountain 10

Palomar Mountain 30

Black Mountain 5 

Volcan Mountain 16

Cuyamaca Mountain 10 

Laguna Mountain 15

The California Spotted Owl Technical Assessment Team (CASPO) evaluated 

the California spotted owl and presented “Areas of Concern” and

“Recommendations” for maintenance of this subspecies in southern California (see 

Verner and others 1992: 7, 16). The purpose of this paper is to review these issues in

the context of new information and to qualitatively evaluate the progress toward

maintenance of this subspecies in southern California since CASPO. 

200 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Spotted Owls in Southern California—LaHaye and Gutiérrez

Spotted Owl Ecology

Spotted owls are found in riparian/hardwood forest, live oak (Quercus

chrysolepis)/bigcone Douglas-fir (Pseudotsuga macrocarpa) forest, and mixed

conifer forest in southern California (Gutiérrez and others 1992). About one third of

the known spotted owl sites are located in each forest type (Gutiérrez and others 

1992). Spotted owl reproduction was found to be higher in live oak/bigcone Douglasfir forest than in mixed conifer forest (LaHaye and others 1997); however, 

reproduction in riparian/hardwood sites was not considered. We speculate that owl 

reproduction in riparian/hardwood sites is similar to live oak/bigcone Douglas-fir 

sites because these forest types are similar in elevation and vegetation association 

(that is, they are typically surrounded by chaparral). 

Table 2—Spotted owl survey effort and banding activity between 1987 and 1998 in the San

Bernardino Mountains of southern California. Table from LaHaye and others (2001). 

Territories Territorial

owls

Juvenile

owls

Number Pct Number Number Pct Number

Year surveyed occupied pairs singles banded1 banded

1987 42 95 32 8 0 15 

1988 78 95 56 18 42 16 

1989 115 90 81 23 42 52

1990 127 86 90 19 72 47

1991 132 76 86 12 79 44

1992 136 65 78 11 91 49

1993 141 63 78 11 90 52

1994 143 68 84 13 86 70

1995 145 70 86 15 93 38

1996 146 67 89 9 93 54 

1997 147 62 80 11 96 41

1998 148 59 76 11 95 23 1

 Percent of the territorial population banded prior to each field season, including newly recruited

territorial owls. 

Spotted owls prey on a variety of insects, birds, and mammals (Smith and others 

1999). The dusky-footed woodrat is their primary prey and appears to be important to

reproductive success in this owl (Smith and others 1999). Spotted owls in southern 

California rarely occur above approximately 2,400 m elevation (Gutiérrez and others 

1995). This is most likely due to the decline in woodrat density at higher elevations 

(Grinnell 1908). 

Little information has been collected on home-range sizes of spotted owls in

southern California. However, reported home ranges in mixed-conifer habitat varied

in size from approximately 400 hectares to 3,300 hectares (Zimmerman and others 

2001). We suspect that these values are near the extremes for this species in this 

habitat type. Some lower elevation owl sites may be smaller than 400 hectares

because suitable habitat is restricted to narrow strips of riparian vegetation. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 201 
Spotted Owls in Southern California—LaHaye and Gutiérrez

Areas of Concern 

CASPO identified a set of potential threats (see Areas of Concern, Verner and others 

1992: 7) that we used as the outline for our qualitative assessment. 

Natural Isolation of Habitat and the Regionally Restricted

Distribution of Spotted Owls

Areas capable of supporting the forested habitat used by this species are uncommon 

in southern California because forest is restricted to higher elevations. Thus, 

opportunities to expand the amount of spotted owl habitat are limited (Verner and 

others 1992). In addition, the continued urbanization of the southern California 

lowlands may exacerbate the difficulties of owl movement between mountain ranges

by creating additional barriers to dispersal (see below). 

Dispersal Habitat 

Spotted owl dispersal habitat includes the forest types mentioned above, as well as a 

variety of other forested and nonforested habitats (Gutiérrez and others 1985).

Dispersal may be necessary to maintain populations of spotted owls within the 

region, and it is potentially critical to the continued existence of several of the 

smallest populations (LaHaye and others 1994, Noon and McKelvey 1992). 

However, no inter-mountain movements were recorded between 1987 and 1998

(LaHaye and others 1994, 2001). We do not know whether movements between 

mountain ranges have always been low or if this is a recent phenomenon reflecting

reduced opportunities for successful dispersal caused by modifications to intermountain vegetation by humans. 

Numerous riparian areas, which once existed at lower elevations and were

potential dispersal corridors, have been degraded by water extraction or lost to 

channelization during the last century. In addition, numerous small coast live oak (Q.

agrifolia) stands have been eliminated or modified by urbanization and are no longer

usable by spotted owls. Some of these live oak stands were known to occasionally

support nesting pairs (Heller 1893, Lillie 1891, WFVZ #21,133, WFVZ #69,985,

WFVZ #84,150, WFVZ #121,717) and may have served as stepping-stones for 

dispersal among the region’s mountain ranges. 

Habitat Trends 

Most montane forests in southern California are administered by the USDA Forest 

Service. However, substantial acreage of forest also exists on private lands. CASPO

was unsure of the trend in the amount of montane forest in this region, and we retain 

that uncertainty today. Similar concerns were expressed about the integrity of owl 

habitat (for example, quality and fragmentation, Verner and others 1992). 

Historic logging, which occurred during the 19th century and from 1950 to 1980 

(Johanneck 1975, Robinson 1989), undoubtedly reduced the amount and quality of 

owl habitat in several mountain ranges in the region. Recent (to 2001) logging 

activity has been limited to infrequent commercial timber sales and localized

fuelwood extraction. Although, spotted owls have reoccupied some logged sites, the 

current suitability of previously logged forest and the extent of reoccupation of these 

202 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Spotted Owls in Southern California—LaHaye and Gutiérrez

forests by spotted owls are unknown. No analysis has been conducted on the effect of 

fragmentation on spotted owls in southern California (Franklin and Gutiérrez 2002). 

Urbanization in the form of primary and vacation homes has degraded or 

consumed some forest in most mountain ranges. The results of spotted owl surveys 

conducted between 1987 and 1998 in the San Bernardino Mountains indicated that a 

large area of potentially-suitable spotted owl habitat, enough to support 10-15 pairs, 

existed between Running Springs and Crestline (LaHaye and others 1999). However,

only four pairs have been found in this area, and owls were found only in 

undeveloped sites. Thus, residential development within montane forests may

preclude spotted owl occupancy, even when closed-canopy forest remains on 

developed sites. 

Forests in southern California have been subjected to major changes in 

disturbance regime during the past 150 years. Logging, urbanization, water 

diversions, wildfire, and mining have produced dramatic local impacts. Intensive 

grazing during the late 1800s (Minnich 1988), fire suppression since 1900 (McBride

and Laven 1976), and air pollution since the 1930s (Miller and McBride 1998) have

affected most forests in the region. These impacts have altered the growth and 

regeneration of forest vegetation (Minnich and others 1995) within owl habitat and 

could be influencing the dynamics of spotted owl prey. While we are unsure how 

these changes have affected this owl, if they negatively affect prey abundance or

availability, maintaining a regional population of spotted owls will require thoughtful

planning and management. 

Water Diversion and Stream Channelization 

Spotted owl habitat associated with riparian areas is subject to direct effects such

as removal of riparian forests due to channelization and other land-clearing activities 

and indirect effects such as vegetation degradation caused by water extraction and 

human recreation. Both may ultimately result in loss of suitable habitat. Riparian 

forests and other woodlands at low elevations have been modified at a faster rate than 

montane forests because they occur in areas targeted for development, flood control, 

or domestic water production. However, the extent of this habitat decline and the 

proportion of the southern California spotted owl population that historically resided

in these forests are unknown. 

Wildfire 

Wildfire is a constant threat in southern California (Keeley and others 1999,

Minnich 1983), and the effects of wildfires are becoming more severe (that is, fires

are larger and hotter, Minnich 1983). In addition, Minnich (1980) suggested that the 

total area of live oak/bigcone Douglas-fir forest has declined during the last century

due to changes in fire regime. Stand-replacing fires may become more commonplace. 

Most resource agencies recognize this problem, but public concerns about air quality,

safety, and liability often prevent application of large-scale prescribed fire programs 

(Biswell 1989). Thus, managers recognize the threat of fire to wildlife habitat and

humans, but little progress has been made toward implementing effective fuels 

reduction programs through controlled burning and other methods. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 203 
Spotted Owls in Southern California—LaHaye and Gutiérrez

Recreational Activities 

Human recreation may be a contributing factor to spotted owl habitat degradation. 

With approximately 15 million people in the southern California area, the demand for 

recreational opportunities is constantly increasing. Much of the focus for this demand 

is on public lands in the mountains and riparian areas. Direct impacts, such as killing 

owls, are probably uncommon. However, indirect impacts such as trampling

vegetation and habitat lost to development of new parks, campgrounds, trails, and 

roads may have significant cumulative effects on spotted owl habitat. We are 

unaware of a region-wide assessment of these impacts. 

Drought 

To the areas of concern listed by CASPO, we add the negative effects of drought and 

air pollution (see below). It is well known that weather can affect demographic

processes in spotted owls (for example, Franklin and others 2000, North and others 

2000). We hypothesize that drought can also have negative long-term effects on their 

habitat. 

The majority of the demographic data for the San Bernardino Mountains 

spotted owl population were collected when annual precipitation was below average.

The results of demographic analyses showed that reproduction was strongly 

dependent on precipitation patterns (LaHaye and others unpublished data3

). However, 

survival was not related to weather. In addition to the direct effect of weather on 

reproduction, the current extended drought is degrading montane forest in many areas 

(that is, increased tree mortality in forests occupied by owls) due to water stress,

which will likely alter the quality and distribution of spotted owl habitat in the region. 

Air Pollution 

Air pollution affects large areas of southern California, especially on the cismontane

sides of the major mountain ranges (Miller and McBride 1998). Pollutants such as 

ozone and nitrogen oxides may affect vegetation growth and subsequent densities of

spotted owl prey. In addition, there may be direct effects on the portion of the owl 

populations that receive the highest concentrations of air contaminants, because birds

lack a DNA repair mechanism for lung tissue (Rombout and others 1991). 

Recommendations 

Below, we review CASPO’s recommendations for the spotted owl in southern

California (Verner and others 1992: 16) and provide additional evaluation and

comment when appropriate. 

Spotted Owl Inventories 

CASPO’s first recommendation was to inventory all potential habitats in southern 

California to provide an accurate estimate of the size of each population. These

 

3

 Unpublished data on file at the Department of Fisheries, Wildlife, and Conservation Biology,

University of Minnesota, St. Paul, MN 

204 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Spotted Owls in Southern California—LaHaye and Gutiérrez

estimates are essential and have considerable influence on the results of modeling 

subpopulation and metapopulation persistence (LaHaye and others 1994, Noon and

McKelvey 1992). Although thorough inventories have not been completed,

Stephenson (1991) estimated the amount of suitable habitat available within all of the 

major mountain ranges and provided estimates of the potential number of owls that

each mountain range could support. Given the recently acquired information on 

demographic variability associated with this species (Franklin and others 2000, 

Seamans and others 2002), we suspect that some of the smaller owl populations may

not currently exist (see table 1). Indeed, the San Bernardino Mountains and San 

Jacinto Mountains spotted owl populations seem to have declined substantially

during the past decade based on recent surveys (LaHaye and others 1999, Gutiérrez 

and others unpublished data4

, respectively). 

Monitoring Demography 

CASPO also felt it was essential to monitor the demography of the owl throughout

the region. When their evaluation was released there were two spotted owl 

demography projects in progress (Gutiérrez and Pritchard 1990, LaHaye and others 

1992). CASPO recommended that two additional demographic projects be initiated to 

estimate correlation of population dynamics and to provide estimates of movement 

rates among subpopulations. These new projects were not initiated. Thus, estimates 

of these parameters are still unavailable. 

In addition, CASPO recommended continued funding for the existing 

demography studies. However, funding was discontinued in 1994 for the San Jacinto

Mountains’ population study and in 1999 for the San Bernardino Mountains’ owl 

study, despite evidence that both populations were declining. 

Maintain Habitat for all Known Pairs 

CASPO argued that the overall numbers of spotted owls in southern California and 

the small sizes of many of the known populations required maintenance of habitat for 

all known pairs. This recommendation was generally accepted and implemented by 

the USDA Forest Service (USDA Forest Service 1993). However, due to incomplete 

surveys and private ownership of some areas containing spotted owl habitat, it is not 

clear how effective this policy has been. A small number of spotted owls were known

or suspected to exist on private lands in several of these mountain ranges. We could 

not determine the degree of protection afforded these owls. To the best of our 

knowledge, the San Bernardino National Forest was the only national forest in 

southern California to develop a formalized plan for protecting spotted owl habitat 

(USDA Forest Service 1989). This plan described the criteria by which a site 

qualified as a designated owl territory, the amount of protected forest to be included

within a designated territory, the desired characteristics of forest to be protected, the

potential disturbance factors, and seasonal closures designed to minimize disturbance

to nesting owls. 

 

4 Unpublished data on file at the Department of Fisheries, Wildlife, and Conservation Biology,

University of Minnesota, St. Paul, MN

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 205 
Spotted Owls in Southern California—LaHaye and Gutiérrez

Formulating Regional Guidelines 

CASPO believed that it was necessary to develop southern California regional

guidelines for effective protection of this species. They recommended assembling a 

team of specialists to formulate these guidelines. This recommendation has not been

accomplished as of 2002. (Editor’s note: the four southern California national forests

completed a California spotted owl management strategy in 2004.)

Conclusions 

Since CASPO (Verner and others 1992), considerably more has been learned about

the ecology of the spotted owl in southern California (Carlson and others 1998, 

LaHaye 1997, LaHaye and others 1992, 1994, 1997, 2001; Smith and others 1999, 

2002). However, the status of the overall population of this species in the region is

still unclear. Demographic estimates continue to indicate the largest spotted owl

population in southern California is declining (LaHaye and others unpublished data5

).

Policy and management of this species do not appear to be clear or consistently

implemented. Further, little effort is devoted towards monitoring the species and its

habitat. 

As the human population in southern California continues to expand, the need 

for additional space and resources is likely to increase pressure on the remaining 

habitat supporting the spotted owl. Continued habitat degradation, in conjunction 

with the unknown cause(s) of the species’ regional decline, leads to an uncertain 

future for the spotted owl in southern California, particularly for the small, isolated

populations. 

Acknowledgments 

We thank M. Brosnan, D. Call, P. Carlson, D. Carpenter, D. Cousins, S. Dark, R.

Enderlein, M. Engle, R. Gamez, M. Hollister, D. Johnson, L. Johnson, P. Johnson, D. 

Kwasney, T. LaGarda, C. Lanning, R. Long, K. Maas, E. Pausch, D. Roberts, J. Saia, 

J. Silva, R. Smith T. Tur, T. Van Blankenstein, A. Vejar, C. Stermer, A. 

Wigglesworth, M. Haldeman, A. Smith, G. Sime, A. Boustany, S. Ng, T. Saladino, T.

Sylvia, M.J. Mazurek, M. Meyers, B. Newman, D. Leslie, K. Mirowski, A. Rogers,

W. Hall, J. Wills, C. Prange, J. Augustine, J. Kelly, M.M. McLeod, M. Painter, M. 

Phillippart, C. Smith, Z. Smith, and H. Tangerman for the many hours, at all times of 

the day and night, they spent helping collect field data. We thank S. Loe and the

personnel of the Mountain Top Ranger District of the San Bernardino National Forest 

for cooperation and assistance provided during this study. We also thank G. Gould 

and the California Department of Fish and Game; J. Palmer and Southern California

Edison; J. Verner and D. Solis; and Region 5 USDA Forest Service. Funding for our 

owl research was provided by the California Fish and Game Department, USDA

Forest Service (Pacific Southwest Region [Region 5]), Southern California Edison, 

Snow Summit Ski Corporation, and the University of Minnesota. 

 

5 Unpublished data on file at the Department of Fisheries, Wildlife, and Conservation Biology,

University of Minnesota, St. Paul, MN

206 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Spotted Owls in Southern California—LaHaye and Gutiérrez

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USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 209 

Bats in the South Coast Ecoregion: Status, 

Conservation Issues, and Research Needs1

Karen L. Miner2 and Drew C. Stokes3

Abstract 

California’s bat fauna is one of the most diverse in the United States. Of the 25 species of bats 

in the state, 24 have been detected in the south coast ecoregion. Many of these species appear

to have experienced population declines in the ecoregion, and 16 are officially recognized as 

sensitive (including one endangered) by wildlife regulatory agencies. Data from recent field 

survey work conducted by bat researchers were compiled in order to provide a tentative 

assessment of the current status of bats within the south coast ecoregion. These data suggest

that the pallid bat (Antrozous pallidus), Townsend’s big-eared bat (Corynorhinus townsendii),

and California leaf-nosed bat (Macrotus californicus) have experienced population declines 

and could be seriously threatened, particularly at lower elevations. This may also be true for

some of the region’s other bat species, such as the western red bat (Lasiurus blossevillii), but

additional research is needed. The Yuma myotis (Myotis yumanensis), Mexican free-tailed bat 

(Tadarida brasiliensis), and big brown bat (Eptesicus fuscus) were frequently encountered in 

both Krutzsch’s (1948) and recent field inventories, so they appear to remain relatively 

common at this time. The major threat to bats in the ecoregion is the loss of habitat (especially

riparian and oak woodland habitats) due to urban expansion as well as extermination or

disturbance of bat colonies. Characterization of species-specific distribution and seasonal 

habitat use patterns is needed so that land managers can address both foraging and roosting 

habitat requirements from a landscape perspective. Research is also needed regarding the 

effects of urbanization, insect control, tree/snag management, bat exclusions, mine closures, 

and recreational activities, specifically rock-climbing, on bat populations. 

Key words: bat conservation, Chiroptera, habitat loss, population status, species diversity

Introduction 

California has the fourth highest diversity of bat species in the United States,

following Arizona, New Mexico, and Texas, with 25 species representing three

families. Twenty-four of these species occur in the south coast ecoregion of the state,

indicating the importance of the region to bat diversity. Over two-thirds of the

region’s bat species are officially recognized as sensitive by the California 

Department of Fish and Game, U.S. Fish and Wildlife Service, and/or Federal land

management agencies at this time (table 1). The California Department of Fish and 

Game (B. Bolster pers. comm.) reports that four additional species have been

proposed to become California Species of Special Concern in the latest draft of 

“Mammal Species of Special Concern in California” (CSC* in table 1). 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 California Department of Parks and Recreation, Southern Service Center, 8885 Rio San Diego Drive, 

Suite 270, San Diego, CA 92108 (e-mail: kminer@parks.ca.gov) 3

 U.S. Geological Survey, Biological Resource Division, Western Ecological Research Center: San

Diego Field Station, 4165 Spruance Road, Suite 200, San Diego, California 92101 (e-mail: 

dstokes@usgs.gov) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 211 
Bats in the South Coast Ecoregion—Miner and Stokes 

Table 1—Bats of California and their legal status. All except Myotis occultus have been 

previously documented in the south coast ecoregion. 

Family/ Scientific name1 Common name Legal status

2

Phyllostomatidae American leaf-nosed bats 

Macrotus californicus California leaf-nosed bat CSC, FSS, BLM

Choeronycteris mexicana Mexican long-tongued bat CSC 

Leptonycteris curasoae yerbabuenae Lesser long-nosed bat FE3

Vespertilionidae Mouse-eared bats 

Myotis lucifugus Little brown bat none

Myotis occultus Arizona myotis CSC, BLM 

Myotis yumanensis Yuma myotis FSC, BLM

Myotis velifer Cave myotis CSC, BLM3

Myotis evotis Long-eared myotis FSC, BLM

Myotis thysanodes Fringed myotis FSC, CSC*, BLM

Myotis volans Long-legged myotis FSC, CSC*, BLM

Myotis californicus California myotis none 

Myotis ciliolabrum Small-footed myotis FSC, BLM 

Lasionycteris noctivagans Silver-haired bat none 

Pipistrellus hesperus Western pipistrelle none 

Eptesicus fuscus Big brown bat none 

Lasiurus blossevillii Red bat CSC*, FSS 

Lasiurus xanthinus Western yellow bat CSC*

Lasiurus cinereus Hoary bat none 

Euderma maculatum Spotted bat CSC, FSC, BLM 

Corynorhinus townsendii Townsend's big-eared bat CSC, FSC, FSS, BLM 

Antrozous pallidus Pallid bat CSC, FSS, BLM 

Molossidae Free-tailed bats 

Tadarida brasiliensis Mexican free-tailed bat none 

Nyctinomops femorosaccus Pocketed free-tailed bat CSC 

Nyctinomops macrotis Big free-tailed bat CSC 

Eumops perotis californicus Western mastiff bat CSC, FSC, BLM 

1

 Scientific names after Koopman (1993), with the exception of Corynorhinus townsendii (Frost and

Timm 1992, Tumlison and Douglas 1992), Lasiurus blossevillii, and L. xanthinus (Baker and others 

1988, Morales and Bickham 1995). 2

 Legal status categories include Federal Endangered (FE), Federal Species of Concern (FSC),

California Species of Special Concern (CSC), Forest Service Sensitive (FSS), and Bureau of Land 

Management Sensitive (BLM). Source: Calif. Dept. of Fish and Game, Special Animals List of July

2002 and species proposed to become California Species of Special Concern (CSC*, DFG—B. Bolster 

pers. comm.) 

3

 Currently known in the state from only two or three recent public health records from urban areas in 

the ecoregion (Constantine 1998). 

Despite the high diversity and sensitivity of bats in the south coast ecoregion, 

this group of mammals has been largely ignored during environmental review of 

proposed projects and in recent multi-species planning efforts. This is primarily due

to the lack of information on the distribution, seasonal habitat associations, and

population status of bat fauna. The purpose of this paper is to bring together recent

bat inventories, in order to attempt a preliminary assessment of population status and 

research needs for this ecologically important group of mammals in the south coast

ecoregion of California. 

212 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Bats in the South Coast Ecoregion—Miner and Stokes 

Historic and Recent Field Studies

Early studies of bats covering portions of the south coast ecoregion of California 

include Grinnell (1918), A Synopsis of the Bats of California; Krutzsch (1948), 

Ecological Study of the Bats of San Diego County, California; and Vaughan (1954),

Mammals of the San Gabriel Mountains of California. While all three studies 

presented species records for the region, only Krutzsch (1948) attempted to assign 

estimates of relative abundance of species with respect to environmental factors, such

as gross climatic zones, topography, life zones, and plant associations (table 2). In 

addition, 18 of the 24 species in the ecoregion are represented in Krutzsch’s study.

This allows us some measure by which to compare and contrast historic and current 

information regarding bat species richness and diversity within the ecoregion. 

Table 2—Relative abundance of bats in San Diego County (A-abundant, C-common, R-rare)

by topographic areas during the 1930s and ‘40s, after Krutzsch (1948). Scientific names as in 

table 1. 

Species 

Coastal 

plain 

Inland

valley 

Western 

foothills 

Mountains Eastern

foothills 

Desert 

Ma. californicus R R C

C. mexicana R R

M. yumanensis C A A C

M. evotis C C

M. thysanodes C C R

M. volans R R

My. californicus C C A A C C

M. ciliolabrum R C C C C

P. hesperus C C A A

E. fuscus C A A A C

L. blossevillii C C C

L. cinereus C C C

C. townsendii C A A A C C

A. pallidus A A A R C C

T. brasiliensis A A A C C

N. femorosaccus R R

N. macrotis R

E. perotis R R C R

The only recent published literature on bat distribution in the region is a review 

of public health records by Constantine (1998), in which he focused on range extensions of ten rare or uncommon species. However, government agencies and private

individuals have conducted inventories of bat fauna at various locations throughout 

the ecoregion in recent years. Written reports or data for some of these efforts have 

been obtained to develop a picture of the current status of bats in the ecoregion. 

These studies vary in duration, intensity, and area covered, but all used at least two of 

the three primary detection methods for bats: mist-net capture, acoustic detection, and

roost searches. The most extensive of the studies is the 3-year study by the USDA 

Forest Service on bat habitat associations within the four southern California national 

forests (Simons and others 2000), conducted for the Southern California Mountains 

and Foothills Assessment (Stephenson and Calcarone 1999). Results from this study

have been divided according to national forest boundaries to provide a more

geographic representation of the data. The most intensive of the studies presented are

those for Orange County (Remington 2000) and Marine Corp Air Station at Miramar 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 213 
Bats in the South Coast Ecoregion—Miner and Stokes 

(MCAS Miramar) in San Diego County (Hunsaker 2001), for which the total nights 

of effort were 68 and 73, respectively, with multiple visits to most, if not all, of the 

sites sampled. Data from the collection of unpublished studies have been summarized

and grouped into two general topographic categories: studies conducted in the coastal 

plain or inland valley areas, with the majority of sampling sites located under 300 

meters elevation, and studies conducted in the foothills or mountain areas with the 

majority of sampling sites above 300 meters (table 3). This division roughly 

corresponds to the topographic divisions in table 2 (after Krutzsch 1948), as well as 

the current location of the expanding edge of urbanization. The general study areas 

are mapped in figure 1. 

There are several caveats to keep in mind when comparing results of these

studies. Selection of sample sites among studies was subjective, and the distributions

of the sites are not likely representative of habitats present within the study area. Bat

researchers typically select sites near water or suspected roosts where the probability

of detecting and capturing bats is higher. The seasonal distribution of sampling effort 

also varied among studies with most, if not all, of the effort focused on the summer 

season. Within the region, some species are more commonly detected during migration (for example, silver-haired bat, Lasionycteris noctivagans) or in winter months 

(Mexican long-tongued bat, Choeronycteris mexicana, and hoary bat, Lasiurus 

cinereus). The studies also differ in the relative emphasis of the survey methods employed. Some species are not readily detected acoustically due to either low call 

amplitude or inconsistent use of echolocation (Townsend’s big-eared bat, 

Corynorhinus townsendii, and Mexican long-tongued bat), while others are not

readily captured in mist nets (western mastiff bat, Eumops perotis). In addition,

species identification based solely on limited acoustic data can be problematic, as

with Myotis species. In consideration of the above, we recognize that the lack of 

species detection within any one study does not necessarily mean that a species is

absent from the study area. However, together these studies do provide some

indication of the frequency at which species are detected during field investigations 

across the ecoregion. 

Both Krutzsch (1948) and recent field studies (table 3) detected 18 species.

However, Krutzsch’s list included the Mexican long-tongued bat, while recent studies include the spotted bat (Euderma maculatum). In general, both historic and recent

studies indicate that species diversity is higher in the foothills and mountains than the 

coastal plain and inland valleys in terms of the relative frequency of detection, 

particularly for the genus Myotis. 

Other Records 

Five other species occurring in the ecoregion, unaccounted for in the above studies, 

include the little brown bat (Myotis lucifugus), cave myotis (Myotis velifer), western

yellow bat (Lasiurus xanthinus), silver-haired bat, and the federally-listed endangered 

lesser long-nosed bat (Leptonycteris curasoae yerbabuenae). These species are

reported in the ecoregion based on a few individual specimens and/or public encounters in urban settings only (Bond 1977, Constantine 1998). Additional information on bat distributions can be obtained through investigation of encounters reported

by the public (Remington 2000, KLM and DCS pers. obs.) These records compliment

recent field studies by providing information for the coastal plain/inland valley 

214 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Bats in the South Coast Ecoregion—Miner and Stokes 

Table 3—Summary of recent bat inventory studies in the south coast ecoregion showing the 

number and percent (shown in italics) of sites at which each bat species was detected. The 

studies are grouped into two general topographic categories: coastal plain and inland valleys 

(A-F) and foothill and mountains (G-N).

Study1

A B C D E F G H I J K L M N

Effort (nights/sites) 

Species 

68 

6 

30 

26 

12 

9 

73 

23 

16 

7 

36 

24 

33 

21 

3 

7 

11 

8 

14 

12 

21 

17 

26 

14 

33 

21 

35 

19 

Ma. 1

californicus 4

M. 6 19 15 6 13 15 7 5 5 3 3 7 4

yumanensis 100 73 65 86 54 71 100 63 42 18 21 33 21

M. 3 1 7 2 3 7 3

evotis 13 13 58 12 21 33 16

M. 1 2 2 1 5

thysanodes 8 12 14 5 26

M. 2 3 3 4

volans 12 21 14 21

My. 3 2 7 8 6 7 9 7 9 12

californicus 50 29 29 38 75 58 53 50 43 63

M. 4 10 12 4 6 10 4 4 6 10

ciliolabrum 57 42 57 57 75 83 24 29 29 53

P. 2 1 7 9 8 3 3 10 9 8 3 3

hesperus 8 4 100 38 38 43 38 83 53 57 14 16

E. 4 17 3 6 12 17 4 7 10 9 9 18 14

fuscus 67 65 13 86 50 81 57 88 83 53 64 86 74

L. 1 1 2 4 2 1 4 1 1 1 5

blossevillii 17 11 9 57 8 5 50 8 6 5 26

L. 1 2 4 4 1 1 2 3

cinereus 4 29 17 50 8 7 10 16

C. 1 3 1 2 4 1 5

townsendii 14 13 14 25 33 5 26

E. 1 1

maculatum 6 5

A. 3 7 3 1 2 2 4 3

pallidus 50 27 13 5 29 25 24 21

T. 6 18 1 11 7 16 11 2 6 8 3 3 5

brasiliensis 100 69 11 48 100 67 52 29 75 67 18 14 26

N. femoro- 2 5 9 7 13 10 6 9 2 7

saccus 33 19 39 100 54 48 75 75 10 37

N. 1 1 3 2 1

macrotis 11 14 13 25 8

E. 6 12 10 7 12 13 2 8 10 6 1 7 11

perotis 100 46 43 100 52 62 29 100 83 35 7 33 58

1

 A) Orange County, Remington 2000; B) Camp Pendleton Marine Corp Base, Brown and Berry 1999; 

C) Pt Loma, USGS 2002 unpublished data D) Marine Corp Air Station, Miramar, Hunsaker 2001; E) 

San Diego National Wildlife Refuge, Stokes unpublished data; F) San Diego County MSCP area, USGS 

2002 unpublished data; G) Southern Santa Ana Mountains, Fisher & Crooks 2000; H) Potrero Creek, 

Riverside County, Stokes unpublished data; I) Ysabel Ranch Preserve, USGS 2002 unpublished data; J) 

Cleveland National Forest, Descanso District, USGS 2002 unpublished data; K) Los Padres National 

Forest, L) Angeles National Forest, M) San Bernardino National Forest, and N) Cleveland National

Forest, Simons and others 2000. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 215 
Bats in the South Coast Ecoregion—Miner and Stokes 

Figure 1—Locations and general area covered by recent bat inventories conducted 

within the south coast ecoregion. Letters correspond to studies as in table 3. 

topographic regions, which are more urbanized now than during Krutzsch’s study

and, therefore, under-represented in recent fieldwork conducted in wildlands. 

Current Population Status 

Declining Populations 

Inferences regarding the population status of bats in the south coast ecoregion can be

made from the information presented above. There is reasonably strong evidence of

population declines within the ecoregion for three species: the pallid bat (Antrozous 

pallidus), Townsend’s big-eared bat, and California leaf-nosed bat (Macrotus 

californicus). 

The pallid bat is a large-bodied species that typically roosts in a variety of 

crevice-like situations and feeds primarily on large arthropods, including several 

terrestrial species (Barbour and Davis 1969). It appears to have undergone a serious 

decline within the ecoregion, particularly in the low-lying areas. Krutzsch (1948) 

described the coastal race as being abundant from sea level through the western

foothills, occurring in riparian, chaparral, oak savannah, and cultivated areas. All of 

the roosts located by Krutzsch were in man-made structures, often in association with 

other common species such as the Yuma myotis (Myotis yumanensis), big brown bat

(Eptesicus fuscus), and Mexican free-tailed bat (Tadarida brasiliensis). Vaughan

(1954) referred to this species as “probably the most common and characteristic bat 

of the citrus belt” at the Pacific base of the San Gabriel Mountains. However, by the 

1970s only one of twelve roosts previously described by Krutzsch as having been 

216 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Bats in the South Coast Ecoregion—Miner and Stokes 

consistently occupied for a decade in San Diego County was still extant (P. Brown 

pers. comm.). In recent years, the pallid bat has been detected in the Transverse

Range (Simons and others 2000), sycamore-oak riparian areas in and around the 

Santa Ana Mountains (Brown and Berry 1999, Fisher and Crooks 2000, Remington

2000), and a few inland valley sites in southern San Diego County (USGS 20024

).

During the last decade in San Diego County we have observed numerous colonies 

occupied by the three common bat species listed by Krutzsch as pallid bat associates;

however, very few pallid bat colonies have been documented. Although this species 

can be found in rural settings, it appears to be intolerant of urban development (E. 

Pierson pers. comm., Remington 2000). Populations will likely continue to decline as

urban expansion encroaches into the foothills of the region. 

The Townsend’s big-eared bat occurs in a variety of habitats, including desert 

scrub, sagebrush, chaparral, and deciduous and coniferous forests, where it feeds

primarily on moths (Barbour and Davis 1969). However, its distribution is strongly 

correlated with the availability of roosting habitat such as caves or cave-like

structures including mines, tunnels, abandoned buildings, and bridges (Pierson and

Rainey 1998a). According to Krutzsch (1948), the Townsend’s big-eared bat was 

“widely and commonly distributed from the seacoast to the desert.” He described it

as common on the coastal plain and abundant in the inland valleys, foothills, and

mountains (table 2). Now it appears that within the south coast ecoregion the species

is relatively rare and limited primarily to the foothill and mountain areas (table 3). 

Pierson and Rainey (1998a) reported that the cismontane maternity colonies 

described by Krutzsch in San Diego County were no longer extant. This species was 

often encountered during recent mine surveys conducted above 850 m within the 

ecoregion, but few maternity colonies were confirmed (Miner and others 2000, 

Simons and others 2000). The Townsend’s big-eared bat is not readily detected in

mist net or acoustic surveys; roost searches are the most efficient survey method for 

this species. However, roost surveys must be undertaken with caution because 

disturbance at the roost can result in abandonment by this species. The loss of or 

disturbance to both maternity and hibernation roost sites has been recognized as the 

primary threat to the Townsend’s big-eared bat throughout the western United States 

(Pierson and others 1999). There are few known extant maternity colonies within the

ecoregion, and none can be considered adequately protected. Due to the species’ 

vulnerability to human disturbance at roost sites, further declines might be expected

as the human population increases within the region. 

The California leaf-nosed bat is a tropical, insectivorous species that ranges

north into the lower Sonoran life zone of southern California, Arizona, and New 

Mexico. It is an obligate cave or mine roosting species and requires roost 

temperatures in excess of 26°C year-round to survive because it lacks the ability to

hibernate (Bradshaw 1962 in Anderson 1969). Historically, the California leaf-nosed

bat was known from a few caves, tunnels, or mines in the inland valleys of the south

coast ecoregion (Bond 1977, P. Brown pers. comm., Constantine 1998, Grinnell 

1918, Krutzsch 1948). Live individuals of this species have not been observed at the 

Los Angeles/Ventura County sites since the late 1940s or early 1950s (Constantine 

1998), nor have they been seen at the northern San Diego County sites since the early

1900s (P. Brown pers. comm.). However, one extant small colony (less than a dozen

bats) continues to persist in a tunnel at an inland valley site in southern San Diego 

 

4

 Unpublished data on file, U.S. Geological Survey, Western Ecological Research Center, San Diego, CA. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 217 
Bats in the South Coast Ecoregion—Miner and Stokes 

County (P. Brown pers. comm., USGS 20025). While the California leaf-nosed bat

can still be found in the California deserts, this species appears to have been all but 

extirpated from the south coast ecoregion. 

Stable Common Populations

Based on the information provided by the various bat inventories (table 3) and the 

authors’ experience investigating calls received by San Diego Bat Conservation 

(SDBC), it is clear that three species—Yuma myotis, Mexican free-tailed bat, and big 

brown bat—are relatively common within the south coast ecoregion and appear to be 

relatively urban-adapted. All three readily roost in man-made structures. The Yuma

myotis forages primarily over open water, although it will forage in other habitats, 

and often roosts near available water sources. The Mexican free-tailed bat covers

large areas while foraging over a variety of habitats. This, along with its generalized

roosting requirements (caves, rock crevices, buildings, and bridges), makes it

probably the most adaptable species in urban landscapes. The big brown bat is

typically found in association with wooded areas, whether natural or planted. 

Although fairly common in both field studies and SDBC public contacts, evidence 

suggests that this species might not persist in small isolated habitat fragments or 

highly urbanized areas (Remington 2000.) 

Stable Uncommon and Rare Populations 

There are several uncommon-to-rare species that appear to have relatively stable 

populations within the ecoregion. These include the western pipistrelle (Pipestrellus 

hesperus), small-footed myotis (Myotis ciliolabrum), long-eared myotis (Myotis

evotis), fringed myotis (Myotis thysanodes), long-legged myotis (Myotis volans), and 

Mexican long-tongued bat. While the desert population of western pipistrelle is 

abundant in rocky areas, the coastal population apparently always has been less 

abundant (Krutzsch 1948). This species appears to still be present in rocky canyon 

habitat of the inland valleys, foothills, and mountains. The small-footed myotis 

occupies the same topographic regions as the western pipistrelle but seems to be 

associated more with riparian and wooded habitats. While still regularly encountered, 

both of these species may be experiencing habitat loss in the lower elevations of their 

range within the ecoregion. 

The long-eared myotis, fringed myotis, and long-legged myotis occur primarily

in forested habitat in the mid- to upper elevations of the region. Based on limited 

historic collections, they likely have never been particularly common, especially the 

latter two species (Grinnell 1918, Krutzsch 1948, Vaughan 1954). Recent field studies in the upper elevations of the ecoregion detected these species at a few sites 

within each study area, indicating that they are patchily distributed throughout the 

forests. These species have been documented roosting in snags or trees in other portions of their range (Barclay and Brigham 1996). However, during the USDA Forest 

Service’s southern California forest bat study, individual bats of these species were

radio-tracked to rock crevices and building roosts (Miner and Brown 1996). The only

maternity colony documented during the study was a building roost of long-eared 

myotis (~ 30 individuals). Timber harvest is the major threat to these species in other 

portions of their range (Barclay and Brigham 1996). While currently this is not a 

serious threat within the ecoregion, there is concern that the removal of snags or

 

5

 Unpublished data on file, U.S. Geological Survey, Western Ecological Research Center, San Diego, CA. 

218 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Bats in the South Coast Ecoregion—Miner and Stokes 

trees, such as is done for hazard abatement, fire suppression, and forest pest 

management, may seriously impact Myotis species. Of particular concern is the lack

of known extant maternity colonies for the fringed myotis and long-legged myotis (B. 

Bolster pers. comm.). 

The Mexican long-tongued bat feeds on pollen and nectar, specializing in the

flowers of agaves and columnar cactus. It will also visit hummingbird feeders and 

other nectar-rich plants (Arroyo-Cabrales and others 1987, W. Peachey pers. comm.),

possibly even farmed avocado plants in the ecoregion (public contact investigations). 

The southern part of the ecoregion represents the northwestern extreme margin of the 

otherwise subtropical distribution of the Mexican-long-tongued bat. It roosts in shallow caves and mines throughout most of its range. However, in California the species 

has yet to be documented in the wildlands and instead is found in residential areas,

roosting in garages, sheds, porches, and under houses built on stilts. The Mexican 

long-tongued bat was first recorded in San Diego County in the late 1940s in 

developed areas along the coast and inland valleys (Olson 1947). Since then, 

individuals have been turned in to public health departments in Los Angeles, Orange, 

and Ventura counties (Constantine 1998). While the Mexican long-tongued bat is 

regularly seen during fall and winter in San Diego residential areas, it is as yet 

undocumented that it remains during the maternity season, and it is presumed not to

breed in California. 

Increasing Populations 

The western yellow bat is another primarily subtropical species. It is known to roost 

among the fronds of palm trees and was first recorded in palm habitat of the California desert in the 1940s (Constantine 1946). Now it is regularly encountered yearround in urban areas throughout the region, primarily in association with planted 

palm trees (Constantine 1998, D. Simons pers. comm., KLM and DCS pers. obs.), 

and it is assumed to breed within the ecoregion (P. Brown, pers. comm.). This

apparent expansion into the ecoregion may be in response to exotic landscape

plantings and possibly to global warming (Constantine 1998). 

Status Unknown 

It is not possible to determine the population status for the 11 remaining species

within the ecoregion based on available information. This is due to either the seasonal or elevational bias of the field studies or the lack of comparable historical data. 

The survey effort for most of the field studies presented above (table 3) was 

concentrated during the summer months. Therefore, tree-roosting species such as the 

silver-haired bat and hoary bat that occur in the ecoregion primarily during migration 

or winter are less frequently encountered. Most hoary bat records within the ecoregion are between late September and May (Vaughan and Krutzsch 1954, authors’ 

public contact investigations). However, some individuals remain all summer in the 

higher elevations of the mountains (KLM and DCS pers. obs., Simons and others

2000, Vaughan and Krutzsch 1954). The higher detection frequency of hoary bats in 

the most recent field studies in San Diego County (studies E and F in table 3) is due 

to spring survey effort, when hoary bats are migrating. The silver-haired bat is a

migratory species associated primarily with coniferous forests and is not known to 

breed in southern California (Barbour and Davis 1969). It is only represented within 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 219 
Bats in the South Coast Ecoregion—Miner and Stokes 

the ecoregion from public contact records (Constantine 1998), presumably encountered during migration. 

Urbanization in the south coast ecoregion has removed habitat primarily at 

lower elevations. For bat species that rely on the impacted habitat types, population 

status would be predicted to decline. Such appears to be the case for the western red

bat and California myotis (Myotis californicus). However, surveys focused on these 

habitat types, as well as urban landscaped areas, are needed before conclusions can

be drawn. The western red bat (Lasiurus blossevillii) is a solitary, migratory species 

that roosts in the foliage of trees and shrubs (Shump and Shump 1982). It has a strong

association with riparian woodlands, especially in the lowlands of the Central Valley 

and southern California (Pierson and others 2000). It has also been observed to roost

in orchards (Constantine 1959, Grinnell 1918, Krutzsch 1948, Vaughan 1954). According to Grinnell (1918), the sexes segregate during late spring and summer, with

females remaining in the lowlands and males moving to higher elevations. However, 

Krutzsch (1948) found both sexes of this species to be common year-round from the 

coast up into the western foothills. After reviewing location records for the species,

E. Pierson (pers. comm.) postulated that coastal southern California could provide

very important wintering as well as summering habitat for the western red bat. 

Additional studies are needed at lower elevations during both winter and summer to 

determine whether this species, like riparian birds, has indeed suffered critical habitat

loss from agricultural conversion, reservoir construction, and urban expansion. The 

California myotis has generally been considered to be common to abundant

throughout all topographic regions of the ecoregion (Krutzsch 1948), particularly in 

oak woodland canyons (DCS pers. obs., Vaughan 1954). However, current data suggest that local declines may be occurring in the more developed portions of the 

coastal plains and inland valleys. 

Bat surveys in urban settings in coastal plain and inland valley areas could help

determine the status of other bat species in the ecoregion as well. Constantine (1998)

documented the first and only two specimens of the federally listed, endangered

lesser long-nosed bat for the state from southern California public health records. 

This species is similar in habit to the Mexican long-tongued bat in that it is migratory, forages on pollen and nectar, and roosts openly in cave-like situations. Like the 

Mexican long-tongued bat, it appears to be found within the ecoregion during the fall

and winter in association with exotic landscaping in urbanized areas. Another species

that is only known in the ecoregion from a few public health records is the cave 

myotis, which previously was only known in the state along the Colorado River

(Constantine 1998). While these records may represent vagrants or inadvertently 

transported individuals, it is possible that these species, like the Mexican longtongued bat and western yellow bat, may be expanding their range into the ecoregion. 

It is difficult to determine current population status relative to historic occurrence information for several species of bats, including the western mastiff bat, big 

free-tailed bat (Nyctinomops macrotis), pocketed free-tailed bat (Nyctinomops 

femorosaccus), and spotted bat. This is due primarily to the fact that acoustic detection is now used more frequently as a survey method, greatly increasing the detectability of these species. Historically, the greatest concentration of western mastiff bat 

roosts documented in the state was in southern California (Pierson and Rainey 

1998b). Many of these roosts were located near the base of the Transverse Range in 

the Los Angeles basin, often in buildings. Natural roosts for this species are located 

primarily in rock crevices on cliffs and large rock outcrops. The western mastiff bat 

220 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Bats in the South Coast Ecoregion—Miner and Stokes 

was detected in all but the USGS Point Loma study represented in table 3, and it was 

previously detected at Point Loma as well (P. Brown pers. comm.). This does not 

necessarily mean that it is abundant within the ecoregion. The detection rate of the 

mastiff bat is greater relative to other species because it has an echolocation call

falling within the human range of hearing that can be heard at distances up to 300

meters. Because of this, its relative abundance can be overestimated. Of particular 

concern is the apparent decline of this species in the northern portion of the Los 

Angeles basin. In their statewide review of the western mastiff bat, Pierson and

Rainey (1998b) were unable to detect this species in areas that had provided 

numerous historic records. The western mastiff bat was also recorded only from one 

site in the Angeles National Forest (table 3), providing further evidence for a possible 

decline in that area. This species typically forages over wide-open spaces of various 

habitats (Pierson and Rainey 1998b). It is likely that urban development of the 

coastal plain has greatly reduced foraging habitat for this species in the ecoregion. 

As with the western mastiff bat, pocketed and big free-tailed bats appear to be

more frequently detected now than historically, both in field studies (table 3) and 

public health records (Constantine 1998). This could be due to increases in their 

populations within the region or to improved survey methods and increased public

contact with these species. Both of these species roost in rock crevices on cliff faces

and have been found to use abandoned rock quarries in San Diego County (KLM and

DCS pers. obs.). The pocketed free-tailed bat is detected throughout the southern

portion of the south coast ecoregion and east through the deserts. The big free-tailed 

bat is much rarer and found further north along the coast than the pocketed free-tailed 

bat (Constantine 1998). 

The spotted bat always has been represented in the region by few records and is

probably so rare as to be easily missed without committing to focused searches.

Likewise, the Transverse Ranges forming the northern border of the ecoregion may

represent the most southern limit of the little brown bat in the state, explaining the 

rarity of records for this species. 

Conservation Issues 

Although specific information from the south coast ecoregion is lacking, the basic 

ecology of bats and information from focused studies conducted elsewhere provide

indications of relevant conservation issues. Bats have specific seasonal roosting 

requirements that may or may not be met within a given roost, area, or region. Most 

bats will change roosts several times over their annual cycle, through longitudinal or 

altitudinal migration or local movements (Barbour and Davis 1969). Recent research

has shown that many tree roosting species will switch roosts every few days (Barclay 

and Brigham 1996). This means that multiple roosts of varying temperature regimes 

need to be available within appropriate habitat and flight distance of the species for a 

population to remain viable. Bats also need adequate foraging habitat within the 

nightly commute distance from a given roost. Commute distances vary among

species (Pierson 1998) and seasonally within a given species (Brown and others 

1995). Therefore, spatial distribution of roosts and foraging habitat may be critical 

for sustainable bat populations. This necessitates that a landscape-level approach be 

taken when considering bat conservation, especially in a region experiencing largescale urban development and increased habitat fragmentation such as the south coast

ecoregion. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 221 
Bats in the South Coast Ecoregion—Miner and Stokes 

Another important conservation issue with respect to bats is the wanton

destruction of or disturbance to bat colonies. Many species of bats aggregate into 

colonies during at least a portion of their annual cycle, making them vulnerable to 

catastrophic events. This is particularly true for colonies that find the microclimatic 

conditions of man-made structures favorable, placing them in close proximity to 

humans. Irrational fears and misconceptions regarding health risks have led to the 

needless destruction of numerous bat colonies. In addition, several species of bats, 

most notably the Townsend’s big-eared bat (Pierson and others 1999), will abandon

roosts if frequently disturbed. Due to low fecundity (typically one pup per year, 

Barbour and Davis 1969), longevity (individuals may live for over a decade, Hill and 

Smith 1984), and the fact that colonial bats typically exhibit high site fidelity from 

year to year (Lewis 1995), recovery or recolonization is slow if it occurs at all. While 

public education can help dispel myths about bats, and humane exclusions can spare 

the lives of bats, physical protection of occupied roosts is crucial for bat 

conservation, particularly for the rarer species. 

Even though there is a wide range of threats to bats (Pierson 1998), the greatest

threat in the south coast ecoregion is urban/suburban expansion and its associated

impacts to roosts and foraging habitat. Loss or conversion of habitat at the lower 

elevations, particularly of riparian and oak woodlands, has likely had significant 

impacts on regional bat fauna. Moreover, once bats take up residence in suitable 

anthropogenic roosts and are discovered, they are often driven out or exterminated.

Even tree trimming activities (for example, palm skinning) can impact bats that

attempt to roost in landscape plantings. The effects of urbanization reach beyond the

immediate developed areas. These edge effects include recreational activities, which

increase disturbance to nearby natural roosts and mines, and increased predation from 

human commensal animals, particularly domestic cats. Thus, the urban-wildland

interface may act as a sort of population “sink,” where bats appear to be common at

first (possibly even attracted to man-made structures) but then decrease in abundance 

or richness with time and further development of the area (Remington 2000). As the 

urban landscape continues to encroach into the foothills, the number and relative 

abundance of bats in the ecoregion is likely to decline. 

Other relevant threats to bats in the south coast ecoregion include water projects

or the development of transportation routes, in which riparian habitat is reduced or 

geomorphic roosting habitat (cliffs and mines) is blasted or inundated. Several 

historic roosts appear to have been lost in this fashion (Pierson pers. comm.). However, dams and highway bridges do provide roosting habitat for some species of bats. 

Many of the dams and bridges within the region currently house bat colonies, 

primarily of Yuma myotis and Mexican free-tailed bats. Bridge replacement or

alteration has the potential to eliminate resident colonies, but California Department 

of Transportation and other jurisdictions are now considering impacts to bats by their 

bridge projects and ways to avoid them. 

Closure of mines for hazard abatement or renewed mining activities can greatly 

impact many species of bats, including several threatened species, such as the Townsend’s big-eared bat (Brown 1995, Riddle 1995, Tuttle and Taylor 1994). The USDA 

Forest Service and Bureau of Land Management are making an effort to conduct bat 

surveys for such projects and address potential impacts. 

Recreational activities such as cave or mine exploration and rock climbing can 

disturb bats, potentially affecting reproductive success and survival or causing roost 

abandonment. While impacts from cave and mine entry are well documented (Hill

222 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Bats in the South Coast Ecoregion—Miner and Stokes 

and Smith 1984, McCracken 1989, Tuttle 1979), impacts to bats from rock climbing

need to be investigated. Once important mine roosts are discovered, protection measures (for example, gating) should be implemented following established guidelines 

for the species involved. 

Prescribed burns and wildfires can create snags and cavities favored by several

species of bats for roosting (Barclay and Brigham 1996), but these snags also may be 

destroyed in a fire or even selectively removed afterwards to reduce perceived

hazards. Another concern stems from the observation that some lasiurine bats will 

roost under leaf litter during winter (Moorman and others 1999, Saugey and others 

1998) and, thus, could be killed when fires (for example, prescribed burns) occur

during winter months. Smoke inhalation also can kill roosting bats if the fire or 

smoke enters the mines or caves. 

Because most of our bats are insectivorous, the application of pesticides is a

threat to bats. Not only is there a potential for reduction of the prey base, but bats also

may be directly poisoned, either through consumption of tainted prey or by being 

sprayed while roosting in trees or orchards. Although chlorinated hydrocarbons, 

which were documented to impact bats (Clark 1981), are now banned in the United 

States, the effects of the now popular organophosphates on bats and other wildlife 

have not been fully investigated. Even application of Bt (Bacillus thuringiensis) may

significantly reduce the prey base, especially moths, which form the majority of the

diet for many species of bats (Ross 1967). For further discussion of conservation 

issues and threats to bats, see Pierson (1998). 

 

Research Needs 

Bats need to be recognized as ecologically important members of natural communities that warrant greater research effort within the ecoregion. Due to the lack of current information, until now bats have been excluded from regional habitat 

conservation planning efforts. General bat inventories need to be conducted across 

the ecoregion during all seasons. Priority should be given to the coastal plains and

inland valleys where habitat loss is proceeding rapidly and species appear to be on 

the decline. These areas also appear to be important to wintering bats and as foraging

areas for far-ranging species that roost in the foothills, such as the free-tailed bats.

Focused, systematic surveys need to be conducted to determine the status and

distributions of species that appear to have suffered population declines and are vulnerable to extirpation. These species include the California leaf-nosed bat, Townsend’s big-eared bat, pallid bat, and western red bat, followed by the western mastiff 

bat, hoary bat, and possibly the California myotis. The availability of suitable roost 

sites is a major limiting factor for many species, and roosting aggregations are extremely vulnerable to disturbance or destruction by humans. Therefore, roost sites,

particularly for the rarer species, need to be located and given the highest level of 

protection. However, it is also important to determine and protect the foraging habitat 

associated with those roosts if the colony is to remain viable. To this end, research

into species-specific home-range size and attributes, including juxtaposition of 

roosting and foraging habitat, would be valuable. This is particularly true with respect to preserve design and management. Land managers also would benefit from 

information on the effects of recreational activities, such as rock-climbing, and the 

effects of prescribed fire. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 223 
Bats in the South Coast Ecoregion—Miner and Stokes 

Given that the bats of the region are almost exclusively insectivorous, research 

also is needed on factors affecting insect distribution and abundance, including the

effects of pesticides, exotic vegetation, and artificial lights. 

The south coast ecoregion, with its rapidly increasing human population and 

associated urban expansion, provides an ideal opportunity to study patterns of urban 

landscape use by bats and the effects of habitat loss and fragmentation on bat species

diversity and reproductive success (Remington 2000). 

Other potential research topics relevant to the ecoregion include the response of 

bats to riparian habitat restoration, the effectiveness of artificial roosts used as 

mitigation, the effects of climatic change on bat distribution, and the value of bats as 

indicators of healthy ecosystems. 

While in the past bat research has been logistically difficult, recent technological

advancements provide valuable tools to aid ecological investigations. We hope that 

current and future efforts will greatly expand our knowledge of this fascinating and 

ecologically important group of animals and provide the information necessary for 

their conservation within the ecoregion and elsewhere. 

Acknowledgments 

We greatly appreciate bat biologists P. Brown, E. Pierson, S. Remington, and D.

Simons for their time and willingness to share data and experiences in the interest of 

bat conservation. We also would like to thank W. Berry, B. Bolster, and S. 

Remington for providing access to unpublished information. We are grateful to T. 

Conkle, R. Fisher, M. Hanley, R. Huffman, D. Lawhead, T. Oberbauer, J. 

Stephenson, S. Weber, and K. Winter for initiating and securing funding for bat 

inventories within the south coast ecoregion. 

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The Santa Margarita River Arundo donax

Control Project: Development of Methods 

and Plant Community Response1

Dawn M. Lawson2, Jesse A. Giessow, and Jason H. Giessow3

Abstract 

A large-scale effort to control the aggressively invasive exotic species Arundo donax in the Santa 

Margarita River watershed in California’s south coast ecoregion was initiated in 1997. The project 

was prompted by the need for Marine Corps Base Camp Pendleton to address impacts to habitat

for federally-listed endangered species and wetlands regulated by the Army Corps of Engineers. 

As of 2000, 27 km of the main stem of the Santa Margarita River had been treated. The methods 

employed were tested in preliminary trials before widespread implementation; additional

techniques emerged during the course of the project. Although the primary target is A. donax, 14 

other invasive exotic species that also threaten riparian ecosystem functions were treated when

encountered. Vegetation-monitoring transects were established in A. donax removal areas to

document the effectiveness of the different treatments and recovery of plant communities. A. 

donax was reduced by over 90 percent after the first treatment and accounted for less than 2

percent absolute cover after three follow-up annual treatments. An experiment testing low-cost 

methods for establishing woody species after A. donax control was also conducted. Cuttings were

installed with no follow-up maintenance. Approximately 30 percent of Baccharis salicifolia and

large (3-6 m) willow cuttings survived 2 years. 

Key words: Arundo donax control, invasive species, riparian restoration, Santa Margarita

River

Introduction 

Arundo donax is an invasive, non-native plant seriously impacting much of 

the riparian habitat throughout coastal California including Marine Corps Base Camp

Pendleton. It is estimated to have infested 68 percent of the riparian vegetation in the 

central portion of the Santa Ana River (Douthit 1993) and is thought to be the most 

serious exotic pest plant problem in southern California coastal rivers (Bell 1998). It

grows up to 10 m tall and can be found in small clumps to monocultures of hundreds 

of hectares (Cummins and Zedler 1998, Dudley 2000). It is not known to produce 

viable seed in the western United States, and so it spreads only vegetatively through 

rhizomes or stem fragments (Dudley 2000). Establishment from vegetative parts requires either burial in deep, moist layers of soil or near-saturated conditions at the

surface for prolonged periods. Therefore, establishment of new patches occurs

primarily in conjunction with flooding. Between flood events, spread occurs outward

from existing patches (Else 1996). Prior to initiation of control on the Santa 

 

1

 An abbreviated version of this paper was presented at Planning for Biodiversity: Bringing Research

and Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 

2, 2000, California State Polytechnic University, Pomona, CA. 2

 Senior Natural Resources Specialist, Southwest Division Naval Facilities Engineering Command, US 

Navy, 1220 Pacific Hwy, San Diego, CA 92132. (e-mail: lawsondm@efdsw.navfac.navy.mil) 3

 Restoration Biologist, Dendra Inc. 1003 Hygeia Ave. Encinitas, CA 92024 (e-mail: jgiessow@cox.net) 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 229 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

Margarita River on Camp Pendleton, A. donax occupied approximately 26 percent of 

the riparian system, including 13 percent of the 2-year floodplain (bankfull channel 

geomorphic surface), 39 percent of the 2-10-year floodplain (active floodplain 

geomorphic surface) and 23 percent of the riparian system outside of the 10-year 

floodplain (terrace geomorphic surface) (Cummins and Zedler 1998, Smith and 

Lichvar 1999). The total area infested (greater than 80 percent A. donax cover) in the 

watershed prior to treatment was estimated at 300 ha (260 ha on Camp Pendleton and

40 ha upstream) on over 40 km of the main stem and tributaries (Giessow, unpubl. 

data). 

Need for Control 

Large stands of A. donax displace native vegetation and associated wildlife species 

and increase flood and fire hazard (Bell 1998, Dudley 2000). On the Santa Margarita 

River, three endangered species have suffered habitat loss resulting from invasion of

A. donax: Vireo bellii pusillus (least Bell's vireo), Empidonax traillii extimus

(southwestern willow flycatcher), and Bufo californicus (arroyo toad). On the Base,

habitat loss resulting from A. donax invasion is considered to pose the greatest threat

to arroyo toads (ACS/ES Camp Pendleton, 1994), which likely are unable to move

through dense stands (Lovich and others 2001). In San Diego County, however, most 

riparian restoration has been done to mitigate impacts to the least Bell's vireo (Kus 

1998). 

Because riparian communities are adapted to high levels of natural disturbance, 

one strategy proposed for recovery of these systems is to remove A. donax and other

invasive exotics and rely on natural flood dynamics for the recovery of native communities. Because of the high cost of restoration of woody riparian habitats, A. donax

removal followed by natural re-colonization by native species achieves a cost savings 

that allows much larger areas to be restored (Bell 1998). On Camp Pendleton, A.

donax control, with or without replanting of native vegetation, serves to mitigate 

temporary and permanent impacts to endangered riparian species. This is based on 

the premises that natural processes will result in the recovery of native habitats that

support the listed species and that control will prevent future habitat losses from anticipated A. donax spread (ACS/ES Camp Pendleton 1994, Babbitt and others 1995). 

Strategies for Control 

Control efforts need to be coordinated and sustained due to the sheer size of the effort 

required to eradicate A. donax (Bell 1998). At the current cost of $49,000 per ha for 

eradication (initial control and 19 years of follow-up), the total cost to eradicate A. 

donax from the Santa Margarita River watershed is estimated at $15,000,000 (J. 

Giessow unpubl. data4

). It is unlikely that such a sum is available from a single

source, even in smaller amounts spread over time, and consequently, it is likely that 

there will be discontinuities in control efforts between the beginning and the end of 

eradication. 

Throughout southern California, restoration of riparian habitat is compromised

by A. donax invasions of restored sites (Stein 1998). Because A. donax establishes

readily from vegetative fragments, but not from seed, it was evident that control at 

 

4

 Unpublished data on file at Dendra Inc. 1003 Hygeia Ave. Encinitas, CA 92024 

230 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

the Santa Margarita River should proceed from the upstream portions of the

watershed downstream. Without this approach, removal in the lower reaches would

be only temporary, because dispersal of vegetative parts downstream would 

reestablish populations. The lack of seed production means that once A. donax is 

removed from a drainage in this manner, it can remain free of A. donax indefinitely. 

Reintroduction by human activities (for example, illegal dumping of yard trimmings) 

can occur, but it is rare enough that periodic checks should be sufficient to prevent its 

reestablishment. The likelihood of discontinuities in the overall control effort 

heightens the need to begin upstream.

Site-specific control methods needed to be developed where A. donax was intermixed or adjacent to habitat for federally listed species, to avoid direct and indirect

impacts to those species as required under the Federal Endangered Species Act. This

was achieved through limits on control activities during the breeding season for 

southwestern willow flycatchers and least Bell’s vireos (March 15 to August 31) and 

the use of methods that minimize damage to native woody species during control

activities. To avoid damaging egg masses of arroyo toads, A. donax control is limited

during the toad breeding season, generally from February 1 to July 15. In addition,

activities are minimized near damp sandy areas during September to avoid impacts to

juvenile arroyo toads. 

The most effective time to begin control using foliar spray treatments is August,

when the plant begins to move photosynthate into the rhizomes after flowering but

before the onset of winter dormancy, when leaves turn yellow, usually beginning in 

November (Bell 1998). Because of the need to avoid impacts to endangered species,

the actual time available for foliar herbicide application is approximately 8 to 10

weeks per year, from mid-September to the onset of winter dormancy. Mechanical

control theoretically can occur at any time, but it is restricted during the spring 

breeding season. Occasional exceptions have been made in informal consultation 

with the U.S. Fish and Wildlife Service, allowing control during the breeding season

in large, solid stands of A. donax and toward the end of the breeding season in 

August if it could be documented that all nesting endangered species had left the 

area. 

In 1994, the U.S. Army Corps of Engineers issued permits for three actions 

under the Clean Water Act that required on-site A. donax control one to four times. In 

1995, at the request of the Marine Corps, the Corps of Engineers modified the

permits to consolidate the control efforts, move them upstream, and extend the length

of control to 5 years. To accomplish this, the Marine Corps partnered with The 

Nature Conservancy, the Mission Resource Conservation District, and private land 

owners to move mitigation commitments as far as possible upstream in the 

watershed. This consolidated project was conducted as an experiment to test control

methods for watershed-wide eradication. Our objectives were to document treatment 

effectiveness on A. donax removal, improve methods as we learned more, document 

and minimize costs of treatment, and monitor plant community recovery. Our 

expectation was that A. donax could be removed on a large scale and that within 5 

years the cover would be at a maintenance level of less than 5 percent. Plant

community recovery objectives were defined by numerical targets for areas of

riparian habitat throughout the Base. When it became clear that watershed-wide 

eradication of A. donax was feasible and the recovery of woody species was

proceeding slowly, we added a restoration component to evaluate inexpensive 

methods of woody-species establishment. Because least Bell’s vireos are one of the 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 231 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

primary target species intended to benefit from ecosystem recovery following A.

donax control, we focused our attention on the recovery of native woody species. As 

of fall 2000, A. donax treatments had been initiated and were in various stages of

completion on 43 km of the river. All of the control projects were initially funded for 

5 years of treatment. 

Methods

Treatment of A. donax

Based on our initial evaluation of methods, we selected foliar application of a 6 

percent glyphosate solution as the primary method of control (Giessow and Giessow 

1998). We used a second approach, the cut stem method, where there was a risk of 

overspray onto native species from foliar application. With the cut stem method, A. 

donax stalks are cut close to the ground and herbicide (100 percent solution) applied 

within two minutes of cutting. In solid stands of A. donax with only isolated small

willows (less than 5-cm diameter stems), the willows were cut back to 1 m in height, 

and plastic bags were placed over them prior to foliar herbicide application in an 

effort to foster post-treatment site recovery. A concerted effort was made with both 

methods to preserve as much native woody vegetation as possible within and

adjacent to treatment areas to enhance site recovery. These individuals provide seed

and propagules for vegetative reproduction. 

Approximately 8 km of river supporting 25 ha of A. donax was treated with

foliar application of herbicide between September and November 1997. Crews used 

4-wheeled all terrain vehicles (ATV) with a 190-liter (50-gallon) power sprayer and 

30-50 m of hose to access sites. Typically, one ATV was used to haul the pesticide 

and spray equipment, with crew members following on foot. When sites were over 

about 1.5 km from the access point, filtered creek water was used to mix with the 

herbicide. Use of the filtered creek water led to increased maintenance costs for the

sprayers, so the cost of travel was balanced against the maintenance costs. Because

the A. donax clumps were too dense to penetrate, coverage of central stems within 

large clumps was obtained by spraying from a 7-m aluminum extension ladder. To 

ensure that all stems were sprayed, patches larger than 10-15 m across were accessed

from 1-m wide trails cut through them. 

Because of access and cost considerations, the foliar treatment was not followed 

by cutting and/or mowing of dead stems. Three primary concerns developed

regarding leaving dead standing biomass: 1) the dead stems would be more likely 

than live A. donax to wash downstream and become lodged behind bridges, 2) dead

standing material is a fire hazard, and 3) the addition of so much dead material to the

river system would alter habitat for sensitive estuarine species such as tidewater 

gobies (Eucyclogobius newberryi). The concern for bridges was high because of past

problems at Camp Pendleton caused by vegetation, including A. donax, during flood

events. We subsequently addressed the fire hazard in part by manually knocking 

down strips of dead A. donax (leaving the stalks on the ground) to act as fire breaks. 

Removing and processing A. donax biomass can be difficult and costly. Huge

volumes of material are generated (up to 45 kg m-2

) when the biomass is removed

(A.H.B.M. Wijte pers. comm.). In addition, removal of biomass is not always

possible in remote sections of the river with little access or in areas where damage to 

habitats supporting federally listed endangered species would result. These concerns

232 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

led to the initiation in 1997 of experimentation with mechanical removal of the entire 

plant. Approximately 3 ha were treated in an initial mechanical control project in the

fall of 1997. Both stems and rhizomes were removed with a clamshell bucket on a

trackhoe or excavator and processed through a tub grinder to minimize potential for

resprouting and make the material re-usable as mulch. To avoid impacts to native 

species, this method was used only in solid stands where access without damage to 

native woody species was feasible. Starting in 1999, we were required by Air

Pollution Control District regulations to spray water for dust control as the biomass

was being ground. After extraction of the stalks and rhizomes with the track hoe,

laborers walked over the site and removed remaining pieces of rhizomes. Scattered 

resprouting from missed pieces of rhizome were either pulled by hand or treated with

herbicide. 

Resprouts from both the mechanical and cut stem methods were often retreated

with foliar spraying the following spring. This was done to prevent the A. donax

stalks from growing so tall that the less effective cut stem method was again required

in the fall to avoid overspray onto native species. Native woody vegetation 3 m or

taller can be avoided using mechanical removal, although A. donax left near those 

plants must be treated using cut stem or foliar herbicide application. 

Vegetation Monitoring 

Twenty-one transects were established in 1997 to monitor the effectiveness of the 

foliar A. donax treatment. The locations of all transects were recorded at sub-meter

accuracy using a Trimble Pro XR GPS (global positioning system) unit. The transects 

varied in length from 12 m to 39 m and extended 2-5 m beyond the A. donax patch

measured. Due to the small size of the mechanically treated area, only two transects

were established, both 50 m in length. 

Initial sampling was carried out from October to November 1997 after the first 

A. donax control treatments had been applied. Re-sampling was conducted in October 

1998, 1999 and 2000. Fall, rather than spring, sampling was done to minimize 

activities in least Bell’s vireo habitat during the breeding season. 

Aerial vegetation and ground cover were recorded every meter along the 

transect line using the point intercept method (Bonham 1989). At each point, a height 

pole was held vertically and the composition of the canopy cover described by 

recording the species of every plant that intersected the pole. Height of each plant 

contacting the pole was recorded to the nearest 0.1 m. The number and type of A.

donax resprouts (rhizome or branch) were recorded within a 1-m belt along the length

of the transect. Surface soil texture was measured for all transects using a 

sedimentation test in which soil aggregates were dissolved; and the proportions of 

sand, silt and clay were determined based on the quantity of soil settling during 

specified time periods. 

The percentage of absolute foliar cover was calculated by summing all of the 

interceptions for each species along a transect and dividing by the number of points 

on the transect. Species were grouped for analysis by life form (herb, tree, or shrub) 

and origin (native or exotic) into the categories A. donax, exotic herbs, native herbs, 

native shrubs, native trees, and no vegetation. The herbaceous categories included all 

non-woody species except A. donax. The dominant species within the exotic and 

native herb communities were determined from a tally of herb-species interceptions

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 233 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

within each treatment. Native trees included Salix lasiolepis, S. laevigata, S. 

gooddingii, and Populus fremontii, while native shrubs included Baccharis 

salicifolia, S. exigua, Vitis girdiana, Toxicodendron diversiloba, Malacothamnus 

fasciculatus, and Sambucus mexicana. The “no vegetation” category included all 

points with no live aerial cover. 

The initial vegetation composition (pre-treatment) at each site was estimated

from the immediate post-treatment data with the aid of aerial photographs taken at a 

scale of 1:3000 in October 1997. Pre-treatment shrub and tree cover was assumed to

be equal to post-treatment cover. Pre-treatment A. donax cover was based on the

physical presence of stalks on the transect. 

A linear regression model was used to evaluate the effects of treatment, 

floodplain position (active floodplain or terrace), and soil texture (expressed as 

percent sand) over time on absolute vegetation cover by life form. A. donax

resprouting (resprout density) over 2 years following foliar herbicide treatment was 

analyzed using linear regression. Annual changes in percent absolute foliar cover 

within treatments were evaluated using Student’s t-tests. 

Site Restoration 

A. donax was mechanically removed from a 25-ha site 11 km from the river mouth 

during fall 1999/winter 2000. A small-scale restoration experiment was carried out in

January 2001 within 30 5x5-m plots to test the effect of mulching on survival of 

cuttings planted to revegetate the site. Fifteen plots were covered with a 25 cm thick

layer of mulch consisting of chipped A. donax ranging from 1-15 cm in length.

Because stem and root fragments of A. donax rarely sprout when spread at this depth 

(Lawson and Smead 2001), we chose it as the mulch material. 

All 30 plots were planted with three pole cuttings and five whip cuttings. The 

pole cuttings were willows (S. exigua, S. lasiolepis, or S. gooddingii); whip cuttings 

consisted of two B. salicifolia and three willows. All plant material was collected 

locally. Pole cuttings 3-6 m in length were planted 2.5-3 m deep in holes that were 

drilled with a 3-m auger. Whip cuttings 50-125 cm long were planted 25-50 cm deep. 

The whip cuttings were scarified at the base to encourage root formation and planted 

using rooting hormone. Typically, the surface soils were moist from 20-30 cm but 

dry until the upper fringe of the groundwater was encountered 0.5-2 m deep. The 

bottom 30 cm of most holes was moist or saturated. A few locations were completely 

dry below the moist surface. Most holes were silty sand for the entire profile,

although some were coarse sand. Poles were watered before and after cuttings were 

planted with approximately 20 liters of water. 

Survival and canopy dimensions were measured after 2 years, in January 2003.

Canopy volume was calculated from the canopy dimensions using volume formulas 

based on the shape of the plant (Ludwig and others 1975 in Bonham 1989). Surface 

soil texture was measured and expressed as percent sand for each plot using the 

sedimentation test described above. 

The effect of treatment (mulch, no mulch) on survival and canopy volume of B.

salicifolia and of willows after 2 years was analyzed using linear regression with 

percent sand and distance from the active river channel as covariates. 

234 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

Results 

Response of A. donax to Treatment 

A wildfire burned 7 of the 21 foliar transects in September 1998 prior to the 1998 

sampling. To control for the effect of fire on our results, the burned and unburned

transects were analyzed separately in all years. 

Live A. donax foliar cover was reduced by more than 90 percent after 1 year and 

by almost 100 percent after 3 years with both the foliar spray and mechanical 

removal (table 1). While it is difficult to provide an exact measure of kill because

there is no practical way to measure the amount of live rhizomes underground, both 

methods effectively killed the plants, including rhizomes, with very little resprouting 

in subsequent years. 

In 1998 and 1999, all live A. donax cover on the burned transects was the result

of rhizome resprouting. In the unburned stands, A. donax stem density averaged 

0.04 m-2 in 1998 and 0.02 m-2 in 1999, while in the burned stands it averaged 0.06 m-2

in 1998 and 0.19 m-2 in 1999. Rhizome sprouting appears to have been stimulated by

the burns; however, the difference in resprouting between the burned and unburned

stands was not statistically significant (F = 1.56, df = 41; p = 0.21).

Table 1—Mean percent absolute cover by life form following foliar spray (n=14 transects),

foliar spray/burn (n=7 transects), and mechanical removal (n=2 transects). 

1997 1998 1999 2000

Foliar Treatment 

(df=3) 

A. donax 68.01 (20.0)2

 a 1.5 (4.1) b 0.0 (0.0) b 1.4 (4.3) b 

Exotic Herb 22.5 (18.6) 30.4 (22.0) 47.6 (31.6) 52.4 (47.8) 

Native Herb 28.1 (32.3) 44.3 (30.5) 27.7 (37.9) 26.9 (20.9) 

Native Shrub 15.4 (23.1) 6.4 (9.3) 9.6 (14.4) 11.3 (20.4) 

Native Tree 23.9 (20.7) 11.7 (11.1) 15.8 (20.1) 22.2 (24.4) 

No Vegetation 8.3 (17.4) 37.1 (22.1) 29.3 (23.4) 25.4 (14.1) 

Foliar 

TreatmentBurned (df=3) 

A. donax 92.2 (6.1) a 1.6 *3 (2.8) b 1.4 (2.5) b 0.0 (0.0) b 

Exotic Herb 9.3 (11.2) 1.8* (2.5) 63.1 (22.4) 88.0 (29.8) 

Native Herb 5.7 (6.7) 3.3* (2.4) 35.4 (13.3) 33.5 (13.6) 

Native Shrub 6.3 (14.9) 0.0* (0.0) 3.5 (7.5) 12.4 (23.5) 

Native Tree 22.5 (23.5) 0.0* (0.0) 7.2 (6.9) 11.4 (9.9) 

No vegetation 0.9 (2.3) 93.8* (6.1) 25.1 (15.9) 6.9 (6.0) 

Mechanical

Treatment (df=3) 

A. donax 100.0 (0.0) a 7.9 (11.1) b 0.0 (0.0) b 0.0 (0.0) b 

Exotic Herb 0.0 (0.0) a 147.1 (22.1) b 100.0 (13.9) c 93.2 (15.2) c 

Native Herb 0.0 (0.0) 52.9 (13.9) 23.5 (27.7) 53.0 (44.3) 

Native Shrub 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 2.0 (2.8) 

Native Tree 0.0 (0.0) 0.0 (0.0) 1.0 (1.4) 0.0 (0.0) 

No Vegetation 0.0 (0.0) 3.0 (4.2) 2.0 (2.8) 3.9 (5.5) 1

 Mean separation using Student's t-tests. Means within a row followed by the same letter are not

significantly different at p = 0.05. 2

 Values in parentheses are one standard deviation. 3

 Means followed by an * were from plots sampled 1 month after a wildfire. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 235 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

If A. donax is not killed completely by the herbicide, sprouting may occur at the 

leaf and branch nodes on the stems. In an earlier study, plants that by their aboveground appearance were thought to be dead resprouted over a year after treatment 

(Giessow and Giessow 1998); we saw this occasionally during this study as well.

These resprouts were easy to kill with one foliar herbicide application. Although we 

did not monitor individual plants, where leaf and branch node sprouts were detected

on transects none were found the year following re-treatment. In 1998, eight sprouts

were counted on two transects; they were treated, and none were detected in those 

transects in 1999. Likewise, five sprouts were detected on two different transects in

1999; by 2000 there were none. 

Plant Community Development 

The influence of floodplain position and soil texture on absolute vegetation cover by

life form was analyzed separately for the burned and unburned foliar-treated sites 

using linear regression. In the unburned foliar-treated transects, the only significant 

response to treatment other than the decrease in A. donax was an increase in native 

herb cover. Native herbs increased in cover as a function of two two-way interactions 

between year and floodplain position (F = 2.72; p = 0.06) and floodplain position 

and percent sand, which ranged from 63 to 98 percent (F=7.19; p=0.01), revealing

that floodplain position influenced native herb growth in certain soil types, but in a 

manner inconsistent across years. In 2000, but not in other years, the percent cover of 

native herbs decreased in coarser soils in the terrace but not in the active floodplain.

No significant changes in vegetation cover occurred on the burned transects other 

than the decline in A. donax cover. 

The post-treatment exotic herb community was dominated by Hirschfeldia 

incana in all years, followed by Conium maculatum, Melilotus alba, and 

Chenopodium album. Artemisia douglasiana was the most frequently encountered

native herb in all years except 1998, when it was the second most frequent. Phacelia

cicutaria was the next most frequently encountered native herb, followed by

Ambrosia acanthicarpa. While absolute cover of herbs varied by treatment and year,

the same species dominated. 

Cover of native trees and shrubs was initially low on all treated areas, and none 

of the changes over the course of the study were significant. However, while not 

statistically significant, absolute cover of both trees and shrubs on the unburned 

transects showed the same pattern, dropping from 1997 to 1998 and then increasing 

each year thereafter (table 1). We also noted interesting though not statistically 

significant differences in cover on different geomorphic surfaces. In each year, the

cover of native shrubs was higher (by 4-25 percent) on the active floodplain than on 

the terrace, while for native trees it was lower (by 6-15 percent) on the active

floodplain than on the terrace. 

With the mechanical treatment, in addition to the response by A. donax there 

was significant inter-annual variation in exotic herb cover but not in native herb

cover (table 1). Exotic herbs increased rapidly in cover the first year after treatment,

followed by a significant fluctuation between the second and third years. Native herb

cover increased post-treatment, but the change was not statistically significant. 

236 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

Site Restoration – Survival and Growth 

Distance from the active river channel to the restoration plots ranged from 30 to

320 m. Percent sand on the plots ranged from 23 to 84 percent. Survival of willow 

pole cuttings averaged 31 percent per plot (n=28 plots, s.d.=36), while for willow 

whip cuttings survival averaged only 1 percent per plot (n=28 plots, s.d.=6). Per plot

survival of the B. salicifolia whip cuttings averaged 28 percent (n=29 plots, s.d.=34). 

Linear regression modeling of the effects of mulching and covariates distance to

river and soil texture (percent sand) on survival significantly predicted willow pole

survival (r2

=0.28; p=0.05). Survival was nearly twice as high (37 percent) on

unmulched plots as it was on mulched plots (20 percent). Mulch treatment interacted 

with soil texture, such that within mulched plots willow survival decreased as percent

sand increased, while in plots without mulch there was no pattern between survival 

and soil texture (F = 3.27; p = 0.08). A significant model was also produced

describing B. salicifolia survival (r2

=0.50; p=0.05). B. salicifolia survival, however,

was unrelated to mulching and was a function of an interaction between distance to

river and percent sand (F=6.45; p=0.02). Survival was highest on finer textured soils 

farthest from the floodplain. 

The average canopy volume on the restoration plots was 1.4 m3

 (s.d.=1.2) for B.

salicifolia and 1.5 m3

 (s.d.=2.2) for the willows. Canopy volume was unrelated to 

mulch treatment, distance to river, and soil texture. 

Costs 

A. donax Treatment 

The costs for A. donax control are highly variable depending primarily on

accessibility of the site and density of the A. donax. The costs reported below exclude 

program management and reporting. For the foliar spray method supplemented with 

the cut stem method, our average annual costs per ha dropped by over 85 percent 

after the initial treatment and did not change significantly after that for retreatments 

(table 2). These costs did not include any manipulation or removal of the biomass.

The cost to create several firebreaks by knocking down dead standing biomass was 

almost $5,000 per ha of firebreak. This work had to be done by hand because the 

sites could not be accessed by mechanized equipment without damaging riparian

woodland habitat. 

Because of the problems with biomass management, the mechanical treatment

was never fully implemented as originally planned. The cost for mechanically 

removing and grinding about 14 ha of A. donax in 1998/1999 was approximately

$19,800 per ha. This includes estimated costs of treating the A. donax with herbicide 

using cut stem and foliar application around native vegetation within and around the

margins of the site and treating resprouts in the first spring, but does not include costs

for biomass management. Treatment costs for the follow-up annual treatments are

Table 2—Labor, herbicide and cost per ha for foliar treatment of A. donax. 

Labor Herbicide Total Treatment

(man hours/ha) (liters/ha) (cost/ha) 

Year 1 (initial treatment) 150 130 $9,900

Year 2 (1st re-treatment) 15 7 $1,350

Year 3 (2nd re-treatment) 12 2 $1,100

Year 4 (3rd re-treatment) 15 2 $1,200

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 237 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

estimated to be the same as for the foliar spray method, because with so little live A.

donax left, the primary costs are associated with getting to the site and searching for 

the few live stems present. 

Site Restoration 

The average labor cost per whip cutting was 0.32 hours and per pole planting was 

1.07 hours. The mulching of the plots required 26 hours of labor plus the cost of the 

equipment. The planting of 180 whip cuttings required a total of 58.25 hours: 10

hours for collecting cuttings and 48.25 hours for planting the cuttings. The planting 

of 115 pole cuttings required 123.5 hours: 30 hours for collecting the pole cuttings 

and 93.5 hours for planting. The time required for mulching is not included in either 

the pole or whip cutting calculations. Planning, report writing, and monitoring are not 

included in these estimates.

Discussion 

Efficacy of Control 

While mid-November is often cited as the date of onset of winter dormancy in A.

donax (Bell 1998), our experience, based on evaluation of plants in the field, is that it 

is often later, particularly near the coast. In addition, we have found that the plants 

are still susceptible to the herbicide after leaf yellowing begins when there is still

green tissue present. A site-by-site evaluation of dormancy can extend the limited

time available for foliar herbicide application. 

Both foliar herbicide application and mechanical removal that includes removal 

of the rhizomes are effective methods of A. donax control and allowed us to meet our 

objective of less than 5 percent cover in 5 years. With both methods we were able to 

minimize damage to existing native trees and shrubs. We found that foliar treatment

was less costly than mechanical and easier to accomplish in areas with poor access.

On the other hand, mechanical treatment resulted in the use of less herbicide, consistent with general Navy policy directing the minimization of pesticide use (U.S. Department of the Navy 2002). Our initial conclusions were that both methods were 

useful in specific situations; however, difficulties and costs associated with grinding 

and using the biomass caused us to discontinue the mechanical removal method in 

2001. Problems included spontaneous combustion of the piled, chipped biomass, presumably caused by increased decomposition of wetted material resulting from water

sprayed to control dust. Insufficient need for the mulch for landscaping or other uses 

on the Base led to a lack of use and build-up of the stockpiles. While resprouting is

low when the material is spread thin (Lawson and Smead 2001), it can be high in

large stockpiles, creating a need for further treatment. 

The two treatment methods affect A. donax plants differently. With foliar spray

treatments, very high kill of the aboveground biomass is achieved with little subsequent resprouting. After the foliar spray treatment, the A. donax plant tissue dies

gradually over several months. We hypothesize that this gradual death is not accompanied by rhizome resprouting because the hormones that suppress resprouting are 

still present in the stems and are only gradually declining. Based on elevated stem 

counts on transects where the stems were removed by wildfire compared to unburned

transects, such an effect may last for a year. However, some resprouting can occur on 

238 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

foliar-sprayed plants with intact stems that have had no live aboveground biomass for

over a year, so follow-up surveys and treatment are needed. 

The low degree of resprouting following foliar spray contrasts with the rapid

resprouting response to other methods where the stems or rhizomes are cut. There are 

also differences between the cut stem and mechanical treatments with respect to 

resprouting. When plants are pulled out of the ground during mechanical removal, 

many of the rhizome fragments resprout quickly. These resprouts are easily killed 

with herbicide application, as evidenced by the reduction in cover from 7.9 percent to 

0 percent between 1998 and 1999 (table 1). Resprouting from entire rhizome masses 

left after the cut stem method requires repeated treatments over several years to kill.

In an earlier study, after 2 years of annual treatment over 1.3 stems m-2 of live A.

donax remained (Giessow and Giessow 1998). At that time, the stand was foliar 

sprayed to meet regulatory requirements. 

Regardless of the method used in the initial treatment of A. donax, retreatments

require approximately 12-15 hours of labor per ha of initially treated material. This

effort includes both searching for resprouts and treatment time. It is important to note 

that the area searched during retreatments may be an order of magnitude larger than

the initially treated area because retreatments cover all suitable habitat within the 

reach of the river in which the initial treatment occurred. Although this is a coarse 

estimate, it is sufficient for developing cost estimates for the entire treatment cycle

and making management decisions. Once the first 4 consecutive years of retreatment 

have occurred, it is advisable to revisit the site less frequently to allow A. donax to 

grow to a noticeable size and optimize retreatment expenditures. 

The importance of understanding riparian zone physical and ecological 

processes to produce consistent, successful riparian restoration has been noted 

(Goodwin and others 1997). This is also true of exotic plant eradication. To achieve 

watershed scale eradication of A. donax, all sites with the target exotic must be 

treated, not just the sites most likely to support a desired plant community within a

specified time frame. Otherwise, A. donax will reinvade downstream sites after 

flooding, compromising natural and restored habitats (Stein 1998). The beneficial 

effects of A. donax removal on a site with low flood return interval may be indirect

and take the form of removing the potential for reinfestation at another site. 

Plant Community Recovery 

While not statistically significant, we detected trends in the response of native shrubs 

and trees to removal of A. donax. We believe that the decline from 1997 to 1998 was 

a result of flooding. Rainfall at the Lake O'Neill weather station (approximately 2 km 

from the study site) in the winter of 1997/1998 measured 802 mm, more than double 

the 125-year average of 351 mm (Office of Water Resources, Camp Pendleton 

unpubl.5

 data). Flood damage to native trees and shrubs was seen on a number of

plots (J. Giessow pers. obs), and similar floods in the local vicinity had been reported 

to destroy up to 40 percent of standing vegetation (Hawkins and others 1997). After 

the initial decline, rainfall measured 218 mm in 1998/1999 and 225 mm in 1999/2000 

(Office of Water Resources, Camp Pendleton unpubl. data), and native trees and 

shrubs showed a steady increase in cover, although the time frame was insufficient 

 

5

 Unpublished data on file at the Office of Water Resources, Marine Corps Base, Camp Pendleton, CA

92055 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 239 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

for cover to equal or exceed the initial cover. The decline and subsequent increase in 

shrub and tree cover probably reflect an interaction of flood damage and increased

moisture availability for plant growth resulting from the A. donax control (Dudley

2000) and flooding (Kus 1998). As would be expected, previously existing trees and

shrubs often showed an increase in canopy cover in areas that had supported dense A. 

donax stands. Plants adjacent to sites, notably Vitis girdiana (wild grape), spread into 

treated areas and can produce significant amounts of cover (Giessow and Giessow 

1998). 

New herbaceous cover developed much more rapidly on the mechanically

cleared sites than on the foliar sprayed sites, and exotic herbs generally accounted for 

much of the difference in cover between the treatments. After mechanical removal, 

the sites were characterized by bare, loose soil—good conditions for weedy species

establishment (Hoshovsky and Randall 2000). The foliar-treated A. donax stands 

developed cover much more slowly. The A. donax litter cover provided few sites 

where seedlings could germinate and survive. However, the herbs that did establish

tended to be large individuals, presumably because they had less competition from

neighboring plants. 

Understanding site potential is key to predicting and evaluating recovery. While 

woody species occur on all three geomorphic surfaces, their presence and ability to 

establish is controlled by the potential for flood damage as flood return interval 

becomes shorter and by lack of establishment events as flood return interval gets 

longer (Harris 1987, Hawkins and others 1997). As a result, not all sites occupied by

A. donax have the same potential natural community. Soil and hydrologic variation in 

riparian areas results in mosaics of plant communities (Platts and others 1987).

Expectations for system recovery after A. donax eradication must take into account

that events that provide bare saturated soil suitable for willow recruitment are 

uncommon on the terraces, and even on the active floodplain they may occur with a 

frequency of only once in 10 years. While the active floodplain is the most suitable 

place for woody riparian species recruitment (Else 1996), the terraces support both 

riparian woodland and riparian scrub habitats (Smith and Lichvar 1999). If woody 

habitats are desired in the terraces after A. donax control, active restoration may be 

needed to achieve this goal within a short time frame. 

Active Restoration 

Our experiments evaluating methods to improve the survival of cuttings showed that 

mulching was counterproductive for establishing willows, whose survival was nearly

twice as high on plots with no mulch. Mulching had no apparent effect on the

survival of B. salicifolia or on the ultimate canopy volume of either this species or 

willows. Soil texture influenced survival of both willows and B. salicifolia. For 

willows, the effect was a simple one of lower survival on coarser soils. For B.

salicifolia, however, survival was highest on finer textured soils farthest from the

floodplain. Finer textured soils have higher available soil moisture (Brady 1974), 

which could explain higher survival in these soils. It is possible that unrecorded

variation in other factors, such as microtopography at the study site, had an effect on

factors important for survival, such as water availability. 

Where natural processes are not expected to lead to a desired plant community 

within a specified time frame, our results show the use of cuttings to be an option to

speed recovery, without the expense of planting and irrigating container stock.

240 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

Planning the installation of cuttings when high rainfall is predicted may be a feasible 

way of increasing survival and thereby decreasing unit costs. Because the plant 

material takes relatively little time to collect and is done at the time of planting, less 

advance planning is necessary, providing more flexibility in scheduling this activity. 

Considering the episodic nature of establishment and growth of woody riparian 

species and its link to flooding (Else 1996, Kus 1998), such a strategy may result in 

more canopy development over a specified time frame. 

Identifying and Tracking New Invasions 

The most neglected part of managing new invasions is the identification of new 

problem species at the earliest possible time. Identifying, tracking, and treating new

invasions so that one problem exotic species is not replaced by another is extremely

important (Hoshovsky and Randall 2000). While this study focused on A. donax

control, Camp Pendleton's program has, in fact, identified 14 species to be controlled. 

Ecosystem managers need to know when new invasive species appear and then

quickly develop and implement strategies to treat them. Watershed-wide coordination

is critical. Having a large-scale program in operation makes it much faster and easier 

to tackle new exotic species that threaten sites where A. donax has been removed. 

This was demonstrated with two new invasive exotic species found on the Santa 

Margarita River: German ivy (Delairea odorata) was identified on approximately

1.5 ha of land on a tributary to the Santa Margarita River, and perennial pepperweed

(Lepidium latifolium) was found over a larger area in the watershed in 2000. Because 

of existing funding and close coordination between the Marine Corps, Mission 

Resource Conservation District, and private landowners, treatment of both species 

was initiated within a month of discovery. 

In the midst of a complicated large-scale task such as the A. donax program, the 

subsidiary task of looking for new species can be overlooked. In addition, the 

resources used to survey can be hard to justify when no new infestations are found, 

even though the increase in control costs from delayed detection can be substantial

and exceed the cost of surveys (Hoshovsky and Randall 2000). While detection of 

new exotic species should draw on existing monitoring work, it needs to be an 

independent task with multiple visits every year initially and every few years in the

long term. It should include vouchering and review of status. The review could be as 

simple as comparison with the California Exotic Pest Plant list (CalEPPC 1999).

Exotic species that are not on the list should be evaluated to determine if they are

recent escapes and a decision made as to whether to carry out localized control, no 

treatment, or full eradication. 

The potential for new exotic species to invade and the detection and control of

all upstream occurrences of A. donax must be addressed. A watershed-based

organization that encompasses both public and private landowners is needed to 

coordinate watershed-based control over the long term and develop a program to

detect new invasions before they spread. A Weed Management Area has been

established for the Santa Margarita River watershed that addresses these issues

(California Department of Food and Agriculture, Santa Margarita Watershed WMA: 

http://smslrwma.org). 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 241 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

Conclusions 

The impetus for A. donax eradication in the Santa Margarita River is based on the 

premise that its removal will result in an increase in riparian cover and functional 

values for endangered species, as well as an avoidance of future habitat losses if A. 

donax were to spread throughout drainages (ACS/ES, Camp Pendleton 1994, Babbitt

and others 1995). A. donax eradication is feasible, and although much has been 

accomplished, substantial funding is needed to complete the job. Strategically timed

and placed restoration and long term exotic species management, including detection 

and control, can insure that the ecosystem-recovery benefits of A. donax eradication 

efforts are realized in a cost-effective manner. 

The temporal and spatial scales selected to evaluate recovery and to plan and 

execute post control restoration will determine to a large extent whether A. donax

control is judged successful in improving riparian habitat for endangered species.

Short-term expectations for succession at A. donax control sites must be balanced 

with the restoration of the natural functioning of the entire system. 

Recovery objectives should be defined for specific management units. To

effectively allocate resources, feasible objectives, including a time frame and 

measurable criteria, must be stated as clearly as possible from the beginning (Briggs

1996). 

Management units should be based on geomorphic surfaces. Sites within units

should be prioritized according to their potential for unassisted recovery. Sites with 

high potential may be a lower priority for scarce restoration funds than sites with 

lower potential. A site on a terrace that has the hydrology to support woody riparian

species but a low occurrence of suitable natural establishment events could be a

target for planting of cuttings. 

The time frame to judge success should be tied to flood events of specific 

magnitudes. The criteria to evaluate restoration success should take patch size into 

consideration and reflect the natural mosaics of plant communities in the 

management unit. 

References 

ACS/ES Camp Pendleton. 1994. Biological assessment: Riparian and estuarine habitat 

Marine Corps Base Camp Pendleton. Marine Corps Base Camp Pendleton, Office of

the Assistant Chief of Staff, Environmental Security. 

Babbitt, B.; White W.; Kobetich, G.C. [Letter (1-6-95-F-02) to Major General C.W. Reinke]. 

1995 October 30. 61 pages. Located at: U.S. Fish and Wildlife Service, Ecological 

Services Office, Carlsbad CA. 

Bell, G. P. 1998. Ecology and management of Arundo donax, and approaches to riparian

habitat restoration in southern California. In: Brock J.H.; Wade M.; Pysek P.; Green 

D., editors. Plant invasions: Studies from North America and Europe. Leiden, The 

Netherlands: Backhuys Publishers; 103-113.

Bonham, C.D. 1989. Measurements for terrestrial vegetation. New York: John Wiley and

Sons, Inc.; 338 p. 

Brady, N.C. 1974. The nature and properties of soils. 8th edition. New York: Macmillan

Publishing Co.; 639p. 

242 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Santa Margarita River A. donax Control Project—Lawson, Giessow, and Giessow

Briggs, M. K. 1996. Riparian ecosystem recovery in arid lands: Strategies and references.

Tucson, AZ: The University of Arizona Press; 159 p. 

California Exotic Pest Plant Council (CalEPPC). 1999. The California Exotic Pest Plant 

Council list: Exotic pest plants of greatest ecological concern in California. Located 

at: http://www.cal-ipc.org. 

Cummins, K.; Zedler P.H. 1998. Riparian vegetation mapping at Marine Corps Base

Camp Pendleton. Geo Insight International, Inc., Ojai, CA. 42 p. Available from 

ACS/ES Marine Corps Base Camp Pendleton. 

Douthit, S. 1993. Arundo donax in the Santa Ana River basin. In: Proceedings of the 

Arundo donax workshop; 1993 November 19; Ontario, CA. Team Arundo; 7-9.

Dudley, T.L. 2000. Arundo donax. In: Bossard, C.C.; Randall, J.M.; Hoshovsky, M.C.,

editors. Invasive plants of California's wildlands. Berkeley: University of California

Press; 53-58. 

Else, J.A. 1996. Post-flood establishment of native woody species and an exotic, Arundo 

donax, in a southern California riparian system. San Diego, CA: San Diego State 

University; 73 p. M.S. thesis. 

Giessow, J.A.; Giessow, J.H. 1998. 1998 annual report: Arundo donax removal project,

Santa Margarita River, CA. November 1998. San Diego: Department of Biology, San

Diego State Univ. 51 p. Available from ACS/ES Marine Corps Base Camp Pendleton. 

Goodwin, C.N.; Hawkins, C.P.; Kershner J.L. 1997. Riparian restoration in the western 

United States: Overview and perspective. Restoration Ecology 5(4S): 4-14. 

Harris, R.E. 1987. Occurrence of vegetation on geomorphic surfaces in the active 

floodplain of a California alluvial stream. American Midland Naturalist 118: 393-405.

Hawkins, C.P.; Bartz K.L.; Neale C.M.U. 1997. Vulnerability of riparian vegetation to 

catastrophic flooding: Implications for riparian restoration. Restoration Ecology 5:

75-84.

Hoshovsky, M.C.; Randall, J.M. 2000. Management of invasive plant species. In: Bossard, 

C.C.; Randall, J.M.; Hoshovsky, M.C., editors. Invasive plants of California's wildlands. 

Berkeley: University of California Press; 53-58. 

Kus, B.E. 1998. Use of restored riparian habitat by the endangered least Bell's vireo

(Vireo bellii pusillus). Restoration Ecology 6: 75-82. 

Lawson, D.M.; Smead, H. 2001. Evaluation of chipped Arundo biomass as mulch. In: 

Kelly, Mike, editor. Proceedings, California Exotic Pest Plant Council symposiums. 

Volume 6: 2000, 2001, 2002; 64-76. 

Lovich, R.E.; Fisher R.N.; Ervin E.L. 2001. Macroinvertebrate community composition of 

Arundo donax in a southern California riparian habitat. Unpublished draft supplied

by author. 

Ludwig, J.A.; Reynolds J.F.; Whitson P.D. 1975. Size-biomass relationship of several

Chihuahuan desert shrubs. American Midland Naturalist 94: 451-461. Cited in: 

Bonham, C.D. 1989. Measurements for terrestrial vegetation. New York: John Wiley and 

Sons, Inc.; 338 p. 

Platts, W.S.; Armour, C.; Booth, G.D.; Bryant, M; Buffort, J.L.; Cuplin, P.; Jenson, S.;

Lienkaemper, G.W.; Minshall, G.W.; Monsen, S.B.; Nelson, R.L.; Sedell, J.R.; Tuhy, J.S. 

1987. Methods for evaluating riparian habitats with applications to management.

Gen. Tech. Rep. GTR-INT-221. Ogden, UT: Intermountain Research Station, Forest

Service, U.S. Department of Agriculture; 177 p. 

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Stein, E.D. 1998. Invasive weed control as mitigation—a shifting paradigm. In: 

Proceedings of the arundo and saltcedar workshop; 1998 June 17; Ontario CA. California

Exotic Pest Plant Council. 

Smith, R.D.; Lichvar, R. 1999. Planning level delineation and functional assessment of

riparian ecosystems at Marine Corps Base, Camp Pendleton. U.S. Army Engineer

Research and Development Center (USAERDC), Waterways Experiment Station,

Vicksburg, MS.; 79 p. Available from ACS/ES Marine Corps Base, Camp Pendleton. 

U.S. Department of the Navy. 2002. OPNAVINST 5090.2. Environmental and natural

resources program manual. Washington, DC.

Wijte, Antonia H.B.M., Professor of Wetlands Plant Biology, California State University at

Long Beach. [Telephone conversation with Dawn M. Lawson], 21 June 2000. 

244 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Diptera Community Composition and 

Succession Following Habitat Disturbance 

by Wildfire1

Michael A. Patten, Jutta C. Burger, Thomas R. Prentice, John T. 

Rotenberry, and Richard A. Redak2

Key words: coastal sage scrub, diversity, flies, postfire, richness 

Extended Abstract 

Introduction 

Both biogeographic (for example, latitude) and local (for example, soil) processes

determine composition and succession of biotic communities. Postfire succession of 

vegetation has been studied intensively in chaparral and coastal sage scrub. Fewer

studies have examined postfire succession of animals, even though fires can

drastically alter their abundance and diversity (Ahlgren and Ahlgren 1960, DeBano 

and others 1998). Work on response to fire has focused on vertebrates, with few 

studies on insects, several yielding conflicting results. Little is known about 

community ecology of Diptera. Flies exhibit high alpha-diversity, show remarkable

variation in foods and habitats (occupying every trophic level), and are highly vagile 

(recolonization is likely to occur quickly). Studies of arthropod succession are 

important for understanding mechanisms that determine community structure, which 

can be critical to land management, conservation, and reserve design (Kremen and 

others 1993). 

We examined Diptera community differences between burned and unburned

sites at family and guild levels. We focused on a mid-successional period (2.5 to 4 yr

after a burn). Full recovery of burned sage scrub requires 5–10 yr and may never 

equal the preburn state (Westman 1981). We hypothesized that (1) Diptera 

communities differ qualitatively between burned and unburned plots; (2) 

recolonization occurs in a predictable order of scavengers, animal feeders (predators, 

parasitoids, and hematavores), plant feeders (pollinators and herbivores), and 

detritivores; (3) local-scale processes drive short-term vegetation recovery because

many sage scrub shrubs reestablish from rootstock and seeds that survive fire; and (4) 

geographic-scale processes drive Diptera community reestablishment because flies

recolonize from surrounding intact areas, not from the disturbed site itself. 

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2

 Department of Biology, Department of Entomology, and Center for Conservation Biology, University

of California, Riverside, California 92521. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 245 
Poster session—Diptera Succession After Wildfire—Patten and others

Methods 

We sampled 12 burned and 12 unburned plots (50-m diameter) in coastal sage scrub 

at the Southwestern Riverside County Multispecies Reserve at Lake Skinner,

California, following a 1993 fire. We quantified vegetation using standard pointintercept and line-transect methods. We sampled arthropods at 3-month intervals 

from March 1996 to December 1997 using malaise traps (1 trap/plot/sampling

period), vacuums (5 transects/plot/period), and pitfalls (7 traps/plot/period). 

Specimens were identified to family and classified into foraging guilds. 

Fluctuations in family richness over time did not differ between unburned and 

burned plots (profile analysis; P > 0.15, n = 8 samples), so data were pooled by site 

across sampling periods. Differences between unburned and burned plots in vegetation and Diptera were tested with MANOVA (multiple analysis of variance). The

association between vegetation and Diptera abundance was determined with Mantel 

tests. The pattern of Diptera recolonization was inferred using isotonic regression of 

mean percent correct classification from discriminant function analyses (DFA) for 

each family (n = 36). Means and SEs were calculated by guild (n = 7). Our H0 was

discrimination of plots is equal across guilds; our ordered expectation HA was

detritivores ≥ pollinators ≈ herbivores ≥ predators ≈ parasitoids ≈ hematavores ≥

scavengers, with at least one strict inequality.

Results 

A total of 12,437 individual flies in 46 families were collected, with Chironomidae 

by far the most abundant family collected. Because of potential swamping, malaisecollected Chironomidae were excluded from analyses. The next best represented 

families were Cecidomyiidae, Anthomyiidae, Tipulidae, Empididae, Chloropidae, 

and Phoridae. 

Burned and unburned plots differed significantly in vegetation structure and 

composition (MANOVA, Wilks' Λ = 0.12, P < 0.01), but community composition of 

Diptera did not differ between plots at the level of family (Wilks' Λ = 0.07, P > 0.5)

or guild (Wilks' Λ = 0.48, P > 0.05). Nevertheless, there was a strong association 

between vegetation and Diptera families (Mantel standardized r = 0.28, P < 0.02) and 

guilds (r = 0.22, P < 0.05). Scavengers occupied burned sites with roughly equal frequency as unburned, so their abundance was a poor discriminator of plot type; detritivores were much more abundant on unburned sites and were better discriminators 

of plot type (table 1). Our data supported the recolonization sequence we predicted

for Diptera guilds (S = 6.38, P < 0.05), underscoring the unequal (but predictable) 

recovery rates. 

Discussion 

Dipteran communities recovered quickly from impacts of fire, consistent with 

previous studies of overall insect abundance (Force 1981, Moya-Raygoza 1995) and

Diptera in particular (Delettre 1994). Burned sites were indistinguishable from 

unburned sites in overall community, with common generalist feeders and scavengers 

equally represented; Lomonaco and Almeida (1995) reported the same pattern. 

Similarity across plots resulted from similarity in abundance of dominant fly 

246 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Poster session—Diptera Succession After Wildfire—Patten and others

families; presence/absence data may yield different interpretations of community 

equality (Patten and Rotenberry 1998) but an ordination with such data produced

comparable results in our case. 

We detected a distinct, predictable pattern of guild recolonization following a 

successional sequence from scavengers to animal-feeders to plant-feeders to detritivores. Our results support the hypothesis that the key to community structure is extrinsic biogeography rather than intrinsic local processes (Cornell and Lawton 1992). 

Local processes play an important role (abundances of some families differed

markedly). These processes probably include microhabitat (Tolbert 1975) and relative 

amounts of shrubs and herbs (Bährmann 1984), which affect guilds (and thus 

families) differently. 

Table 1—Percent correct classification of burned and unburned plots using discriminant

function analysis. Means were calculated across families within each guild (thus the different

sample sizes for each guild). The ordered expectation of isotonic decrease is significant (0.01 

< P < 0.05).

Diptera guild Mean % (± SE) 

Sample size correctly classified per guild (ni)

detritivore 60.1% ± 2.6 14

pollinator + herbivore 54.2% ± 3.0 9

predator + parasitoid + hematavore 50.4% ± 4.7 12 

scavenger 45.1% ± 6.8 11 

Acknowledgments 

M.L. Buffington, K.A. Campbell, T. Hosch, E.E. Porter, and C. Slotnik assisted in 

the field and lab. C.E. Dunning and the Rotenberry lab provided helpful criticism of 

this presentation. The Metropolitan Water District of Southern California (agreement 

no. 4602) funded this work. 

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248 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Patterns and Processes of Arthropod 

Community Succession after a Fire1

Jutta C. Burger2, Michael A. Patten2,3, John T. Rotenberry3,4, and 

Richard A. Redak2,4

Key words: coastal sage scrub, detritivores, diversity, herbivores, predators 

Extended Abstract 

Introduction 

At present little is known about arthropod recolonization and community succession

after disturbance. Arthropods are critical in ecosystems as members of food webs and 

accelerators of nutrient cycling. The degree to and speed with which they return to 

disturbed habitat is thus of great interest to researchers and land managers alike. 

We compared abundance and diversity of arthropods in mid-successional (after

a 1993 wildfire) and mature California coastal sage scrub (CSS), a vanishing 

ecosystem in southern California (Westman 1981) that is threatened primarily by

urban development (Davis and others 1994), associated increases in wildfire intensity 

(Minnich 1983), and invasion by non-native grasses (Alberts and others 1993). 

Our general purpose was to determine whether arthropod communities in midsuccessional habitat are qualitatively different from those in mature habitat.

Specifically, we investigated whether foraging guild structure, abundance, and

diversity differ in relation to fire history and whether there is evidence of failure in 

community recovery after a fire. Lastly, we compared our results for the entire 

arthropod community to those for just the Diptera, a highly volant insect order within 

which all guilds are represented. 

Methods

We collected arthropods from 12 burned and 12 unburned 50-m-diameter plots at the 

Southwestern Riverside County Multispecies Reserve by Lake Skinner, California.

Arthropods were sampled with pitfall (7 traps/plot/sampling period) and vacuum (5

transects/plot/sampling period) at 3-month intervals from March 1996 to December

1997. Vegetation structure and percent cover were sampled using standard pointtransect and line-intercept methods. All arthropods were identified to morphospecies 

within family with the exception of the Apterygota, Acari, Collembola, Isopoda,

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2

 Department of Entomology, University of California, Riverside 3

 Department of Biology, University of California, Riverside 4

 Center for Conservation Biology, University of California, Riverside 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 249 
Poster Session—Arthropod Succession—Burger and others 

Psocoptera, Thysanoptera, Aphididoidea, and Coccoidea, which were counted. Each

arthropod taxon was assigned to a foraging guild (detritivore, flower visitor,

hematavore [blood feeder], herbivore, parasitoid, predator, or scavenger). 

Abundance and richness were compared between burned and unburned plots 

using t-tests with Dunn-Šidak corrections for multiple comparisons. Detrended 

correspondence analysis (DCA) was performed on abundance per site of all 

arthropods identified to morphospecies. Axis identity was determined by correlating 

plot scores on each axis (DC I and DC II) with vegetation variables. Families 

occurring only once were omitted from analysis. 

Results 

A total of 52,926 arthropods were collected with vacuum and pitfall samples. The 

distribution of arthropod guilds was more even in unburned than burned CSS. 

Proportionally more detritivores and herbivores were found in burned areas, and 

more flower visitors occurred in unburned sites. Overall, more arthropods were 

collected from burned sites, including significantly more detritivores and marginally

more herbivores and predators. Abundance of arthropod orders followed a similar 

pattern to that of guilds, with orders associated with detritivore and herbivore feeding 

guilds tending to be more abundant in burned sites, but only Psocoptera (detritivores) 

were significantly more abundant in burned sites. 

Species richness per site of the six dominant arthropod orders (Hymenoptera, 

Araneae, Homoptera, Diptera, Coleoptera, Hemiptera) was remarkably similar across

burned and unburned sites. Assemblages of families appeared to be qualitatively

different among burned and unburned sites. Detrended correspondence analysis of 

family abundances produced a relatively short gradient (length = 1.14; 4.0 represents 

complete turnover). However, sites were separated according to fire history along 

DC I and along a grass cover gradient on DC II (fig. 1). Families positively

associated with mature, unburned sites were Saturniidae (Lepidoptera), Corinnidae 

(Araneae), and Ptinidae (Coleoptera). Families associated with low grass cover were

Kinnaridae (Homoptera), Cucujidae (Coleoptera), and Dermestidae (Coleoptera). All 

orders strongly associated with DCA axes I and II were relatively uncommon in our 

samples. Therefore, their associations independent of other orders are tenuous. 

Conclusions 

Results suggest that arthropod guilds are more evenly distributed in mature vs. midsuccessional coastal sage scrub. Mid-successional sites supported a higher 

abundance, but not a higher richness, of arthropods. Much of this difference in

abundance was a result of higher numbers of detritivores in burned areas. Arthropod 

assemblages at the family level were associated with fire history and correlated with

degree of grass cover, primarily as a result of the distribution pattern of relatively rare

families (fig. 1). Grass invasion appeared to be affecting arthropod community 

composition in coastal sage.

The guild structure of all arthropods in burned sites differed from that of just the 

Diptera (see Patten and others, this volume). Taxa that are less mobile than Diptera

may be subject to vastly different conditions affecting recolonization. These include 

250 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Poster Session—Arthropod Succession—Burger and others 

purely local processes, such as the ability to survive and reestablish after a fire from 

populations on site or nearby rather than recolonizing solely via immigration. 

Acknowledgments 

M. Buffington, K. Campbell, T. Hosch, M. Misenhelter, E. Porter, T. Prentice, and C. 

Slotnick assisted in collecting arthropods in the field and identifying them in lab. M. 

Al-Siddiqi, E. Bonilla, N. Davis, M. Escobar, T. Gaines, S. Kensinger, and L. Trong 

assisted in sorting arthropods and data entry. J. Bethke provided technical assistance.

C. Dunning provided helpful comments. This project was partially funded by a grant 

from Metropolitan Water District (agreement 4602).

Figure 1—Detrended correspondence analysis of arthropod assemblages on burned

and unburned sites. The dashed line is a heuristic to show the near complete simple 

separation of burned and unburned plots based on arthropod community 

composition. 

References 

Alberts, A.C.; Richman, A.D.; Tran, D.; Sauvajot, R.; McCalvin, C.; Bolger, D.T. 1993. 

Effects of habitat fragmentation on populations of native and exotic plants in

southern California coastal scrub. In: Keeley, J.E., editor. Proceedings of the

symposium on the interface between ecology and land development in California. Los 

Angeles, CA: Southern California Academy of Sciences; 103-110. 

Davis, F.W.; Stine, P.A.; Stoms, D.M.1994. Distribution and conservation status of coastal

sage scrub in southwestern California. Journal of Vegetation Science 5: 743-756. 

Minnich, R. 1983. Fire mosaics in southern California and northern Baja California. 

Science 219: 1287-1294. 

Westman, W.E. 1981. Factors influencing the distribution of species of Californian

coastal sage scrub. Ecology 62: 439-455. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 251 

Monitoring the Effects of Natural and 

Anthropogenic Habitat Disturbance on the 

Ecology and Behavior of the San Diego 

Coast Horned Lizard (Phrynosoma 

coronatum blainvillei)

1

Tandora D. Grant and Allison C. Alberts2

Key words: body size, foraging preference, habitat disturbance, home range size, sage 

scrub 

Extended Abstract 

The sage scrub communities in which San Diego coast horned lizards (Phrynosoma

coronatum blainvillei) typically occur are now restricted to only 10 to 15 percent of 

their former range and are currently subject to rapidly accelerating rates of loss. As a

result of urban and agricultural development, pesticide use, collecting for the pet 

trade, off-road vehicle activity, and other human disturbance, San Diego horned

lizards have undergone serious population declines in the last decade. In some areas,

introduced exotic ant species have displaced the native harvester ants on which 

horned lizards depend. Because relatively little is known of the biology of coast 

horned lizards, basic information on their habitat and resource requirements, as well 

as their behavior and reproductive physiology, is critical to developing effective 

management plans for their survival. 

Although San Diego coast horned lizards appear able to persist in semidisturbed habitats, the extent to which these areas can support viable populations over

the long term is unknown. To address this information gap, we compared several

aspects of the behavioral ecology of the San Diego coast horned lizard between 

undisturbed Riversidian sage scrub and habitat disturbed by either fire or historic 

grazing on the Southwestern Riverside County Multi-Species Reserve during 1995

and 1996. 

In 1996, we radio-tracked 11 male and 9 female horned lizards. Once located, 

each lizard’s position was recorded to + l m using a GPS (global positioning system) 

equipped to receive real-time differential corrections. For the 10 males and 4 females

for which enough data points were available to measure home range size reliably,

home range size was estimated using two methods: the convex polygon method and a

grid cell technique that produces a two-dimensional frequency distribution describing 

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California Polytechnic University, Pomona, CA. 2 Center for Reproduction of Endangered Species, Zoological Society of San Diego, PO Box 120551, 

San Diego, CA 92112

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 253 
Poster Session—Ecology of San Diego Coast Horned Lizard—Grant and Alberts 

spatial utilization of the home range. While there was no significant difference in the

home range size of adult males and females, there was a tendency for convex 

polygon home ranges in disturbed habitat to be larger than those in undisturbed 

habitat (U=10, df=12, P=0.15). Using a grid size of 10 m2

, differences in the spatial 

utilization of home ranges in disturbed and undisturbed habitat were significant 

(U=5.5, df=12, P=0.04). In addition, lizards in disturbed habitat overlapped more 

conspecifics on average than those in undisturbed habitat (U=6, df=12, P=0.04). 

To document foraging preferences of horned lizards, we dissected scat from 

disturbed and undisturbed habitats on the reserve and identified the species consumed

based on head morphology. In both the number of ant species consumed (T=2.49,

df=32, P=0.02) and total number of prey types (T=2.18, df=32, P=0.04), horned 

lizards in undisturbed habitat exhibited richer diets than those in disturbed habitats 

(fig. 1). 

Figure 1. 

254 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Poster Session—Ecology of San Diego Coast Horned Lizard—Grant and Alberts 

We measured snout-vent length (SVL) and body mass of 27 adult males, 25

adult females, and 206 juveniles in order to compare morphology of horned lizards in

disturbed and undisturbed habitats. Using a two-factor analysis of variance with sex

and disturbance level as independent variables, we found that individuals of both 

sexes from undisturbed habitat were larger in SVL (F = 13.68, df = 1, 48, P = 0.0006)

and body mass (F = 14.03, df = 1, 48, P = 0.0005) than those from disturbed habitat. 

In contrast, analysis of covariance indicated that there were no differences in growth 

of hatchlings in SVL (F = 1.40, df = 1, 184, P = 0.24) or body mass (F = 1.59, df = 1, 

184, P = 0.21) between disturbed and undisturbed habitats.

In summary, horned lizards in disturbed habitats utilize larger home ranges with 

lower plant diversity and greater overlap among individuals. Increased home range

size may represent a compensatory response to the lower species richness of insects,

particularly ants, found in disturbed habitats. Horned lizards in disturbed habitats exhibit smaller adult body size than those in undisturbed habitats. That growth rates of 

hatchlings during their first season are similar across habitats suggests that differences in adult size are not the result of differences in early growth of hatchlings. 

Rather, these differences may be due to differential growth later in the maturation 

period or to differences in predation pressure. The initial response of horned lizards

to habitat disturbance appears to be behavioral rather than physiological. The greater

spatial requirements of horned lizards in disturbed habitats should be considered in 

determining the area needed to support a viable population over the long term. 

Acknowledgments 

We are grateful to the Metropolitan Water District and to the Reserve Management

Committee for their sponsorship of this project. We would also like to thank Lori 

Jackintell, Jeff Lemm, Monica Olivera, and Catherine Wertis for invaluable field 

assistance. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 255 

Fire Management in Some California 

Ecosystems: A Cautionary Note1

Hartmut S. Walter, Teresa Brennan, and Christian Albrecht2

Key words: Bishop pine, chaparral, closed-cone conifer, mixed-conifer forest, 

regeneration, stand renewal

Extended Abstract 

Introduction 

Fire has been recognized as a natural and important physical factor in many ecoregions of North America. We wish to point out that our understanding of the biocomplexity of our natural ecosystems is far from complete; in particular, the role of fire in

vegetation succession and ecosystem health deserves more scrutiny where biodiversity conservation is a primary or major goal of management. We present four case

studies from southern California that are evidence of species and community persistence and renewal in the absence of wildfire or in the presence of low frequency fire

events. A simplistic application of the “fire is good and necessary” paradigm may put

certain taxa and habitats at risk. 

Study Areas and Methodology

1. Santa Cruz Island off Santa Barbara contains three distinct stands of Bishop pine 

(Pinus muricata remorata). This pine species has long been considered to belong to

the group of closed-cone and obligate fire conifers (Vogl and others 1977). The present study searched for evidence of spontaneous cone opening, germination, and 

stand renewal in the absence and presence of fire. Field observations were made on 

live and dead trees regarding pine cone status. Cones were also collected and monitored for opening and seed release. Seeds were planted in commercial soil and mulch 

in the open air. Study plots on the island were monitored as to the number of seedlings per hectare. The opportunity to assess the pine’s response to a hot fire arose 

when a prescribed fire event went out of control in 1994 (Walter and Taha 2000). 

The research hypothesis was defined as Bishop pines of Santa Cruz Island do not 

require periodic fire events for seed germination and stand regeneration. 

2. The North San Diego region in the Carmel Valley near Del Mar is unusually

rich in endemic plant and animal taxa. Monitoring and conservation evaluation of 

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2 Department of Geography, University of California, Los Angeles; P. O. Box 951524, Los Angeles, CA 

90095-1524

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 257 
Poster Session—Fire Management—Walter and others 

several undisturbed chaparral parcels at and near Carmel Mountain was carried out in 

1995-96. Dozens of vegetation transects were made, comparing burned and unburned 

chaparral patches (Walter 1996). The research hypothesis was worded as The unusual

arborescence, composition, and structure of this chaparral ecosystem are a result of

a long absence of fire. 

3. The James San Jacinto Mountains Reserve near Idyllwild (Riverside County)

is the site of a long-term post-fire succession study (Walter, in preparation) at the 

chaparral-mixed forest ecotone. A 5x80-m transect was established on completely

burned ground at an elevation of 5,800 feet after a major wildfire in August 1974. 

Vegetation succession has been monitored for 25 years. The initial research

hypothesis tested was Post-fire succession includes a short-term chaparral stage of 

circa 15-20 years duration followed by mixed forest regeneration and dominance. 

4. The James Reserve is surrounded by a complex, unburned forest community

near Hall Canyon. This mixed conifer-oak forest is species-rich and appears to

contain a number of different age cohorts. For this study, we censused three slope

transects of 1x100-m at 5,500 feet elevation for species abundance and tree height.

The research hypothesis for this forest surrounding the James Reserve was This 

species-rich and mature forest does not require fire for stand renewal. 

Results 

Bishop Pine on Santa Cruz Island 

After the drought years of the early 1990s a significant portion of the island’s Bishop 

pines died. Subsequently, aided by sheep removal, seedling and sapling pines began

to replace the dead trees and expand from the remnant forest stands into surrounding 

grassland. In 1998 and 1999, detailed studies and experiments on cones of island 

Bishop pine (Walter and Taha 2000) revealed that (1) some cones open while still on 

the tree under a normal temperature regime, (2) all cones may release their seeds 

when exposed to normal summer temperatures and solar radiation, and (3) seeds

germinate easily and grow well in normal soil. 

In 1994, study plots near pine snags contained hundreds of healthy pine seedlings; their density reached more than 1,400 seedlings per hectare. The December

1994 hot wildfire event, however, killed most remaining pine trees as well as thousands of seedlings in the affected area. Five years later (1999), only a small number 

of seedlings (about 10/ha) were detected on the burn site. 

North San Diego Chaparral Ecosystem

Carmel Mountain and a few adjacent chaparral patches constitute rarely seen mature

and over-aged chaparral with dominant shrubs such as Ceanothus verrucosus, Xylococcus bicolor, Comarostaphylis diversifolia, Cercocarpus minutiflorus, and 

Adenostoma fasciculatum. In canyon bottoms, this community forms dense arboreal 

thickets because one or more recent fire events jumped across slopes without burning 

these habitats. 

On the high mesa of Carmel Mountain there is a shallow valley with open habitat dominated by shrubs, bunch grasses, rocky ground, and several small trees of mission manzanita (Xylococcus bicolor) attaining record heights of 5 to 7 m. Nearby, an 

258 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Poster Session—Fire Management—Walter and others 

old California sagebrush (Artemisia californica) has become woody, with a diameter

of about 15 cm. The vegetation must not have burned for ages, that is, over several 

historic fire cycles or generations. This ecosystem differs from frequently burned 

chaparral. It is likely quite stable and not senescent. 

Chaparral-Mixed Forest Ecotone Succession (San Jacinto

Mountains) 

After 25 years of succession, the chaparral community (Ceanothus leucodermis 

Arctostaphylos glauca) is still a dominant matrix and shows no sign of senescence.

Resprouting oaks (Quercus chrysolepis and Q. wislezinii) and sapling pines (Pinus

coulteri) are a minor part of the chaparral vegetation in and around the transect. It 

may take more than 100 years to bring back a mixed forest stand resembling the prefire vegetation at this ecotone location. 

Hall Canyon Forest Ecosystem (San Jacinto Mountains) 

The mixed conifer-oak forest of Hall Canyon, located on the east slope of the San Jacinto

Mountains, has been subjected to several natural mortality factors (drought, insects, and 

wind throw) over the past 25 years. However, it has not experienced a stand-burning fire 

in more than 200 years. An analysis of woody vegetation data (seedlings included) from

three 1x100-m transects shows a mixed forest dominated by canyon live oak, black oak, 

incense cedar, and sugar pine. The stand structure of Hall Canyon is complex, with 

dominant species having average heights of 0.099 m, 0.24 m, 1.05 m, and 0.53 m 

respectively. The low average height values obtained are representative of the large

number of seedlings and saplings observed and are also indicative of the varied age

structure within the forest. Preliminary research results for the mature conifer-oak forest

of Hall Canyon show significant regeneration of dominant tree species after episodes of 

severe drought, insect infestation, wind throw, and in the absence of fire. 

Discussion 

The continental distribution of the Bishop pine is peculiarly coastal and insular, 

suggesting that this species is actually persisting as a fire evader, although it cannot

persist in more interior regions where the occurrence of hot fires is more likely. We 

conclude, therefore, that fire is not a necessary agent for regeneration of the island 

Bishop pine populations, of the species as a whole, and possibly of other closed-cone 

conifers as well. The Santa Cruz Island Bishop pine can be categorized as a 

“pyrofugal” taxon (Walter and Taha 2000). 

Some North San Diego chaparral patches represent a unique biodiversity asset. 

We conclude that old and mature chaparral remains viable and species-rich in this 

coastal study region; there is a low incidence of natural fire events. Mature and firefree chaparral constitutes a unique habitat due to the quality of its structure as well as 

its species composition (lichen richness, and so forth). In light of the generally short 

successional cycles seen in California’s chaparral ecosystems, this habitat qualifies as

old growth or ancient chaparral in analogy to the old growth forests of the Pacific

Northwest. Any control burn of this habitat would eliminate a unique plant and 

probably also a corresponding animal community. Patches of ancient chaparral 

deserve priority conservation management. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 259 
Poster Session—Fire Management—Walter and others 

Our research hypothesis was not validated at our ecotone succession transect in

the San Jacinto Mountains. This surprising result provides evidence that long-term fire

cycles have in the historic past included between-fire intervals of 100 years or more 

resulting in an open conifer-oak forest with a minor understory of tall chaparral shrubs.

Cultural fire suppression should be ruled out as a contributing factor to extended firefree intervals in this part of California. The Hall Canyon forest history of the last 25

years offers further support for forest regeneration and rejuvenation in the absence of 

fire. Catastrophic wind throws and a severe and prolonged drought episode followed by 

insect infestations have opened up the forest canopy, created excellent breeding

opportunities for cavity-requiring vertebrates and sunlight for tree seedlings. 

References 

Agee, J.K. 1993. Fire ecology of Pacific Northwest forests. Washington, DC: Island Press. 

Aplet, N.J.; Olson, J.T.; Sample, V.A., editors. 1993. Defining sustainable forestry. 

Washington, DC: Island Press. 

Berg, K. 1982. Hall Canyon annotated floristic list. Riverside, CA: University of California, 

James San Jacinto Mountains Reserve, Natural Land and Water Reserves System. 

Carroll, M.C., Laughrin, L.L.; Bromfield, A.C. 1993. Fire on the California Islands: Does it

play a role in chaparral and closed cone pine forest habitats? In: Hochberg, F.G., 

editor. Third California Islands symposium: Recent advances in research on the California Islands. Santa Barbara, CA: Santa Barbara Museum of Natural History; 73-88. 

Hobbs, E. 1980. Effects of grazing on the northern population of Pinus muricata on Santa

Cruz Island, California. In: Power, D.M., editor. The California Islands: Proceedings of

a multidisciplinary symposium. Santa Barbara, CA: Santa Barbara Museum of Natural 

History; 159-165. 

Spurr, S.H.; Barnes. B.V. 1980. Forest ecology. 3rd edition. New York: John Wiley & Sons. 

Vogl, R.J.; Armstrong, W.P.; White, K.L.; Cole, K.L. 1977. The closed-cone pines and 

cypress. In: Barbour, M.G.; Major, J., editors. Terrestrial vegetation of California. New 

York: John Wiley & Sons; 295-358. 

Walter, H.S. 1977. Effects of fire on wildlife communities. In: Proceedings of the 

symposium on the environmental consequences of fire and fuel management in

Mediterranean ecosystems; 1977 August 1-15; Palo Alto, California. Gen. Tech. Rep.

WO-3. Washington, DC: Forest Service, U.S. Department of Agriculture; 183-192. 

Walter, H.S. 1996. Biological resource conservation and mitigation issues: Neighborhood 

8A, north city of San Diego. Los Angeles, CA: Research Report. 106 p.

Walter, H.S.; Taha, L.A. 2000. Regeneration of Bishop pine (Pinus muricata) in the 

absence and presence of fire: A case study from Santa Cruz Island, California. In: 

Browne, D.R.; Mitchell, K.L.; Chaney, H.W., editors. Proceedings of the fifth California

Islands symposium; 1999 March 29 to April 1; Santa Barbara, California. San Diego,

CA: U.S. Dept. of Interior, Mineral Management Service (OCS Study MMS 99-0038);

172-181. 

Wehtje, W. 1994. Response of a Bishop pine (Pinus muricata) population to removal of

feral sheep on Santa Cruz Island, California. In: Halvorson, W.L.; Maender, G.J.,

editors. The fourth California Islands symposium: Update on the status of resources.

Santa Barbara, CA: Santa Barbara Museum of Natural History; 331-340.

260 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Spatial and Temporal Variation in 

Ephemeral Pool Crustacean Communities1

Janette Holtz2, Marie A. Simovich, and Thomas Philippi 

Key words: habitat loss, rainfall pattern, species diversity, vernal pools

Extended Abstract 

Vernal pool habitat losses in San Diego County, California, are estimated at 

95 percent and will increase as development continues. The majority of San Diego’s

remaining pools are located at the Marine Corps Air Station, Miramar. Crustacean 

communities in eight vernal pools in San Diego County were sampled over 3 years 

that differed in rainfall amount and pattern. In 1994, rains occurred late but were 

moderate in quantity, and species richness ranged from 1 to 22 species; 1998 was a 

high rainfall El Niño year, and pools varied from 0 to 15 species; 1999 was a low 

rainfall La Niña year, and pools contained 4 to 7 species. No spatial pattern was seen

among neighboring pools. Longer-lived pools had more species, including the more 

rare species; smaller pools included the more common taxa (table 1). Diversity 

within pools varied between years, although the highest rainfall year did not result in

the highest diversity (fig. 1). These data indicate the necessity of surveying multiple

pools over a period of several years to establish baseline data for these complex 

communities. Therefore, adequate monitoring of restored or created pools must 

extend over several pools and several years. Certain species present within a cyst

bank may not hatch every year due to weather conditions or other factors. Some 

species hatch repeatedly after a filling but may not breed and replenish the bank. 

 

1

A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2 Department of Biology, University of San Diego, San Diego, CA; and Savannah River Ecology Lab, 

University of Georgia, Drawer E, Aiken, SC.

USDA Forest Service Gen. Tech. Rep. PS-GTR-195. 2005. 261 
Poster Session – Ephemeral Pool Crustacean Communities – Holtz and Others

Table 1—Number of species of vernal pool crustaceans in 4 different pools sampled in 3

different years (A = 1994, B = 1998, and C = 1999). Pools 1 and 2 persisted much longer 

than pools 6 and 7. Pool 7 was not sampled in 1999. 

Long-lived Pools Short-lived Pools

Pool number: 1 2 6 7

 Year A B C A B C A B C A B C

Anostraca 

Branchinecta sandiegoensis x x x x x x x x

Branchinecta juveniles x x x x x x x x x x x

Total # of species 1 1 (1) 1 1 (1) 1 1 (1) 1 1 -

Cladocera 

Macrothrix hirsuticornis x x x x x x x x x

Ceriodaphnia dubia x x x x x x x

Moina micrura x x x

Alona cf diaphana x x x x x

Simocephalus sp. x x x

Total # of species 5 4 1 5 3 2 3 2 1 0 1 -

Ostracoda

Cypris pubera x x x x x x x x x x

Cypris sp. 1 x x x x x x x x

Cypris sp. 2 x x x x x

Cypris pustulosa x x x x x

Limnocythere glypta x x x x x

Potamocypris sp. x x x x

Prionocypris sp. x

Cypris virens x x x

Cypridopsis vidua x

Herpetocypris sp.

Ostracod N

Unknown x

Total # of species 7 7 3 7 8 3 3 2 0 1 2 -

Copepoda 

Acanthocyclops vernalis x x x x x x x x x x x

Acanthocyclops robustus x x x x x x x x x

Hesperodiaptomus 

 franciscanus

x x x x x

Total # of species 3 3 2 2 3 1 3 2 2 2 2 -

TOTAL # OF SPECIES 16 15 7 15 15 7 10 7 4 4 6 -

262 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Poster Session – Ephemeral Pool Crustacean Communities – Holtz and Others

0

2

4

6

8

10

12

14

16

18

20

1 2 3 4 5 6 7 8

Pool

S

pecies Rich

n

es

1994

1998

1999

Figure 1—Comparison of species richness in 1994, 1998, and 1999 in eight vernal

pools in San Diego County. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 263 

Habitat Ephemerality and Hatching 

Fractions of a Diapausing Anostracan 

(Crustacea: Branchiopoda)1

Marie A Simovich2, Thomas E. Philippi3, Ellen T. Bauder4, and 

Jacob A. Moorad5 

Key words: Branchinecta sandiegonensis, diapause, fairy shrimp, San Diego County, vernal pool 

Extended Abstract 

Diapause allows aquatic organisms to survive periods of drydown in intermittent 

pools. However, often not all of the individuals hatch in response to a filling event. 

This prolonged diapause can be a bet-hedging adaptation to unpredictability in the 

duration of filling events. Under the simplest bet-hedging model of selection on 

prolonged diapause, the fraction of eggs hatching in any filling event should 

approximate the fraction of filling events lasting long enough for successful

reproduction. We quantified filling durations in seven vernal pools in San Diego 

County inhabited by the endemic fairy shrimp Branchinecta sandiegonensis

(Branchiopoda: Anostraca) for all winter and spring events from 1983 to 1996. For

any given rainfall event, larger, deeper pools held water longer than smaller, 

shallower pools. However, the distributions of filling durations did not differ 

significantly among these pools, so selection should be similar across pools. 

Approximately 28 percent of the filling events lasted 17 days or longer, the 

developmental time to first reproduction (under field conditions) in B.

sandiegonensis. Laboratory hatching experiments determined a maximum hatching 

rate of 28 percent at 10o

C, with lower hatching percentages at higher and lower 

temperatures. The close fit between the probability of adequate filling duration and 

the hatching fraction, and the similar distributions of filling durations among these 

pools, suggests that creating pools with hydrology suitable for B. sandiegonensis may

require care to match the filling durations to which this species appears adapted.

References 

An expanded version of this report was published as Philippi, T.E.; Simovich, M.A.; Bauder,

E.T.; Moorad, J.A.; Moorad, J.A. 2001. Habitat ephemerality and hatching fractions 

of a diapausing anostracan (Crustacea: Branchiopoda). Israel Journal of Zoology

47(4): 387-396. 

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2 Department of Biology, University of San Diego 3 Savannah River Ecology Lab, University of Georgia 4 Department of Botany, San Diego State University 5 Department of Marine and Environmental Science, University of San Diego and Department of

Biology, Indiana University

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 265 

Bee Diversity Associated with Limnanthes

Floral Patches in California Vernal Pool 

Habitats1

Joan M. Leong2 and Robbin W. Thorp3

Key words: Andrena limnanthis, Andrenidae, Jepson Prairie, solitary bees, species 

richness, wetlands 

Extended Abstract 

As with other kinds of wetland habitats in California, approximately 90 percent of

vernal pool habitat (estimated) has been lost in California. In southern California, 

losses are estimated to be even greater. The flora of these endangered habitats is 

reasonably well known, especially the spring flowering annuals that are found in or at 

the margins of vernal pools (for example, Blennosperma, Limnanthes, Lasthenia, and

Downingia species). However, the insect visitor fauna associated with these floral

species is much less well known, and what work has been done has been limited

primarily to the northern Central Valley region. The insect visitors associated with

vernal pool plants are of ecological importance because many vernal pool plant 

species are self-incompatible and require outcrossing to reproduce. Previous studies

indicate that species of solitary bees are especially dominant members of the insect

visitor fauna associated with vernal pool plants. One of our recent studies examined

the bee fauna associated with floral patches of Limnanthes douglasii ssp. rosea using 

a new sampling or monitoring method. We used colored pan traps or water traps to 

assess the species richness and abundance of the bee fauna associated with seven L. 

douglasii ssp. rosea populations or patches. These Limnanthes populations were 

located at Jepson Prairie Preserve near Sacramento. Prior to this study, basic

quantitative information on bee species richness and abundance in these endangered

wetland habitats had been lacking. 

We found remarkably high levels of bee species richness and abundance 

associated with the Limnanthes patches sampled at Jepson Prairie Preserve. A total of 

1,970 bees were trapped over four sampling dates. Of the 1,970 bees caught, 1,598 

were individuals of a native solitary bee, Andrena limnanthis. Andrena

(Hesperandrena) limnnathis is a specialist on the pollen of L. douglasii ssp. rosea

and a few closely related Limnanthes species. Our study revealed that 32 bee species

were foraging in these L. douglasii ssp. rosea patches over a 3-week period.

Approximately two-thirds of the bee species belong to the family Andrenidae, many 

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2 Department of Entomology, University of California, Davis, and Department of Biological Sciences, 

California State Polytechnic University, Pomona 3 Department of Entomology, University of California, Davis

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 267 
Poster Session—Bee Diversity in Vernal Pool Habitat—Leong and Thorp 

of which are thought to be endemic to vernal pool habitats. The mean number of bees 

caught per sampling unit, averaged over patches and dates, was 118.8 ± 20 SE 

(standard error), range 32-283. The mean number of bee species caught per sampling

unit, averaged over patches and dates, was 9.3 ± 0.8 SE, range 5-18. The number of 

A. limnanthis individuals and the number of bee species with intermediate levels of 

abundance was higher than expected. The shape of the bee species abundance curve 

was most similar to the log series or lognormal theoretical species abundance models. 

These results suggest that the diversity of vernal pool plant visitors may be 

underrated by most wetland ecologists and professionals. Many of the same vernal 

pool plants that host a variety of oligolectic or specialist bees and non-oligolectic 

bees in central and northern areas are also found in southern California vernal pool 

habitats. Thus, we propose that a similar level of bee diversity may occur in southern 

California vernal pool habitats. Because surprisingly little is known from these 

southern areas, we recommend that similar studies of the bee fauna of southern 

California vernal pool habitats be undertaken. Our results here will provide baseline 

data for future monitoring efforts. Such information is critical to biologically

informed decisions involving the conservation and restoration of biodiversity in the 

remaining vernal pool habitats in California. 

268 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Atmospheric Nitrogen Deposition and 

Habitat Alteration in Terrestrial and Aquatic 

Ecosystems in Southern California: 

Implications for Threatened and 

Endangered Species1

Mark Fenn, Mark Poth, and Thomas Meixner2

Key words: air pollution, amphibians, montane forests, N-enrichment, nitrate, runoff 

Extended Abstract 

Recent studies in the transverse ranges (including Class I Wilderness areas) of 

southern California have emphasized the strong linkage between levels of air 

pollution-related atmospheric nitrogen (N) inputs into montane watersheds and levels 

of nitrate in surface and subsurface drainage waters (fig. 1). Nitrate concentrations in

streamwater in southern California are the highest in North America for wildland 

watersheds. High nitrate concentrations in runoff are a result of excessive ecosystem 

enrichment with N from atmospheric deposition. When N enrichment results in the 

removal of biological N limitation and the N retention capacity of the system is 

exceeded, the ecosystem is described as “N saturated.” Nitrogen-saturated forests are

analogous to over-fertilized agricultural fields with high nitrate runoff. The primary 

symptom of an N-saturated ecosystem is high nitrate concentrations in the soil

solution, in soil leachate, and in groundwater and streamwater (Fenn and others

1998). Vegetation in N-saturated ecosystems contains high levels of N, and excess N 

also builds up in litter and soil organic matter, especially under fire suppression 

regimes. Stand densification is likely to be exacerbated by N deposition. 

Due to the interaction of N deposition with land management activities, it is

possible that past, present, and future land management practices (including fire

suppression, introduction of invasive species, and forestry practices) could minimize 

or exacerbate the adverse effects of N deposition on terrestrial and aquatic

ecosystems. Hydrologic flowpaths in a watershed also influence the impact of 

atmospheric N deposition on aquatic ecosystems. In summary, chronic N deposition 

results in excess N in terrestrial, riparian, and aquatic habitats. This dramatic change 

in the chemical environment of these habitats has high potential to upset the normal 

communities of vegetation, microbes, and micro- and macro-flora and fauna either 

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2 Plant Pathologist and Microbiologist, respectively, USDA Forest Service, Pacific Southwest Research 

Station, Forest Fire Laboratory, 4955 Canyon Crest Drive, Riverside, CA 92507, mfenn@fs.fed.us; and 

Professor of Hydrology, Department of Environmental Science, University of California, Riverside, CA

92521

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 269 
Poster Session—Nitrogen Deposition and Habitat Alteration—Fenn and others

via direct effects on sensitive organisms or via cascading effects on the food chain. 

As an example, evidence suggests that elevated nitrate levels in pond water may be 

toxic to larvae of several western frog species that are in decline (Marco and others

1999). Because of the excessively high N enrichment of many southern California

ecosystems, the effect on key species is a neglected area of research that should be 

addressed. 

Average nitrate concentrations in streams in Devil Canyon in the western San 

Bernardino Mountains fall within the low end of the range of concentrations that 

cause developmental, physical, and behavioral abnormalities in sensitive amphibian 

species (Rouse and others 1999). However, peak concentrations in streams in the 

western San Bernardino Mountains are more than twice as high as average values and

may have acute effects on sensitive amphibian species. Tadpole and larval stages 

appear to be the most sensitive life stages. Amphibian prey and predators also can be 

affected by elevated nitrate, a fact that may alter predation pressure and ecological 

balance. However, studies on the effects of elevated streamwater nitrate on aquatic

biota have not yet been done in the San Bernardino Mountains. 

Recent studies suggest that chronic N deposition may be a factor in the shift

from native shrub species in coastal sage scrub ecosystems to exotic annual grasses –

a widespread phenomenon in the remaining coastal sage habitats in southern 

California (Padgett and others 1999). Changes in vegetative communities may also 

be occurring in the forest understory in N-saturated forests, but this has not been

investigated sufficiently. In high N deposition sites in the San Bernardino Mountains, DC5 DC7 DC1 DC2 DC4 DC6 DC3 SGW4 SGW3 SGW2 SGW1 SGW5 SGW7 SGW11 SGW8 SGW10 SGW9 SGW12 SGW6 Average NO3

- Concentration (μmol L-1)

0

20

40

60

80

100

120

140

160 SGW Southern

Sites

SGW Northern

Sites

Devil Canyon

Sites

Stream Sites

N deposition

35 kg/ha/yr

N deposition

10-15 kg/ha/yr

N deposition

2-6 kg/ha/yr

Figure 1—Average streamwater nitrate concentrations in Devil Canyon (high N 

deposition area in the western San Bernardino Mountains), along the southern side

of the San Gorgonio Wilderness (SGW) (moderate N deposition), and on the northern

side of the SGW (low N deposition; Fenn and Poth 1999). 

270 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Poster Session—Nitrogen Deposition and Habitat Alteration—Fenn and others

the ecosystem is highly N enriched as evidenced by high N concentrations in foliage

(of understory and overstory species), litter, mineral soil, soil solution, soil leachate,

and trace gas emissions from soil (Fenn and others 1996). Forested sites further away

from N pollution sources (motor vehicle emissions and dairy farms) receive

atmospheric N deposition loads which are at least 5-10 fold lower, and plant growth

is highly N limited – which is the normal condition for these forests. Thus, the longterm ecological and environmental impacts of decades of elevated atmospheric N

inputs into wildland ecosystems in southern California remains an important and

sometimes under-appreciated issue for natural resource managers, especially in the

face of ever-increasing human populations. 

References 

Fenn, M.E.; Poth, M.A. 1999. Temporal and spatial trends in streamwater nitrate

concentrations in the San Bernardino Mountains, southern California. Journal of

Environmental Quality 28: 822-836. 

Fenn, M.E.; Poth, M.A.; Johnson, D.W. 1996. Evidence for nitrogen saturation in the San

Bernardino Mountains in southern California. Forest Ecology and Management 82:

211-230. 

Fenn, M.E.; Poth, M.A.; Aber, J.D.; Baron, J.S.; Bormann, B.T.; Johnson, D.W.; Lemly,

A.D.; McNulty, S.G.; Ryan, D.F.; Stottlemyer, R. 1998. Nitrogen excess in North 

American ecosystems: Predisposing factors, ecosystem responses, and management 

strategies. Ecological Applications 8: 706-733. 

Marco, A.; Quilchano, C.; Blaustein, A.R. 1999. Sensitivity to nitrate and nitrite in pondbreeding amphibians from the Pacific Northwest, USA. Environmental Toxicology

and Chemistry 18: 2836-2839. 

Padgett, P.E.; Allen, E.B.; Bytnerowicz, A.; Minnich, R.A. 1999. Changes in soil inorganic 

nitrogen as related to atmospheric nitrogenous pollutants in southern California.

Atmospheric Environment 33: 769-781. 

Rouse, J.D.; Bishop, C.A.; Struger, J. 1999. Nitrogen pollution: An assessment of its threat

to amphibian survival. Environmental Health Perspectives 107: 799-803. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 271 

Surveys for California Red-Legged Frog 

and Arroyo Toad on the Los Padres 

National Forest1

Valerie K. Hubbartt and Thomas G. Murphey2 

Key words: amphibians, monitoring, range expansion, stream channels, suitable habitat

Extended Abstract 

Starting in the spring of 1999 through the fall of 2000, USDA Forest Service 

biologists have conducted surveys throughout the Los Padres National Forest for the 

federally-listed California red-legged frog (Rana aurora draytonii) and arroyo toad

(Bufo californicus). Sites known to have frogs or toads present were monitored for

breeding and possible impacts of Forest Service activities. Potential impacts to frogs

and toads include roads, trails, campgrounds, off-highway vehicles, grazing

allotments, and introduced species. 

Historical locations and areas having suitable habitat for both species were

visited to confirm occupancy. Day surveys were conducted to confirm presence of 

eggs and tadpoles, in the breeding season, and to determine areas that had potentially

suitable habitat for adults. Night surveys were performed to establish presence or 

absence of adult frogs and toads. When adults were present, the locations of 

individuals were recorded and mapped, and pertinent habitat information was 

collected. Habitat information included land use, dominant vegetation, air and water 

temperatures, pool depth and location, and general hydrological information about 

the site. 

Approximately 267 kilometers of stream channel have been surveyed. Survey 

results are summarized as the maximum number of adults seen within sub-watersheds

(table 1). These numbers should not be interpreted as exact counts of current 

populations. Surveys through August 2000 have resulted in 11 new locations on the

Los Padres National Forest for red-legged frogs and an expanded range for arroyo

toad on upper Piru Creek near Castaic Mine and Goldhill Campground in Ventura 

County, based on past surveys (Stephenson and Calcarone 1999). Arroyo toad 

breeding was confirmed in these drainages: Piru Creek, Aqua Blanca Creek, Sespe

Creek, Indian Creek, Mono Creek, and the Santa Ynez River. Arroyo toad breeding 

occurred from April to the end of May, and toads were still calling as late as the end

of June, 2000. Breeding for California red-legged frog occurred from February to the 

end of March and at all sites that contained more than one frog. 

 

1 A poster version of this paper was presented at Planning for Biodiversity: Bringing Research and 

Management Together, a Symposium for the California South Coast Ecoregion, February 29-March 2, 

2000, California State Polytechnic University, Pomona, CA. 2 USDA Forest Service, Los Padres National Forest, Santa Maria, CA 93453. 

USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 273 
Poster Session—Amphibian Surveys—Hubbartt and Murphey

Table 1—Results of May 1999 to August 2000 amphibian surveys on the Los Padres National

Forest. 

Maximum # of Total Area 

Sub-watershed Adults Observed Surveyed (km.) Creeks/Rivers Surveyed 

California 

Red-legged Frog

Arroyo 

Toad 

Arroyo Seco 0 0 8.7 Arroyo Seco, Santa Lucia and Tassajara 

Carmel 50 0 18.0 Carmel River and Miller fork of Carmel*. 

San Antonio 0 0 3.6 North Fork San Antonio 

Monterey Coastal 

Streams 

 14 9 17.0 Little Sur, Sycamore Cyn.*, Laffler Cyn.*, 

Hare Cyn., Limekiln, Kirk, Mill, Wild

Cattle, Prewitt, Plasket, Willow, Salmon, 

Alder, Villa, Redwood Gulch, Dutra and 

San Carpoforo 

Salinas 6 0 21.7 Salinas* and Navajo

San Luis Obispo 

Coastal Streams 

 18 0 5.9 Morro, East. fork Morro* and Chorro 

Hausna 0 0 1.6 Trout 

Arroyo Grande 5 0 5.7 Big Falls*, Little Falls* and Lopez Cyn. 

Cuyama 23 0 21.4 Branch*, Alamo*, Cuyama, Stephens Cyn., 

Pine Cyn.*, Aliso Cyn., Bates Cyn. and 

Rancho Nuevo 

La Brea 78 0 20.0 North fork La Brea*, South fork La Brea*

and Main fork La Brea*

Sisquoc 34 1 12.6 Sisquoc*#

 and Horse Cyn.

Manzana 35 0 7.2 Manzana* and Davey Brown 

Santa Ynez 242 70 50.0 Santa Ynez*#

, Oso, Cachuma*, Mono*#, 

Indian*#

, Blue Cyn.*, Agua Caliente*#, 

Fox* and Alder*

Matilija 0 0 0.8 N. Fork Matilija 

Sespe 0 54 26.2 Sespe#

, Middle Lion, Rock and Rose Valley

Piru 0 44 47.0 Piru#

, Agua Blanca#, Canton Cyn., Reasoner

Cyn., Dominquez Cyn. and Lime Cyn.

* Drainages with California red-legged frogs present. 

# Drainages with arroyo toads present. 

Future surveys will be conducted to further identify areas of potentially suitable 

habitat and to locate new breeding areas. Inventorying and monitoring will continue 

where Forest Service activities may result in adverse impacts on the listed species. 

References 

Stephenson, John R.; Calcarcone, Gena M. 1999. Southern California mountains and

foothills assessment: Habitat and species conservation issues. General Technical 

Report PSW-GTR-172. Albany, CA: Pacific Southwest Research Station, Forest Service,

U.S. Department of Agriculture; 402 p. 

274 USDA Forest Service Gen. Tech. Rep. PSW-GTR-195. 2005. 
Pacific Southwest Research Station

PO Box 245

Berkeley, CA 94701

www.fs.fed.us/psw

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