Ecosystems of California 9780520962170

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EC OSYST E M S O F CA L I F O R N I A

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Ecosystems of California

Edited by HAROLD MOONEY and ERIKA ZAVALETA

Graphics edited by MELISSA C. CHAPIN

U NIVER SIT Y OF CALIFOR NIA PR ESS

University of California Press, one of the most distinguished

Library of Congress Cataloging-in-Publication Data

university presses in the United States, enriches lives around the world by advancing scholarship in the humanities, social sciences,

Ecosystems of California / edited by Harold Mooney and Erika

and natural sciences. Its activities are supported by the UC Press

Zavaleta ; graphics editor, Melissa Chapin.

Foundation and by philanthropic contributions from individuals

       pages cm

and institutions. For more information, visit www.ucpress.edu.

Includes bibliographical references and index. ISBN 978-0-520-27880-6 (cloth : alk. paper)

University of California Press

ISBN 978-0-520-96217-0 (ebook)

Oakland, California

1.  Ecology—California. 2.  Ecosystem management—California.  I. Mooney, Harold A., editor. II. Zavaleta, Erika, 1972–editor.

© 2016 by The Regents of the University of California

QH105.C2E36 2016 577.09794—dc2

2015016442

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The paper used in this publication meets the minimum requirements of ANSI/NISO Z39.48-1992 (R 2002) (Permanence of Paper). 8

To our students, whose enthusiasm inspired and motivated us to craft something more substantial than we could have on our own

To our families, whose love and support make our work possible

In memory of Rafe Sagarin, 1971–2015

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CONTENTS

CONTRIBUTORS xv PREFACE AND ACKNOWLEDGMENTS xix MARINE AND TERRESTRIAL MAPS OF CALIFORNIA xxi–xxii

1

Introduction / 1

3 Fire as an Ecosystem Process / 27

ERIKA ZAVALETA and HAROLD MOONEY

JON E. KEELEY and HUGH D. SAFFORD

The Need for This Book / 1 The Early Development of Ecology in California: A Different Path Taken / 2 Emergent Patterns / 2 Structure of the Book / 4 Recommended Reading / 5 References / 5

Introduction / 27 Fire Intensity or Severity / 29 Fire in a Historical Context / 32 Landscape Patterns of Burning / 33 Fire Departure Index / 34 Postfire Recovery of Plant Communities / 35 Fire Effects on Soils, Hydrology, and Carbon Storage / 38 Fire in Social Context / 39 Fire Management / 39 Global Changes / 40 Summary / 41 Acknowledgments / 41 Recommended Reading / 41 Glossary / 42 References / 42

PART ONE. DRIVERS

2

Climate / 9 SAM F. IACOBELLIS, DANIEL R. CAYAN, JOHN T. ABATZOGLOU, and HAROLD MOONEY

The Nature of the California Climate / 9 General Climate Features / 10 Warm Season / 12 Cool Season / 14 Atmospheric Rivers / 14 Ocean Influence / 14 El Niño and La Niña / 15 The Wide Range of Climates in California / 16 Spatial Variability of Temperature / 17 Seasonal and Diurnal Temperature Variation / 17 Spatial Variability of Precipitation / 18 Interannual Variability of Precipitation / 19 Spatial Variation of Incoming Surface Solar Radiation / 20 Regional Features / 20 Summary / 22 Acknowledgments / 23 Recommended Reading / 23 Glossary / 24 References / 24

4

Geomorphology and Soils / 47 ROBERT C. GRAHAM and A. TOBY O’GEEN

Overview of Geologic Processes / 47 Sierra Nevada / 48 The Central Valley / 51 Coast Ranges / 55 Klamath Mountains / 56 Cascade Range / 58 Modoc Plateau / 60 Basin and Range Province / 60 Transverse and Peninsular Ranges / 64 Coastal California / 66 Summary / 68 Recommended Reading / 68 Glossary / 68 References / 70

5

Reconstructing the Past: Methods of Historical Ecology / 133 Misty Origins: Mesozoic Ecosystems (252– 66 Ma) / 134 Recognizable Taxa with Nonanalog Associations: Paleogene and Neogene (66–2.6 Ma) / 136 The Ice Age Rollercoaster: Quaternary Environments (3 km from the coastline), kelp forests and estuaries, and selected current and former coastal dune sites (labeled). Major interstate (white) and state (light orange/red) area roads are shown. At the scale of the individual site, within Carmel Bay, distributions are shown of kelp forest, rocky intertidal, and sandy beach ecosystems, in relation to daily mean lower low tides and daily mean higher high tides, with blue ocean shading indicating depth and bathymetry. Data from the U.S. Geological Survey, National Hydrography Dataset (NHD); California Department of Fish and Wildlife, Ocean Imaging; the Seafloor Mapping Lab of California State University– Monterey Bay; and ESRI. Map: Parker Welch (Center for Integrated Spatial Research [CISR], UC Santa Cruz), Emily Saarman.

TERRESTRIAL ECOSYSTEMS

Terrestrial ecosystems of California. The state’s three major desert regions are labeled: Great Basin, Mojave, and Colorado. Data from the U.S. Geological Survey, Gap Analysis Program (GAP); Cal Fire, Fire Resource and Assessment Program (FRAP); and the U.S. Geological Survey, National Hydrography Dataset (NHD). Map: Parker Welch (CISR, UC Santa Cruz).

ONE

Introduction ER I K A Z AVALE TA and HAROLD MOONE Y

It is no small task to understand the characteristics of the various ecosystems that occupy California, since there are more different types here than anywhere else in the United States. This is the joy and marvel of living in a state where in a short distance, you can be in a completely different landscape from the one you live in. It could be the coast, a beautiful lake, a magnificent river, dramatic deserts, or the rich and diverse forests that occupy the state—​or, for that matter, all of those landscapes depending on the direction you pick. As pioneering California biologist and university president David Starr Jordan (1898) wrote, “There is from end to end of California scarcely a common mile.” The task that we lay out in this book is to bring together the necessary knowledge to understand the nature of California’s diverse landscapes and their biotic and social character so we can not only appreciate their uniqueness but also position ourselves to ensure that they are being utilized sustainably, at present and into the future. We take the position that these diverse systems are not totally isolated from one another but rather are linked and interacting to varying degrees. Thus we take on the challenge of assembling information on all of California’s systems, managed and natural, terrestrial and aquatic. We look not only at the status and dynamics of these systems at present but also at how these systems came to be

and how they are likely to look in the future. We understand that although California is somewhat of an ecological island, isolated by mountains and climate from much of the continent that it bounds, it is not unconnected from the rest of the nation and world. The lessons we gather from a comprehensive look at this singularly diverse and complex state can, we hope, inspire integrative and dynamic thinking about ecological systems elsewhere.

The Need for This Book More great books about the natural world and our connections to it have been written about California than probably anywhere else. So far, however, these books have not included a comprehensive, process-oriented ecology of the entire state. California contains more distinct ecosystem types by far than any other U.S. state. California has also had a long, complex, and unique history as a dynamic social-ecological system. It is both the most biodiverse and the most demographically diverse state in the nation. It hurtles towards a complex, unique, and uncertain future. In this book we strive to focus on process in every sense—​on the cross-scale processes that define the boundaries and character of each of California’s 1

ecosystems; the interacting social and ecological dynamics that have shaped the ecosystems we see today and continue to shape them going forward; the articulations between systems defined ecologically and systems defined by human enterprise, from fisheries to agriculture to cities; and the prospects for stewardship now and in the coming century. The result is a book with twice as many chapters as we first planned, and even so with gaps. Why did we feel we needed this book? First, although many courses are taught about California’s ecology in the state’s institutions of higher learning, there has been no comprehensive text available to cover the diverse ecosystems of the state from terrestrial to aquatic to marine; their distribution, structure, and composition and functional attributes; and least of all the range of natural and managed systems. We both have been teaching university courses about California’s ecology for a long time without an accompanying text. One of us (Hal) taught Ecosystems of California at Stanford for some forty years using materials culled from a great variety of sources. Especially in a state renowned globally for its university system, it ought to be easier for people to teach and learn about California’s tremendous diversity of ecological systems. Beyond the classroom, a comprehensive overview of the state of knowledge about California’s many ecosystems could serve as a valuable reference for researchers, decision makers, and stewards. Among other things, each of the chapters in this book addresses knowledge gaps that highlight opportunities and needs for new scholarship; history that provides context for what we see today; and past, current, and future threats and losses that highlight opportunities and needs for policy and management action. The volume is intended as a source book for teaching and for conservation, policy, planning, and decision-making in California. Second, we wanted to capture the complexity and dynamism of a more comprehensive set of systems in California than other works have. The field of ecology has evolved over the past several decades to embrace perpetual dynamism, historical contingency, cross-scale interactions, and emergent features. This approach has replaced a more static view and a focus on structure, inviting more attention to function and processes. Much of this contemporary perspective on ecology has been explored but not synthesized in California, and it deserved to be brought together in one place. Finally, an up-to-date treatment of the history, ecology, management, and future of each of California’s ecosystems could undergird targeted research, informed management and stewardship, and inspired decision-making and citizenship. It could make plain what we know—​a nd what we need to know. It could also provide a view of the state’s ecology as integrated with and dependent on, rather than artificially separated from, the state’s economy, culture, and human communities. Because we care about the diversity and ecology of California, we want to explore and articulate how contingent its past has been and its future will be on our collective choices and decisions.

Photo on previous page: California brown pelicans (Pelecanus occidentalis californicus) and early winter swell, Scotts Creek, ­California. Listed as endangered in 1970 due to severe DDT exposure throughout the U.S., brown pelicans recovered gradually after DDT was banned in the early 1970s and were officially delisted in 1999. Their recovery reminds us that California’s ecosystems today reflect a history of conservation successes as well as human impacts. Photo: Ed Dickie (eddickie.com). 2  In troduction

The Early Development of Ecology in California: A Different Path Taken What we know best and least about California’s ecosystems is colored by the unique development of ecology as a science in this state. California is an island ecologically, and it was also an island in the way our knowledge of natural systems developed through time. This yielded both innovations originating in the state and gaps peculiar to California. Michael Smith (1987) noted, for example, that early Californian scientists were influenced by their “observation of California’s physical environment [that] profoundly affected their thinking—​both about science and about their new home. The peculiarities of their social environment influenced the ways in which they sought support for science and eventually prompted their efforts to arbitrate between the land and its occupants. These forces, compounded by geographic separation from other scientists, contributed to a professional role for earth and life scientists that differed in significant ways from that of their Eastern counterparts” as did the “rich intermingling of environmental influences [that] checkerboards California with dissimilar ecosystems . . . [that] compress into dramatic adjacency in California” in contrast to the gradual nature of ecosystem change (through space) in other parts of the country. California did not achieve statehood until 1850, three-quarters of a century after the establishment of the United States. In 1853 a small group of citizens in San Francisco founded the California Academy of Sciences to provide a home for the state’s growing natural history collections. The San Francisco Bay Area became the initial hot spot for the early buildup of local scientific capacity at the academy as well as at the emerging universities. In 1868 the University of California was founded, and early faculty including Joseph LeConte, E. L. Greene, Willis Lynn Jepson, and Joseph Grinnell combed the state to catalog its geological, plant, and animal diversity. They rapidly accumulated deep knowledge about both the state’s biogeography and the ecological and behavioral traits of the specimens that they collected. The scientific basis for the optimal use of California’s natural resources by society was fostered from the outset because the new University of California was designated from its beginnings as a land grant university. A College of Agriculture was established in 1875, and in 1913 a forestry division was added. In much of the U.S. and beyond, scientists involved directly in studying natural systems often avoided those with a human presence. In California, environmental science and protection of natural systems coevolved more closely, reducing the lines dividing ecology from conservation. In the early twentieth century, leading ecologists in the state became alarmed at the extensive overexploitation of its natural wealth and took direct actions to ensure its protection. They became directly involved in conservation efforts and helped ignite a national conservation movement through their roles in founding the Sierra Club and the Sempervirens Club. Then as now, California had the opportunity for leadership in bridging scientific knowledge to sound conservation and resource management policies and actions.

Emergent Patterns The chapters in this book were written by a diverse cadre of experts on California’s ecosystems. Each team of authors, of

FIGURE 1.1 Wildfire sunrise in Lassen Volcanic National Park. Photo: Ed Dickie (eddickie.com).

course, brought its own set of perspectives to the chapter in question and chose to a large degree how best to structure the chapter to reflect its subject. This inevitably produced differences in the emphases of each chapter on particular ecological scales, taxonomic groups, trophic levels, and so on. Nevertheless, we strived for strong coverage in every chapter across ecological scales (from organismal adaptations to species interactions to feedbacks and interactions between community and ecosystem scales), time periods (from historical to future), and applications (from basic ecology to conservation, management, and policy implications). As a result, variation among chapters reflects genuine differences in knowledge, state, and stressors of each system, among other things. For example, it became clear in the process of developing the chapter on alpine ecosystems (29) that very little research has been done on the ecosystem ecology of California’s alpine; this is reflected in the chapter’s coverage of that topic. By the same token, consistency among chapters reveals some genuine, emergent patterns that we did not necessarily expect at the outset. A strong pattern emerging across many ecosystems from chapters here deals with the historic moment we inhabit in California’s conservation trajectory. Chapter after chapter describes the recovery of species and systems from a nadir of environmental quality that occurred roughly forty to fifty years ago. To a surprising extent, some trends have been successfully reversed, such as declining air quality in California’s cities (see, e.g., Chapters 7 and 39) and declining species from island endemic plants to peregrine falcons to sea otters in the face of stressors from DDT to overharvesting and invasive species (see, e.g., Chapters 11 and 34). Older legacies, like the widespread damage caused by over-

grazing and gold mining in the nineteenth century, have both continued to fade with time and been actively addressed by restoration and remediation efforts (see, e.g., Chapters 5, 19, and 33) and changes for the better in management of such forces as fire and grazing (see, e.g., Chapters 3, 23, and 37) (Figure 1.1). These successes are tempered by at least two other trends: the ongoing loss of habitat to continued urban and exurban development (see, e.g., Chapters 5, 22, 25, and 31) and the emergence of climate change as a growing, ubiquitous force influencing California’s ecosystems (see, e.g., Chapters 14, 17, 26, 27, and 29, among many others). California’s ecosystems can be profound integrators of the complexity around interacting environmental changes at various temporal and spatial scales; for example, the acid neutralizing capacity of Emerald Lake in the Sierras declined steadily from about 1920 to 1970 before beginning to climb from the late twentieth century to the present, possibly reflecting first the effects of rising fossil fuel burning and acid deposition, then the effects of advancing snowmelt under increasing temperatures (Chapter 32). Examples from throughout the volume make clear the prevalence of interactions of this nature across ecosystems. Across the state, matter and energy move downstream, such as the movement of sediment, nutrients, and other compounds from forest to river to estuary to intertidal and kelp forest ecosystems (Chapters 17–​19, 28, and 33); and upstream, such as in the fires that burn from the montane forest up into the subalpine forest (Chapters 3 and 28), the nutrients moved by anadromous fishes from the ocean to freshwaters and nearby terrestrial settings (Chapters 33 and 35), and the air masses that carry moisture, pollutants, and even propagules from the coast to the mountains (e.g., ChapIn troduction  3

ters 2, 7, and 27). Source and sink relationships abound; entire heterotrophic ecosystems in California, subsidized by flows of energy and matter from other ecosystems, range from its estuaries (Chapter 19) to its cities (Chapter 39). Offshore, upwelling associated with the California current generates tremendous marine diversity and productivity that move among the porous boundaries of intertidal, rocky reef, estuarine, open ocean, and sandy beach ecosystems (Chapters 16–​ 20) in forms ranging from carcasses to suspended particulate organic matter. A third, emergent pattern from across systems concerns north-south distinctions in dynamics, threats, and in some cases perspectives on ecology that arise from differences in what studies have emphasized as well as in the underlying systems themselves. These contrasts and gradients across the latitudinal range of the state are most clear for ecosystem types that span a large part of the state. For example, the dynamics and dominance of invasive plant species vary markedly from south to north in coastal sage scrub (Chapter 22) and grasslands (Chapter 23). The historical and present roles of fire vary from south to north in chaparral ecosystems (Chapter 24); treeline elevation shifts (Chapters 28 and 29); and dominant species turnover in systems ranging from desert scrub (Chapter 30) to montane forests (Chapter 27) and kelp forests (Chapter 17). Finally, although statewide development patterns have altered coastal, low-elevation ecosystems more and interior, high-elevation ecosystems less (e.g., highly altered and fragmented coastal sage scrub, intermediate desert [low, interior], and relatively unfragmented subalpine and alpine), the condition of particular ecosystems tends to vary more, and in complex ways, along north-south lines than from west to east. For example, coast redwood forests (Chapter 26) experienced historically high rates of clearing throughout the coast but are now more affected by fire, climate change, and disease to the south than in the state’s north. Grasslands (Chapter 23) are more fragmented, converted, and invaded in the south than in the north within the coastal and interior valley grassland types.

Structure of the Book We designed this book to be used in a variety of ways to reflect a diversity of needs. We provide this brief overview as a road map to the book and a guide for selecting sections and chapters for reference, teaching, and study. Throughout the volume, chapters include recommended further readings and chapter-specific glossaries of technical terms not described in the text. We also chose to keep the references with each chapter to facilitate use of individual chapters for reference and teaching. The first part of the book following this introduction examines overarching drivers of patterns and processes on the California landscape (“Drivers,” Chapters 2–​7). Although we initially conceived of them as more or less abiotic drivers, in reality they occupy a range from strongly abiotic (e.g., oceanography, climate) to largely influenced by biotic forces (e.g., fire, soils). All incorporate, again to varying degrees, the effects of human activities. In particular, the chapters on atmospheric chemistry (7) and population and land use (5) explicitly focus on human drivers, and the chapter on fire as an ecosystem process (3) deals extensively with human effects on fire regimes. The chapter on population and land use is a hybrid between a historical and a driver chapter; it is 4  In troduction

the abridged story of how human population and land use dynamics and in the postcolonial era have shaped the lay of the land today. We felt that if we placed it in the history part of the book (“History,” Chapters 8–​10), we would erroneously convey that land use and human habitation are less-than-critical drivers of ecosystem patterns and processes in California. The next three chapters focus on history: the paleohistory of vegetation and animals, respectively, and an ecological history of indigenous Californians and their roles in shaping the California we see today. The chapter on paleovegetation (8) necessarily tackles the stage for vegetation history in the region and covers the geomorphological, climate, and ecosystem history of California. The chapter on vertebrate prehistory (9) focuses on mammals as the best-understood group in the paleorecord and traces their history in the region from sixty million years ago. The chapter on indigenous California (10) illustrates the degree to which scholarly debate and changing political contexts can influence, over time, conceptions of how people related to and shaped their environment in the past, and how that history has influenced today’s ecosystems. The next five chapters (“Biota,” Chapters 11–​15) describe overarching biotic patterns, threats, and concepts as a foundation for the rest of the book. The chapter on biodiversity (11) provides an overview of the state’s biological diversity and its spatial distribution, threats, and success stories across five focal taxonomic groups (plants, birds, mammals, invertebrates, and herptiles [amphibians and reptiles]). Subsequent chapters give an overview of the state’s major patterns of terrestrial vegetation (12) and of the special, widespread issues of biological invasions (13) and climate change impacts (14) in relation to the state’s ecology. The final chapter in this section (15) describes emerging understanding and framing of the relationships among biological diversity, ecosystem functioning, ecosystem services, and natural capital, which recur in nearly every subsequent chapter. The next part of the book describes the state’s ecosystems (“Ecosystems,” Chapters 16–​33). This section proceeds roughly from the offshore Pacific Ocean towards land, then inland and upward in elevation, and finally down the eastern mountain slopes to the desert. Chapters 31–​33 double back and tackle the major freshwater systems (wetlands, lakes, rivers) distributed across the state’s terrestrial ecosystems. Each chapter in this part describes the process-based ecology of an ecosystem: its distribution and the factors that shape it; key constituent species, species interactions and processes such as disturbance regimes; ecosystem dynamics, including trophic interactions and the cycling of elements; and ecological history, including the influences of people. Each chapter also describes the services to society associated with that ecosystem; the threats and challenges it faces; and likely future trends, challenges, and opportunities for its management. In this section we had to make some tough decisions about how to lump and split, and where to accept uneven coverage that reflected space constraints as well as knowledge gaps. For example, we decided not to split estuarine subtidal and salt marsh systems into separate chapters but to include both within a single chapter on estuaries (19), or to split deserts (30) into the various desert types that characterize eastern California but to combine them in one large chapter. Enough is known about each subsystem for a book, but this book had to fit between one set of covers. The chapters vary in length and balance among components, reflecting author choices as well as varying availability of knowledge about particular

systems. For example, the chapter on alpine ecosystems (29) focuses especially on geomorphological processes and biotic communities, with less emphasis on ecosystem dynamics, simply because less is known about them specifically in California’s alpine. The next part of the book describes managed ecosystems (“Managed Systems,” Chapters 34–​39), defined variously by the societal endeavors that created them and proceeding roughly from less to more strongly human-altered managed systems. California’s islands (Chapter 34) are a microcosm of many of California’s coastal systems but are distinct both ecologically, as islands, and in terms of their management for diverse purposes ranging from military to conservation to recreation. California’s marine fisheries (Chapter 35) are defined by the organisms they harvest, while forestry (Chapter 36) is an activity defined by the ecosystems that can support it, and range (Chapter 37) is what livestock do when they are let out to forage. The chapter on agriculture (38) emphasizes the history and economy of an activity that defines much of the state’s landscape and water use and supplies much of the country. The chapter on urban ecosystems (39) enters relatively new terrain, applying ecological concepts to the city as an ecosystem and examining how it supports biodiversity and ecosystem services in unique ways. The final two chapters in the book return to stewardship (“Policy and Stewardship,” Chapters 40 and 41). The chapter on regulation for resource conservation (40) traces efforts to regulate land use and stewardship over California’s history. The chapter on stewardship, conservation, and restoration in the context of environmental change (41) pulls together themes and cases from across the book to examine effective paths to sustain California’s ecological legacy into the future. We live in exciting times; California is in a state of accelerating change, but so is our knowledge about its ecological dynamics and their articulation with conservation and management efforts. In keeping with the spirit of pervasive dynamism that kindled this effort, we have already begun to think about the next edition of this volume; we would be grateful for your feedback. Finally, we hope this book will build understanding and guide stewardship of California’s incredible diversity, with appreciation for the layers of historical and dynamic forces reflected in the landscapes of this great state.

Recommended Reading A diversity of great books tackle California’s ecology from many angles. They include the literary—​f rom John Muir’s The Mountains of California (1875) to John McPhee’s “Los Angeles against the Mountains” (1989) and Assembling California (1993) to the anthology of Jack London’s California works, Golden State (Haslam 1999). They include the taxonomic—​ from Jim Hickman’s great Jepson Manual of the state’s vascular plants (1993), now in its second edition (Baldwin et al. 2012), to Inland Fishes of California (Moyle 2002). They include a great many natural histories, including Allan Schoenherr’s A Natural History of California (1995) as well as scores of guides to particular systems, regions, groups of organisms, weather, glaciers, ethnobotany, and geology. They include conservation surveys—​t he Atlas of the Biodiversity of California (Parisi 2003) summarizes tremendous information and knowledge in maps; while Life on the Edge: A Guide to California’s Endangered Natural Resources centers on wildlife, threats, and their management (Thelander and Crabtree 1994).

We cannot fail to mention the works of historical ecology, like Laura Cunningham’s A State of Change (2010) and the carefully researched California Grizzly (Storer and Tevis 1996); the comprehensive Manual of California Vegetation (Sawyer et al. 2009) and Terrestrial Vegetation of California (Barbour et al. 2007); and the many whole volumes exploring the ecology of individual ecosystems (like California Grasslands, Stromberg et al. 2007), processes (like Fire in California’s Ecosystems, Sugihara et al. 2006), and threats (like biological invasions in California’s Fading Wildflowers, Minnich 2008). Many books—​ from Ray Dasmann’s rousing Destruction of California (1965) to the cautious optimism of volumes like In Our Own Hands (Jensen et al. 1993)—​lay plain the conservation challenges that we face in California and impel us to act. Finally, a number of valuable California ecosystem resources exist online. The California Naturalist Program (http://calnat. ucanr.edu/) provides naturalist certification and training to members of the public interested in stewardship. It is one of many excellent efforts to build citizen science and service in the state. Considerable data and educational resources reside on the websites of the California Academy of Sciences, Berkeley’s Jepson Herbarium and Museum of Vertebrate Zoology, the California Invasive Plant Council, the California Native Plant Society, and many state agency sites. We have surely forgotten to mention volumes and efforts that we cherish, but their sheer numbers make it hard to call them all up at once.

References Baldwin, B. G., D. H. Goldman, D. J. Keil, R. Patterson, T. J. Rosatti, and D. H. Wilken, editors. 2012. The Jepson manual: Vascular plants of California. Second edition. University of California Press, Berkeley, California. Barbour, M., T. Keeler-Wolf, and A. Schoenherr. 2007. Terrestrial vegetation of California. University of California Press, Berkeley, California. Cunningham, L. 2010. A state of change: Forgotten landscapes of California. Heyday Books, Berkeley, California. Dasmann, R. F. 1965. The destruction of California. Macmillan Company, New York. Haslam, G., editor. 1999. Jack London’s Golden State: Selected California writings. Heyday Books, Berkeley, California. Jensen, D. B., M. S. Torn, and J. Harte. 1993. In our own hands: A strategy for conserving California’s biological diversity. University of California Press, Berkeley, California. Jordan, David S. 1898. California and the Californians. The Atlantic Monthly 82:793– ​8 01. McPhee, J. 1993. Assembling California. Farrar, Strauss & Giroux, New York. ———. 1989. Los Angeles against the mountains. Pages 183–​272 in J. McPhee. The control of nature. Farrar, Strauss & Giroux, New York. Minnich, R. A. 2008. California’s fading wildflowers: Lost legacy and biological invasions. University of California Press, Berkeley, California. Moyle, P. B. 2002. Inland fishes of California. University of California Press, Berkeley, California. Muir, J. 1875. The mountains of California. The Century Company, New York. Parisi, M. 2003. Atlas of the biodiversity of California. California Department of Fish and Game, Sacramento. Sawyer, J., T. Keeler-Wolf, and J. Evens. 2009. A manual of California vegetation. Second edition. California Native Plant Society, Sacramento. Schoenherr, A. 1995. A natural history of California. University of California Press, Berkeley, California. Smith, M. L. 1987. Pacific visions. California scientists and the In troduction  5

environment 1850–​1915. Yale University Press, New Haven, Connecticut. Storer, T. I., and L. P. Tevis. 1996. California grizzly. University of California Press, Berkeley, California. Stromberg, M., J. Corbin, and C. D’Antonio. 2007. California grasslands: Ecology and management. University of California Press, Berkeley, California. Sugihara, N. G., J. W. van Wagtendonk, K. E. Shaffer, J. Fites-

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Kaufman, and A. E. Thode. 2006. Fire in California’s ecosystems. University of California Press, Berkeley, California. Thelander, C. G., and M. Crabtree, editors. 1994. Life on the edge: A guide to California’s endangered natural resources: Wildlife. Biosystems Books, Santa Cruz, California.

PAR T ON E

DRIVERS

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T WO

Climate SAM F. IACOBELLIS, DAN IEL R . CAYAN, JOHN T. ABAT ZOG LOU, and HAROLD MOONE Y

The Nature of the California Climate California is one of the most geographically and ecologically diverse regions in the world, with landscapes ranging from sandy beaches to coastal redwood rainforests to snow-covered alpine mountains to dry desert valleys. Despite the variety of landscapes, much of California experiences what is termed a “Mediterranean-type” climate, sharing very similar climate and vegetation with other low- to mid-latitude regions on the west sides of continents, including the (European) Mediterranean Basin (e.g., di Castri and Mooney 1973, Minnich 2006), parts of South Africa, south and southwestern Australia, and central Chile. All of these regions have cool winters with intermittent wetness and hot, dry summers. This type of climatic regime has characterized California for millions of years and started to develop between 7 and 4 Ma ago as the California Current became colder and blocked entry of the seasonal North American monsoons emanating from Mexico (Adams and Comrie 1997). By 2.6 Ma ago, California’s vegetation elements resembled what we see today (see Chapter 8, “Ecosystems Past: Vegetation Prehistory”). In today’s California the basic progression of temperature and precipitation throughout the year is the same in nearly all parts of the state—​although the actual amount of precipitation

and the temperatures obtained are very dissimilar and explain to a large degree the variability of ecosystems that exist (Figure 2.1). The shading in the bottom row of Figure 2.1 indicates significant variability in monthly precipitation throughout the state—​monthly precipitation commonly ranges from less than 50% to over 150% of its long-term average, and the highest variations occur in the winter months when mean precipitation is greatest. In fact, the ups and downs of precipitation are a fundamental feature of California climate, not only within daily and monthly totals but also over annual and longer periods. California has the highest variability of year-to-year precipitation in the conterminous United States (Figure 2.2a). Within these fluctuations are multiyear periods of consecutive wet or dry years, notorious spells that impact ecosystems throughout the state. Reconstruction of the past four hundred years of northern California winter precipitation from tree ring analysis (Figure 2.2b) provides evidence that the high level of interannual variability observed in recent decades by weather instruments is not unusual, and that consecutive wet or dry years can last up to a decade or longer. The decoupling of moisture and energy of the Mediterranean climate type poses “problems” for organisms. In winter, 9

Temperature (˚C)

Sacramento

San Francisco

Bakersfield

San Diego

El Centro

30 20 10

Precipitation (mm)

150

100

50 0

J FMA M J J A SOND

J F M A M J J A SO N D

J FM AM J J A SOND J F MAM J J A SOND

J F MAM J J A SOND

Month FIGURE 2.1 Climatological monthly mean temperature (top row) and precipitation (bottom row) at five sites in California. Values computed using available station records of at least sixty years from the National Weather Service (NWS) Cooperative Observer Program (COOP). The shading represents the standard deviation around the monthly means.

when moisture is available, temperatures are too low for high rates of metabolic activity. In contrast, during summer, when energy is ample and temperatures are favorable for growth, little or no water often is available (Figure 2.3). The activities of virtually all organisms in California can be related to the juxtaposition of available water and suitable temperatures. Warm summers and autumns are occasionally punctuated by hot, extremely dry air masses, sometimes with strong winds from the interior, that stress vegetation and other organisms. Dry, windy conditions promote wildfires, which in turn affect the evolution and patterning of the vegetation on the landscape. For many ecosystems in California, biotic activity tends to be limited to the spring. This is the crossover point of adequate moisture and equable temperatures. During the dry summer months, vegetation becomes water-limited. The biota of vast areas of the state goes into dormancy to avoid the extended drought period. Those plants that can find a water supply, such as by a stream or lakeside, or those with roots that can penetrate deeply into the soil where water reserves are held, can remain active during the seasonal dry period. Although there are important variations in different regions and individual landscapes, discussed below in this chapter, this basic pattern of alternating dry and wet period is found in all areas of the state. It is often remarked that California experiences a drought every year during the summer period, owning to the absence of storms, clear skies, and high amounts of solar radia-

Photo on previous page: Visible satellite image of California and the adjacent Pacific Ocean, January 17, 2011. While sunny conditions prevailed over most of California on this winter day, dense tule fog extended throughout the Central Valley, and marine stratus hugged the coastline in the northern portion of the state. Embedded within the marine stratus are ship tracks created as aerosols from ship exhaust alter cloud radiative and reflective properties. The snowcapped Sierra Nevada Range and brown desert regions in the southeast part of the state further illustrate the diversity present in California’s climate at a moment in time. Geostationary Earth Orbiting Satellite (GOES) image courtesy of the National Oceanic and Atmospheric Administration (NOAA). 10  Drivers

tion reaching the surface, which produces high evaporative demand. Actual evapotranspiration (AET) is the combined flux of water to the atmosphere from soil surfaces and plants by evaporation and transpiration. AET is driven by solar energy and wind but depends on the availability of water. Potential evapotranspiration (PET) is the atmospheric demand of water from the soil and free water surfaces, which represents the amount of evapotranspiration that would occur if water were not a limiting factor. There are marked seasonal and spatial structures of PET and AET in California, where PET peaks in summer due to strongest energy availability and AET peaks in spring and early summer when water and the evaporative demand are jointly high (see Figure 2.3). Arid and semiarid areas have PET in considerable excess of AET and are thereby water-limited. Wetter areas of the state and those at high elevations, where energy is lower, exhibit AET values that are closer to PET and are energy-limited. During persistent dry spells, the landscape shifts toward a greater proportion of water-limited area, while during wet spells it shifts toward more energy-limited area (Hidalgo et al. 2005). Recent investigation of tree mortality indicates that increased water limitation is a primary mechanism of tree mortality in California’s lower-elevation forests (Das et al. 2013).

General Climate Features The North Pacific High pressure system, among all of the large-scale features of global atmospheric circulation, has the most immediate impact on California climate. A network of subtropical high pressure centers, including the North Pacific High, result from the uneven heating of the Earth’s surface. A net radiative influx (more solar radiation reaching the Earth’s surface than is vented from the surface by terrestrial infrared radiation) within latitudes equatorward of approximately 38° and a net loss of radiation poleward of approximately 38° results in low surface pressure near the equator and higher surface pressure near latitudes of 30°N and 30°S. Due to the position of the land masses, these high-pressure systems do

A Fraction 0.1 0.2 0.3 0.4 0.5 0.6 0.7

B

FIGURE 2.2 California precipitation

variability.

November-April normalized precipitation

A Coefficient of variation (standard

deviation/mean) of water year (October– ​September) precipitation at long-term monitoring stations across the conterminous United States, from water year 1951–​2 008. Source: From Dettinger et al. 2011.

2 1 0

B Reconstructed winter precipitation

-1 -2

1600

1800

1900

Year

Sacramento

San Francisco

Bakersfield

2000

San Diego

DEF PPT AET PET

250 200 mm

1700

(normalized by the standard deviation) for northern California Coast ranges based on analysis of blue oak (Quercus douglasii) tree rings. Annual estimates are plotted in blue and a five-year smoothed version is plotted in black. Source: From Stahle et al. 2013.

El Centro

DRAFT

150 100 50 0 J F M A M J J A S O ND

J F M A M J J A S O ND

J F M A M J J A S O ND J F M A M J J A S O ND

J F M A M J J A S O ND

Month FIGURE 2.3 Climatological monthly water balance at five sites in California estimated using a modified Thornthwaite soil water balance model. Individual components include precipitation (PPT), potential evapotranspiration (PET), actual evapotranspiration (AET), and the deficit (DEF, defined as the difference between PET and AET). When AET is greater than PPT, the difference is obtained from soil moisture. Evapotranspiration values obtained using the Penman-Monteith method for a reference grass surface and a prescribed soil water holding capacity of 150 mm. Source: Necessary climatic data was obtained from the PRISM Climate Group, Oregon State University, .

not form a continuous belt but rather form distinct high-pressure centers. The North Pacific High is a prominent example of these systems and is normally centered between the Hawaiian Islands and the west coast of North America. It is referred to as “semipermanent” because it is nearly always present in the region even though the strength and location of the highpressure system varies over time. Within the high-pressure system, the air from aloft descends, inhibiting vertically extensive cloud and storm formation. As the air descends, it also warms adiabatically, decreasing relative humidity, so the air within the highpressure system is generally warm and dry. In the Northern Hemisphere the air flow around the high pressure is clockwise (counterclockwise in the Southern Hemisphere) and is responsible for the general clockwise-flowing ocean currents in much of the North Pacific. The circulation around the North Pacific High also exposes much of California to prevailing northwesterly winds. However, local wind patterns can vary greatly due to the diverse topographic features in the state and the related thermal conditions.

Warm Season The North Pacific High reaches its maximum intensity and furthest northward extent during the summer months (Figure 2.4a). Associated with the expansion of the high-­pressure system, the North Pacific storm track becomes less active and recedes northward. Descending air within the high suppresses upward motion and inhibits rainfall. These two factors create dry conditions—​precipitation occurs very rarely in summer months throughout most of California, except for occasional thunderstorms in the mountains and desert areas. Because they lie on the eastward and northward flanks of the Pacific High, California coastal waters are driven by northwesterly winds that prevail throughout spring, summer, and early fall. The northwesterly winds “push” surface water away from the coast that is then replaced by the upwelling of colder water from below. Upwelling along with the southward flow of cool water by the California current (see Chapter 6, “Oceanography”) maintains relatively cold water, between 50°F and 60°F (10–​15°C), along most of the California coastline during summer. The relatively cool water along the California coast helps to promote a cool and moist marine layer in the lowest 300 to 1,000 meters of atmosphere. A key feature that distinguishes California from other regions of the United States is a low-level temperature inversion, which is a crucial feature in developing the marine layer and low-level clouds (Iacobellis and Cayan 2013) and also in trapping and accumulating particulates and pollution (Iacobellis et al. 2009; see Chapter 7, “Atmospheric Chemistry”). Normally, air temperature cools with height, but below the top of a temperature inversion air temperature increases with height. Controlled by the strength and position of the North Pacific High and other large-scale factors, temperature inversions are a persistent feature along the California coast and inland valleys. Along the coast the inversion is known as a “subsidence inversion,” caused by descending air in the North Pacific High combined with the cool air above the ocean surface. Interior valleys also frequently experience inversions, especially at night in the winter season due to low solar insolation, radiative cooling of land surfaces, and drainage of cool air into low-lying areas. During strong inversions the afternoon temperature at San Diego near sea level along the coast 12  Drivers

can be 8°C cooler than that at Alpine, in the adjacent mountains at 530 meters (approximately 1,730 feet) elevation. The subsidence inversion is present, to varying degree, throughout much of the year, especially in spring and summer. Because the density of air under the inversion is vertically stacked and very stable, the low-level temperature inversion forms a “cap” that inhibits vertical mixing of surface air. This can lead to elevated pollutant levels near the surface, particularly in coastal areas with mountainous topography to the east (e.g., Los Angeles and San Joaquin Valley air basins). The stability of the temperature inversion also leads to a lens of horizontally extensive marine stratus clouds, a nearly ubiquitous spring and summer feature of the California coastline that often extends from well offshore across coastal terrain that is less than approximately 600 meters (2,000 feet) elevation. Coastal fog occurs when the stratus cloud base is low enough to intersect the ground or the ocean surface. These clouds are most dense in evening through morning hours but still reflect considerable incoming solar radiation, which suppresses daytime temperatures along the coast. The marine layer and stratus regime is an important regulator that affects a wide variety of ecosystems through its moderation of solar radiation, temperature, and humidity. Periodically during summer months, a pulse of warm moist air flows northeastward into the southern part of the state from the Gulf of California, mainland Mexico, and sometimes the eastern subtropical Pacific. These pulses are part of the Southwest Monsoon, which commonly operates in northern Mexico, Arizona, and New Mexico, leading to frequent summer thunderstorms. When conditions align, the warm, moist air stream can be augmented by the remnants of hurricanes dissipating off the Pacific coast of Mexico. The warm and moisture-laden air often permits thunderstorm development and large precipitation events in the desert regions of southeasterly California. In contrast to most of California, these southeasterly deserts receive a substantial fraction of their annual precipitation during the summer months (see panel for El Centro in Figure 2.1). During spring and summer, winds in the coastal regions are usually from the northwest. The prevailing northwesterly direction may be broken along the Southern California Bight when winds take a southerly direction during periods when the Catalina Eddy takes hold. This eddy forms in the lee of the mountainous coastline, which takes an abrupt change in the direction at Point Conception. The Catalina Eddy most commonly occurs during periods of moderate-to-strong northwesterly winds off the Central California coastline (Kanamitsu et al. 2013). Some of California’s windiest locations occur in the vicinity of mountain passes, which direct air from cooler climates (usually coastal) to that of the warmer climates (often deserts or valleys). One such location separates the cool, coastal climate of the San Francisco Bay region from that of the Central Valley (Altamont Pass); another lies in the cool upper reaches of the Tehachapi Mountains located close to the hot Mohave Desert (Tehachapi Pass); and the third separates the southern California coastal plain from the Colorado Desert (San Gorgonio Pass). These winds tend to be strongest and most prevalent during summer, when the coast-to-interior temperature contrast is largest. As the interior region warms up and air near the surface rises, cooler marine air from the west rushes in through the mountain passes to replace it. These winds have a strong diurnal character and are the strongest in the a ­ fternoon when daytime heating and resulting thermal gradients are greatest.

A. Strong Pacific High

B. Weak Pacific High 3 2

8

100

1

10

L1009

H1031

2

10 24

10 16

H1018

B. Weak Pacific High 3

2

8

100

1

10

1000

10 18 26 Temperature (°C)

L1009

99

8 100

00 10

0

16

Height (km)

6 101

Height (km)

3

2 1 0

-4

4 12

Temperature (°C)

L983 2

10

10

16

08

10

1016

08

10

H1018 1016

2010 JAN 24 00Z

FIGURE 2.4 Satellite infrared imagery overlaid with surface pressure isobars (yellow lines) on days when the North Pacific High was strong (A) and weak (B). The inset in the upper right corner of each plot shows the vertical temperature at Oakland, California, for each day. A strong temperature inversion is evident on the day with the strong North Pacific High, but nonexistent on the day with the weak North Pacific High.

1016

1016

A. Strong Pacific High

24

Tem

L983

99

2010 JUL 31 00Z

2010 JUL 31 00Z

0

08

24

H1031

1

10

10

1004

2

8

24 10

1000

10 18 26 Temperature (°C)

100

00 10

0

16

Height (km)

6 101

Height (km)

L1004

3

201

Cool Season During winter months the North Pacific High weakens and shifts to the southeast. At the same time, the North Pacific storm track also migrates southward and intensifies due to a strengthened north-south atmospheric temperature gradient. As a consequence, the storms that form along the jet stream are more intense and more likely to directly impact California because blocking by the North Pacific High is reduced (see Figure 2.4b). The winter storms that impact California generally are spawned to the west and north over the North Pacific and propagate over the state in a northwest to southeast direction. Sometimes they can also move in a more westerly or occasionally southwesterly direction. The direction of the storm center, and most important the direction of the winds within the storm, is critical due to the strength and orientation of the winds with respect to topographic features (Dettinger et al. 2004). Interaction of storm winds and mountain ranges can create very heavy precipitation—​as air encounters the mountains, it is forced upward and cools. If the air cools to the condensation threshold, the water vapor in the air will condense, leading to enhanced precipitation. When the direction of the wind is perpendicular to the orientation of the mountain range, precipitation is maximized. As discussed later, California’s mountain ranges present a variety of directional orientations and thus exhibit differing responses to a given wind direction. Many areas of California, particularly high elevations and mountain passes, are prone to strong winds as winter storms move through the state. As a typical storm approaches from the northwest, winds near the surface are generally from the southwest and then swing around to westerly and finally northwesterly as the storm leaves the region. Certain locations are more susceptible to winds during different phases of storm passage due to topographic features such as the directional orientations of canyons and passes.

Atmospheric Rivers Occasionally, relatively narrow streams of air with high concentrations of moisture originating from the tropics or subtropics intercept the coast of California (Figure 2.5). Interaction of these moisture belts with storms forming along the jet stream can result in extreme amounts of precipitation (Dettinger et al. 2011). These moisture-laden air streams have sometimes been called a “pineapple express” but are now referred to as an atmospheric river. Because they often reside over a given location for several hours, retaining their active dynamics and generating heavy precipitation, atmospheric rivers can be extremely productive and are now understood to be the source of many of the floods along the West Coast. A study by Ralph et al. (2006) found that all of the significant floods from October 1997 to February 2006 in the Russian River area were associated with atmospheric rivers crossing the California coastline. In many cases, atmospheric rivers transport warm moist air into California so that storm precipitation occurs as rain up to unusually high elevations before transitioning to snow. These high-elevation snow levels can add to the heavy precipitation in producing flood flows due to the larger fraction of the landscape generating rainfall runoff along with possible melting of low- and mid-elevation snowpack due to rain-on-snow. There is evidence that unusually severe and lengthy atmo14  Drivers

spheric river episodes can appear in California, perhaps every couple of hundred years. A legendary historical event began on Christmas Eve in 1861, when high amounts of precipitation starting falling in the Sierra Nevada and lowland watersheds. Reports indicate that precipitation, in some form, persisted for forty-three days in central California and that virtually the whole state was affected by this wet spell. With the copious runoff, the Central Valley turned into a gigantic lake and the tidal inflow into San Francisco Bay was curtailed by freshwater discharge from the Sacramento and San Joaquin River runoff (Brewer 1930, Roden 1967). It is estimated that some eight hundred thousand cattle died and parts of Sacramento, the state capital, were submerged at times under 3 meters of flood water (Dettinger and Ingram 2013). The warm-moisture “conveyer belt” that operates during atmospheric river storms can also cause considerable flood damage at higher elevations. During a vigorous atmospheric river during December 26, 1996, through January 3, 1997, large amounts of warm rainfall fell in the northern and central Sierra Nevada. Along an elevational gradient, this storm produced 9.4 centimeters (3.7 inches) of rain in Sacramento, 24.4 centimeters (9.6 inches) in Auburn, and 75.4 centimeters (29.7 inches) in Blue Canyon and melted snow above the 3,050-meter level (about 10,000 feet). Copious runoff overwhelmed dams, and widespread flooding occurred. In addition to the climate-related hazards associated with atmospheric rivers, they are important contributors to annual precipitation across much of California. Dettinger et al. (2011) found approximately 30% to 45% of the annual precipitation fell in association with atmospheric river events, and Dettinger (2013) identifies atmospheric rivers as important “drought busters.”

Ocean Influence The climate of California is strongly influenced by its proximity to the Pacific Ocean. Water has a much larger heat capacity than air or soil, so an enormous reservoir of heat can be stored in the upper ocean. As a result, the ocean is much more difficult to cool or warm than are the nearby land masses. Thus the ocean and its marine air masses do not change temperature as much and the diurnal and annual variation of the ocean surface temperature and the air above it are much less than those of the air over land masses. In most cases, the ocean is cooler than the adjacent land in summer and warmer than land in winter. A similar relationship generally holds during the course of a day, with the ocean being cooler than land during the day and warmer than land during night. Consequently, the ocean creates an “air-conditioned” zone that pervades across the coastal lowlands to locations inland where marine air can easily penetrate. Ocean temperature offshore of California is relatively cool (compared to similar latitudes) due to frequent upwelling and horizontal advection of cold water in the California current system (see Chapter 6,“Oceanography”). Sea surface temperatures (SSTs) along California’s coast vary from about 10°C to 15°C (50°F to 59°F) north of Point Conception. South of Point Conception within the Southern California Bight, where the northwesterly winds are interrupted so that upwelling is diminished and the transport of the California current is reduced, SSTs are generally warmer, especially in summer, and have a larger seasonal variation of about 13°C to 22°C (55°F to 72°F). Because the upper ocean holds a massive amount of heat,

SSM/I IWV Image: 16 Feb 2004

17 Feb 04 Daily 3.187 streamflow rank

Record Top 0.2% Top 1% Top 2% Remainder of sites

40N

30N

144W

1

2

128W

136W

3

4

5

120W

6

7

IWV (cm) FIGURE 2.5 Composite satellite image of integrated water vapor (IWV) in centimeters (color scale at bottom of map) constructed from multiple polar-orbiting swaths during an atmospheric river event on February 16, 2004, between approximately 1400 and 1830 UTC (Coordinated Universal Time). The impact on river systems along the U.S. West Coast is shown using percentile rankings of daily streamflow on the following day, February 17, 2004, LST (Local Standard Time) (add eight hours for UTC) as indicated by the color-filled circles (e.g., red, yellow, and green circles indicate daily streamflow at the 98th, 99th, and 99.8th percentile levels, respectively; a daily streamflow greater than any previously recorded value is denoted by a blue circle). Gauges with less than thirty years of recorded data are not shown. Source: From Ralph et al. 2006.

the sea surface temperature is relatively resistant to change. Due to molecular and turbulent exchanges, the air in the boundary layer over the ocean tends to quickly equilibrate to the SST. Prevailing northwesterly winds transport this cool ocean air onto the California coastal regions and into accessible inland valleys. This marine layer influence moderates the range of temperatures that would otherwise exist in these regions, an effect that is particularly noticeable during summer months. In addition, since most of the storms that impact California track from over the ocean, the surface air within these systems is usually close to the ocean temperature. As a result, coastal and inland valleys of California are generally insulated from the cold winter conditions found at similar latitudes at other locations in the United States.

El Niño and La Niña Next to the annual cycle, the most important pattern of climate variation in California is the El Niño / Southern Oscil-

lation (ENSO). El Niño and its opposing phase partner, La Niña, are parts of a coupled ocean-atmosphere phenomenon rooted in the tropical Pacific and involving fluctuations in heat content, ocean surface temperature, and tradewinds. ENSO has a variety of global expressions but can be quite strongly felt in California’s oceanic and atmospheric climate. Through transfers of heat, moisture, and momentum, tropical ENSO conditions are transmitted (teleconnected) to the mid-latitude ocean and atmosphere including the California region. Near the equator, surface winds (the tradewinds) generally blow from east to west. Ocean surface currents, driven by the tradewinds, transport water in a westward direction, resulting in a sea level about a half of a meter higher in the western than in the eastern Pacific. Movement of the water away from the coast of South America drives the upwelling of colder deep water. During “normal” conditions, sea surface temperature (SST) decreases from west to east along the equatorial Pacific, with relatively cold water along the South American coast and very warm water (SST >30°C) in the Western Tropical Pacific sometimes Climate  15

ELEVATION

TEMPERATURE JANUARY MEAN

Elevation (km) 0

0.5

1 1.4

2

PRECIPITATION JULY MEAN

ANNUAL MEAN

Temperature (ºC) 3

4

-5 0

5 10 15 20 25 30 35

Precipitation (cm) 0

50

100 150 200 250

FIGURE 2.6 Maps showing (from left to right) elevation, January mean temperature, July mean temperature, and annual mean precipitation throughout California. Data obtained from the PRISM Climate Group, Oregon State University, http://prism.oregonstate.edu.

referred to as “the warm water pool.” Associated with this warm water are low surface pressures, strong atmospheric convection, and resulting deep cumulus clouds and heavy precipitation. Periodically, at irregular intervals of about two to seven years, the easterly tradewinds over the tropical Pacific slacken or occasionally even reverse direction (Rasmussen and Carpenter 1982, Philander 1990). In response, the SST increases in the Central and Eastern Tropical Pacific, upwelling diminishes along the equator and along the South American coast, convective activity shifts from the western toward the central and eastern equatorial Pacific, and dryness sets in over central America. This condition is called El Niño and can have a dramatic impact on weather and ecosystems both locally, in the Tropical Pacific, and with recognized global manifestations (Rasmussen and Wallace 1983, Philander 1990, Wolter and Timlin 2011) as well as strong regional impacts along the U.S. West Coast, particularly in California (Sette and Isaacs 1960, Monteverdi and Null 1997). During El Niño, the increase in SST in the Central and Eastern Tropical Pacific often consumes a vast region that alters heat and moisture fluxes to the extent that it impacts global atmospheric circulation patterns. One primary impact is a shift and intensification of the subtropical branch of the jet stream across the Pacific that bring winter storms further south during winter toward California. As a result, California can be (but is not always) subjected to more frequent winter storminess and precipitation during an El Niño event (Redmond and Koch 1991, Gershunov and Barnett 1998, Cayan et al. 1999). During the excep16  Drivers

tionally strong El Niño events of 1982–​1983 and 1997–​1998, California experienced enormous storm sequences and extraordinarily heavy precipitation, snowpack, and coastal storm impacts (Flick 1998, Bromirski et al. 2003, National Research Council 2012). In addition, these storms are generally warmer, contain more atmospheric moisture, and tend to track in a more westerly or southwesterly direction as they cross the California coast. The La Niña phase of the ENSO phenomenon occurs when tradewinds in the tropical Pacific become stronger than normal. In contrast to El Niño, La Niña SSTs are cooler than normal in the Central and Eastern Tropical Pacific, convection is strong in the far western equatorial Pacific, and upwelling along the equator and along South America is strong. In many La Niña events, associated atmospheric circulation changes result in a northerly shift in the jet stream across the North Pacific, limiting the frequency of winter storms and precipitation in California. In contrast to the often-increased California precipitation during El Niño years, La Niña years have quite consistently yielded subpar winter precipitation (Redmond and Koch 1991). This La Niña–​d iminished precipitation association, like its El Niño-increased precipitation counterpart, is strongest in southern California and fades in northern California.

The Wide Range of Climates in California Most of California is dominated by cool, intermittently wet winters and warm dry summers. Overlain on this general pat-

Spatial Variability of Temperature As expected, the broad-scale view of surface air temperature in California shows an increase from north to south. However, there is considerable complexity and spatial variation in temperature that cannot be explained simply by latitude (Figure 2.6).

Coastal Influence Throughout the state, coastal regions generally tend to be warmer in winter and cooler in summer than interior regions due to the moderating effect caused by the large thermal inertia of the ocean. Temperatures over the land that are closely connected to the coast are influenced by marine air masses transported from directly over the ocean (Lebassi et al. 2009), with air temperatures very close to the SST (Figure 2.7). The oceanic influence on surface air temperatures decreases as one moves inland and is sharply limited by coastal hills and mountains. In some cases, such as along the Big Sur coast of central California, this transition can occur over distances less than 3 kilometers due to the confluence of maritime air and steep topography, resulting in dramatic gradients in temperature and humidity. The oceanic influence on air temperatures has important consequences for annual range of temperature. Coastal locations have a low range of temperatures, while further inland the difference between summer maximum and winter minimum temperatures increases sharply. This temperature range is an important determinant of the distribution and adaptability of species and ecosystems across the region.

Deserts The desert regions of southeastern California are warmer than most other regions of the state throughout the year. This is partly due to their southerly location but is highly dictated by the year-round influence of the descending (and warming) air on the flank of the North Pacific High. Another factor that contributes to the region’s relative warmth is the adiabatic compressional heating of eastward-flowing air masses that traverse the Peninsular and Transverse ranges. In addition, during daytime hours, due to the lack of moisture and vegetation, incoming solar radiation quickly warms the surface since little solar energy is used to evaporate water, resulting in higher maximum temperatures. Conversely, at night, radiational cooling of the surface is very efficient because of the lack of clouds and extreme lack of water vapor in the air, so the surface can get very cold. As a result, deserts exhibit very strong diurnal temperature changes—​t his impacts the distribution and adaptability of ecosystems in these regions.

25

Annual mean precipitation (cm x 10-1) Annual mean daily temperatrue maximum (°C)

tern is extraordinary spatial and temporal variability. Considerable spatial variability arises from the inherent structure within the atmosphere and because of the influence of California’s varied setting and complex landscape. For example, Abatzoglou et al. (2009) used objective measures to identify eleven regional modes of temperature and precipitation variability within California. This variability is an inextricable part of California’s climate and ecosystems.

20

15

10

5 Temperature (ºC) Precipitation (cm x 10-1) 0 32

34 Los Angeles

36

38

40

San Francisco

42 Crescent City

Latitude FIGURE 2.7 Plot showing the latitudinal dependence of annual precipitation (triangles) and annual mean daily temperature maximum (circles) at coastal California locations based on COOP (Cooperative Observer Program) station measurements.

Elevation and Slope Throughout the year in most situations, relatively cool temperatures are found at the higher elevations, particularly along the Coast Ranges and the Sierra Nevada. As elevations rise from sea level, atmospheric pressure decreases. Since air temperature depends strongly on pressure, temperatures also generally decrease with height (an exception is the inversion layer discussed earlier in this chapter). On average in the free atmosphere air temperature decreases with height at a rate of 6.5°C per kilometer. The marked topographic features in California mean that associated air temperature changes due to elevation are large, as are the changes in species and ecosystems (e.g., Thorne et al. 2009). The patterning of the biota of California is also strongly influenced by the effect of topography on microclimate. Aspect (slope direction) also affects radiation, local temperature, and hence site water balance. These in turn determine biological processes and the distribution of organisms. North-facing slopes in the Northern Hemisphere receive considerably less radiation, particularly in the winter, than do south-facing slopes.

Seasonal and Diurnal Temperature Variation The range of temperatures is a critical factor for the viability of many species. Given its diverse landscapes, California has a broad range of temperature variation. Temperature variability, both diurnally and seasonally, along California’s coast is moderated by the insulating effects of the Pacific Ocean. Further inland, temperature variation is much greater (Figure 2.8). The largest annual temperature variations are found in the southeastern deserts of California. There, daytime temperatures during summer can exceed 120°F (approximately 50°C). At night, the surface cools quickly through Climate  17

precipitation by mountain and lowland topography, along with the larger-scale changes in precipitation that occur over the north-to-south and west-to-east extents of California’s boundaries.

ANNUAL TEMPERATURE VARIATION (Mean JULY Max) - (Mean JAN Min)

Topography

Temperature (ºC) 0

5

10

15

20

25

30

35

40

FIGURE 2.8 Annual temperature variation as measured by the difference between the mean July daily maximum and mean January daily minimum temperatures. Source: Data based on years 1981 through 2010 obtained from the PRISM Climate Group, Oregon State University, http://prism.oregonstate.edu.

the emission of infrared radiation. This radiational cooling is enhanced by the low concentration of atmospheric water vapor and cloud-free skies generally found in the desert regions. During winter, when the nights are longer, minimum temperatures can fall as low as 30°F (-1°C). The spatial distribution of diurnal temperature variations is similar to the annual values shown in Figure 2.6; however, the magnitude of the differences are considerably lower.

Spatial Variability of Precipitation California displays enormous variation in annual precipitation, ranging from 300 cm year-1 along its northern coastal regions (see, e.g., Minnich 2006). This remarkable spatial variability in precipitation results from the strong amplification or suppression of 18  Drivers

When moisture-laden air within storm systems sweeps over mountains such as the Sierra Nevada, the air is forced upward in a process known as orographic uplift. As the air rises, it cools due to decreasing atmospheric pressure. If the air cools sufficiently, it will reach its saturation point and water vapor will begin to condense, forming clouds and precipitation along the windward side of the mountain range. Orographic lifting amplifies precipitation along the windward side of mountain ranges. Conversely, an opposing “rain shadow” effect occurs on the leeward side, where the air mass begins to sink and warm as the atmospheric pressure increases with declining altitude. This effectively ends further condensation, and the clouds often quickly evaporate. Also, the air on the lee side is drier because moisture in the air mass has been depleted by precipitation on the windward side. Increased precipitation on the windward sides and rain shadows on the leeward sides are most pronounced when the air is flowing across (perpendicular to) the orientation of the mountain range. Most regions on the leeward side of California’s mountain ranges exhibit rain shadows—​notable examples are found on the east side of the Coast Range and the east side of the Sierra Nevada (Figure 2.9). A “rain-out” of moisture also occurs as storm-driven Pacific air masses cross the mountain blocks in California, as evidenced by the much drier, sparsely vegetated west slope of the White Mountains compared to the west slopes of the Coast Range and Sierra Nevada. Figure 2.9 shows seven locations along an east-west transect across central California, where the mountain ranges lie mostly parallel to the coast. As storm air masses move off the Pacific and across the state, larger amounts of precipitation are observed on the windward sides and smaller amounts on the leeward sides. The Sierra Nevada, with the highest elevations, exhibits the largest difference between windward and leeward precipitation and experiences greatest precipitation during storms with westerly to west-southwesterly wind flow (Pandey et al. 1999) and during warm storms with moist air masses (Cayan and Riddle 1993). While most of California’s mountain ranges are generally oriented north-south, others, such as the Transverse Mountains of southern California, lie in a more east-west configuration. Some winter storms—​for example, during some of California’s strong El Niño winters, contain relatively warm air and large amounts of moisture and produce a circulation with southwesterly or southerly winds. These storm systems tend to produce large amounts of precipitation in just a few hours along the windward side of the Transverse Range. Such high-intensity precipitation events can lead to disastrous results, including landslides and debris flows, particularly if a recent wildfire has denuded the mountain slopes (USGS 2005).

Latitude In addition to the variability created by the numerous mountain ranges, precipitation generally increases with latitude.

Meters 4000 2400 1600 900 500 200 0

Santa Cruz 76

Hollister 39

Coalinga 19

Fresno 28

Grant Grove 108

Independence 12

Death Valley 6

Mean annual precipitation (cm) FIGURE 2.9 Annual mean precipitation values at selected sites along an approximate latitudinal transect across California. The color map denotes elevation. Precipitation data based on selected COOP (Cooperative Observer Program) station measurements with at least forty years of data.

This is easiest to see along the immediate coastline, where the impact of mountain ranges on precipitation is minimized (see Figure 2.7). Most, but not all, winter storms are disturbances that are embedded in the fast, upper-level atmospheric jet stream. Typically these storms initially make landfall along the north coast of California and then propagate south and east. As these storms move in a general southeastward direction across California, they often weaken to the point that there is less frequent storm activity in southern California than areas north of Point Conception. In contrast to the more frequently occurring frontal storm system are “cut-off low” circulations that often form at lower latitudes and may deposit rain and snow in southern California but not to its north. Cut-off lows, as eddy-like circulations that are separate from the primary flow at upper levels, often propagate very slowly or remain stationary and can produce many hours of precipitation that is sometimes quite heavy. Overall, however, the southern portions of the state experience diminished storminess and less precipitation, which creates a south-to-north precipitation gradient—​compare mean annual precipitation in San Diego (32° 44’N) of about 25 centimeters to that in Eureka (40° 48’) of about 100 centimeters. In contrast, air temperatures along the immediate coastline display only a modest variation with latitude of about 7°C due to the moderating influence of the eastern North Pacific.

Interannual Variability of Precipitation From an ecological as well as economic point of view, highly variable precipitation is one of the most important characteristics of California’s climate. The relatively short seasonal window of precipitation (see Figure 2.1) compounds the influence of strong variability in the winter storm track across the northeastern Pacific. Variability in precipitation from synoptic to multiyear time scales in California occurs throughout the state (see Figure 2.2). In relative terms, the variability of California precipitation is the greatest in the United States (Dettinger et al. 2011). This arises from the off-and-on activity of storminess in California and its tendency to take a more

northerly track across the North Pacific in some years and a southerly track in others (Dettinger et al. 1998). Examination of yearly records at locations throughout the state finds frequent, multiyear periods of either deficit or surplus precipitation amounts (see, e.g., Figure 2.2b). Extended periods of either deficit or surplus precipitation can place significant stress on both terrestrial and marine ecosystems. During winter, much of the precipitation that falls over the Sierra Nevada and higher-elevation watersheds in other mountain ranges is in the form of snow. The snow at these high elevations typically remains frozen until temperatures warm later in spring (Lundquist and Loheide 2011) (Figure 2.10). This explains why the water content in the Sierra snowpack usually reaches a maximum during March or April, while precipitation amounts peak during January. During heavy precipitation years such as the great El Niño of 1982–​ 1983, high snow depths can extend into May (Figure 2.10; California Department of Water Resources 1983). The natural water storage provided by mountain snowpack has important implications for many ecosystems, since it ­prolongs the presence of water in the soil and allows for more gradual runoff than occurs in settings that are rainfall-dominated. Snow-fed watersheds produce less damaging flood conditions compared to rainfall basins or those whose snow melts immediately after falling. This is also a major concern of climate change adaptability studies. Projected increases in temperature are expected in California to result in both (1) less snow and more rain, leading to more immediate runoff; and (2) an earlier period of snowmelt (see Chapter 14, “Climate Change Impacts”). Through these effects warmer winters and springs, especially if they include heavy precipitation events, could lead to increased potential for damaging floods. Changes in the timing of the spring runoff due to snowmelt would have important consequences for soil moisture values at many locations in the state. Stewart et al. (2005) describe an advance of snowmelt–​d riven streamflow across much of the western U.S. since 1950, and Hoerling et al. (2013) found that streamflow maxima in many snowmelt-fed streams of the southwestern U.S. occurred earlier during 2001–​2 010, compared to the 1950–​1999 base period. An earlier runoff Climate  19

Snow (cm)

700 600 500 400 300 200 100 0

Heavy snow (1983) Moderate snow (1984) Light snow (1977)

40N

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Mt. Shasta (41.4°N)

Meadow Lake (39.4°N)

35N

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Highland Meadow (38.6°N)

Albedo (%) 5

Jan

Feb

Mar

Apr

May

Jun

Month FIGURE 2.10 Monthly snow depth at three California locations. Bars indicate monthly averages over the 1950–​1999 period. The lines represent selected years when the snow depth was low, moderate, or high. Source: California Department of Water Resources Snow Survey Records.

would generally lead to reduced soil moisture later during in the summer months when soils are already critically dry due to the normal summer drought.

Spatial Variation of Incoming Surface Solar Radiation The amount as well as spatial and temporal distribution of solar radiation supplies the energy needed by organisms to photosynthesize. Solar radiation also provides much of the energy needed to melt snow and to drive evaporative processes. At the top of the atmosphere, the amount of incoming solar radiation is a function of latitude and time of year and is easily calculated. However, how much of that solar radiation reaches the surface is not so trivial. Solar radiation at the surface is determined by slope and aspect of the land surface and is also strongly affected by cloud cover, with aerosol and water vapor concentrations playing a minor role. A major producer of clouds in California is cyclonic storm systems, which are especially frequent and generally best developed in the northern part of the state, and more common in winter than in summer. Rain shadows of numerous mountain ranges inhibit cloud formation and thus add complexity to the distribution of clouds (and of sunshine) within the state (Figure 2.11). The far southeastern part of the state is rarely impacted by northerly cyclonic storms and lies more or less in a large rain shadow, under the influence of a regional descending air mass (described earlier in the chapter). This 20  Drivers

7

9

11

13

15

17

19

21

23

25

FIGURE 2.11 Annual mean cloud albedo (solar reflectance) measured by Geostationary Orbiting Environmental Satellites (GOES) imagery over the 1996–​2 012 period. Source: Satellite data obtained from the National Oceanic and Atmospheric Administration (NOAA) Comprehensive Large Array-data Stewardship System (CLASS) website at and the University of Wisconsin Space Science and Engineering Center.

region has a large number of cloudless days that, in part, determines the high temperatures that prevail in this region. These cloudless zones are among the most arid regions of the state, which have a strong influence on the flora and fauna there and have led to the development of new solar energy generation facilities (see Chapter 30, “Deserts”). Marine stratus clouds play a major role along the coast of California and are a key factor in determining the amount of solar radiation reaching the surface. These clouds, produced by the effects of the North Pacific High and the cool SSTs off the California coast, are most frequent during summer months when available solar radiation is at its highest, which amplifies their impact. They create a marked seasonal increase in cloud albedo (equivalent to decrease in surface solar radiation) along California’s coastline (see Figure 2.11). The contrast is largest in the northern portions of the state and decreases southward. The reduction of solar radiation by marine stratus along the immediate coastline is an important contributor to the relatively cool temperatures in these regions.

Regional Features Heat Waves Summer heat waves impose a major stress on ecosystems in California, often occurring during periods when soil and plant moisture have already declined and temperature is already relatively hot. Heat waves usually occur just a few times in a given summer and typically persist for just a few days, but

they have been observed to last for more than one week (Gershunov et al. 2009). Heat wave events whose strongest expression is during daytime hours have occurred irregularly during the past several decades, but without major upward or downward trends. A rising trend has been observed of nighttime heat waves, anomalously warm during the day but with exceptionally warm nighttime temperatures (Gershunov et al. 2009, Guirguis et al. 2014). The atmospheric circulation anomalies responsible for most great daytime- and nighttime-type California heat waves are remarkably similar, consisting of a prominent anticyclone aloft above Washington State. This feature reinforces a strong, surface pressure gradient between a highpressure anomaly over the Great Plains and a low off the California coast. In contrast to daytime heat waves, which are generally characterized by low relative humidity, nighttime heat waves have been accompanied by heightened humidity. The lack of relief from hot daytime conditions when nighttime temperatures remain elevated can have harsh consequences for ecosystems, livestock, and people.

Piru Padua/Grand Prix Old

Simi

Ventura Oxnard

San Bernadino Riverside

Los Angeles Anaheim Long Beach

Roblar 2 Paradise Escondido

Cedar

San Diego

Otay Tijuana

Santa Ana Winds

Ensenada

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The Santa Ana is a foehn -like wind in southern California that results when high pressure develops over the interior to the east and north of coastal California. Often this configuration arises when a cool, dry air mass flows downslope from high-elevation basins in the western North American interior toward lower atmospheric pressures off the Pacific coast. As the air descends to lower elevations, it undergoes compression and warms. Since the air is sourced from higher altitudes, it is already dry, and as it warms its relative humidity is further reduced. Compression of this air mass through mountain passes often produces winds of 40 to 60 km hr-1 and in extreme cases may yield wind speeds in excess of 100 km hr-1. Hot, dry conditions in summer in California are conducive to wildfires and are especially common in California’s mountain forest zones (Wildfire Today 2013). These summer wildfire conditions are distinct from Santa Ana events (Westerling et al. 2003), which are most common during fall and winter and are especially dangerous in spawning fires at lower elevations west of the mountain fronts in southern California (Hughes et al. 2011). Santa Anas are a regional phenomenon expressed most strongly within the coastal margin and nearby mountain passes of southern California. In many cases Santa Ana winds are set up by much larger-scale weather patterns that drive northeasterly winds along and across the gradient of high-amplitude atmospheric pressure cells that develop in the West Coast and Great Basin region during fall and winter (Abatzoglou et al. 2013). These hot, dry winds frequently occur toward the end of the dry season, when soils and watersheds are at their driest after the long summer drought, and thereby severely increase the environmental stress on many ecosystems. If fuel moistures remain low in early fall due to lack of widespread wetting rains, the potential for Santa Ana–​ driven wildfires heightens. Santa Ana winds are often responsible for the rapid expansion of destructive wildfires, (Westerling et al. 2004) as occurred in southern California in October 2003 (Figure 2.12). The October 2003 fires, emblematic of other strong Santa Ana–​d riven fires in southern California (Keeley et al. 2013), burned vast areas of native grassland and chaparral—​a nd also destroyed many homes and businesses when the paths of these fires ran into developed areas. Because of

-50

20N 160E

180

160W

-250 -200 -150 -100 -50

140W 0

120W

100W

50 100 150 200 250

700 hPa height anomalies FIGURE 2.12 (Top) Smoke from southern California wildfires (named in italics) on October 26, 2003, as seen from NASA/MODIS (Moderate Resolution Imaging Spectroradiometer) satellite imagery. Active fire perimeters are outlined in red. (Bottom) The bottom panel shows the 700 hPa geopotential height anomalies on October 25, 2003 (ignition date). Source: From Westerling et al. 2004.

the extreme winds and drying in Santa Anas, resulting wildfires can rapidly cover enormous areas; the October 2003 fires burned more than 300,000 hectares, destroying ecosystems and causing more than three billion dollars in damage.

Sea Breeze Because of the high heat capacity of water, land surfaces are usually much warmer than the ocean during summer daytime hours. This horizontal land-ocean temperature gradient creates a pressure minimum over the warm land, which sets Climate  21

up a circulation that draws in cool air from over the ocean. This wind system is called a sea breeze and is most common and most pronounced during the summer, when the landocean temperature contrast is at a maximum. The onshore flow of cool marine air helps to moderate the temperatures along the coastal margin, providing cooler and moister conditions on days when the sea breeze develops. Larger-scale synoptic weather patterns can reinforce the sea breeze or create an opposing pressure gradient negating the sea breeze. During periods when the sea breeze is suppressed, coastal areas can become much warmer than normal, increasing stress on ecosystems (and humans) accustomed to the normally cool coastal temperatures. As the sun sets, the land cools more quickly than the ocean, and the temperature gradient can reverse—​w ith low pressure forming over the water and a breeze blowing from land to the water. This is called a land breeze and primarily occurs during winter when the nights are longer, giving the land more time to cool. A particular example of a sea breeze with important impacts on a large region occurs during summer east of the San Francisco Bay. On most summer days, the inland Sacramento Valley quickly warms during morning hours. At the same time, the land-ocean thermal gradient increases rapidly and a sea breeze develops. Advection of this cool moist marine air advances inland through the San Joaquin Delta and often into the western Sacramento Valley and is known as the Delta Breeze. The Delta Breeze is most pervasive when subsiding air from the overhead North Pacific High is weakened, allowing a deeper marine layer. With a deep marine layer, the cool Delta Breeze can penetrate into Sacramento and beyond into the Sacramento Valley, bringing substantial relief from hot summer temperatures.

Tule Fog During winter months, light winds are a typical feature in the San Joaquin and Sacramento Valleys. In winter, the ground is often wet enough to moisten the boundary layer. At night, cold air drains down the sides of the Sierra Nevada to the east and the Coast Ranges to the west, facilitating the formation of a strong nocturnal temperature inversion. With little wind mixing and sufficient humidity, shallow radiation fog easily forms overnight underneath this inversion. This radiation fog is also known as tule fog after the tule grass wetlands in the region. During daylight hours the fog drastically reduces visibility and reflects large amounts on incoming solar radiation, preventing the surface from warming and aiding the persistence of the fog. Although it is relatively shallow (approximately 60 meters thick), the fog can endure through much of the day and often will not dissipate at all, setting the stage for even greater production the following night. The daily cycle of fog is interrupted by increasing winds of approaching winter storms that mix the air and break down the thermal inversion. It is not unusual each winter season to have prolonged periods of fog (two to three weeks) due to a lack of storm activity. Air temperatures in the coastal mountains and Sierra Nevada are typically above normal during such prolonged events due to high-pressure subsidence and ample sunshine on landscapes that lie above the inversion. Fog keeps the surface cool during the day by reflecting sunlight but also keeps the surface warmer at night by absorbing and reemitting

22  Drivers

infrared energy. The resulting consistent, cool temperatures are an asset to fruit tree production, a major agricultural business in these valleys.

Lightning Lightning strikes play a key role in igniting wildfires in California (e.g., Abatzoglou and Brown 2009). The occurrence of lightning varies greatly according to geography and topography, season, and weather pattern (Figure 2.13). Lightning occurrence is highest in summer, peaking in August when surface warming is greatest, and tends to occur most frequently in afternoon hours with peak heating of the land surface and enhanced tendency for rising atmospheric motions. Lightning strikes are more frequent at higher elevations. Weather patterns that promote lightning in most California regions exhibit strengthened high-pressure cells stationed over the western U.S. (van Wagtendonk and Cayan 2008). These circulations deploy moist, monsoonlike air masses, and favor hot afternoon temperatures that promote rising motions, especially over mountain terrain. The occurrence of lightning strikes is highest in California’s mountains and deserts and lowest in coastal and valley regions. In the mountains relatively frequent lightning occurrence falls on forest ecosystems with heavy fuel loads. This combination can result in active wildfires. Fortunately, lightning usually does not occur during strong winds, and lightning is suppressed during Santa Ana events when vertical motions are downward, so this ignition source is quelled during some of the most extreme fire weather conditions.

Summary California’s Mediterranean climate, comprised of cool, wet winters and hot, dry summers, together with numerous mountain ranges and valleys throughout the state produce a unique and diverse set of ecological landscapes. The seasonal pattern of precipitation and temperature is a stress on many ecosystems during summer, when temperatures are conducive to growth but necessary water is lacking. Precipitation in California is strongly linked to seasonal variations in the North Pacific High and the extratropical storm track. During summer months the North Pacific High is at peak intensity and greatest poleward extent, effectively blocking storm systems from California. The North Pacific High weakens and moves equatorward away from California during winter months, allowing storm systems forming over the North Pacific to track across the state. California’s desert regions experience considerably less precipitation than other parts of the state and receive the bulk of their precipitation during the summer months due to thunderstorm activity. Annual precipitation amounts throughout California vary significantly over a range of time scales, from synoptic periods to decades. This extremely high variation is an important factor for ecosystems and subjects them to large year-to-year and even extended, multiyear periods of below- or aboveaverage moisture availability. Some of this strong variability stems from the occurrence of El Niño and La Niña events, which often result, respectively, in significantly higher and lower than normal annual precipitation totals throughout much of the state. Spatially, the variation across the Califor-

A

0-5 Strikes yr-1 100 km-2 5-10 Strikes yr-1 100 km-2 10-15 Strikes yr-1 100 km-2 15-20 Strikes yr-1 100 km-2 20-25 Strikes yr-1 100 km-2 25-30 Strikes yr-1 100 km-2 30-35 Strikes yr-1 100 km-2

B 1000000

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Central Valley

North Coast

Southeast Deserts

Northeast Plateaus

Klamath Mountains

Cascade Range

Sierra Nevada

200 kilometers

FIGURE 2.13 Lightning strikes. Source: From van Wagtendonk and Cayan 2008. A Lightning strike density (strikes yr-1 100 km-2) in California, for the period 1985 through 2000. B Number of lightning strikes by month and bioregion in California, for the period 1985 through 2000. The monthly pattern is consistent

with more strikes occurring during the summer months than during the winter months.

nia landscape of temperature and precipitation is strongly associated with latitude and topography. There is a general trend of cooler temperatures and increasing precipitation with increasing latitude. However, topography plays a strong role also, as the height and orientation of the numerous mountain ranges have profound influence on precipitation distribution. In general, as winds within storm systems encounter mountainous terrain, precipitation is enhanced on windward slopes and reduced on the leeward sides. A diverse array of regional climatic features throughout California have a profound impact on ecosystems. Cloud cover from wintertime cyclonic storms and summer marine stratus along California’s coast strongly regulate the solar radiation reaching the surface. Santa Ana winds aggravate wildfire threat by producing unseasonably dry and warm conditions at many locations in southern California. During summer months sea breezes transport cool marine air along California’s coastal regions. Prolonged periods of tule fog often occur each winter in the San Joaquin and Sacramento Valleys, reducing visibility and resulting in consistently cool daytime and nighttime temperatures. Thunderstorm activity, most commonly in California’s mountain and desert regions, reaches a maximum frequency during summer months, with associated lightning strikes playing an important role in igniting wildfires in California.

Acknowledgments The authors would like to acknowledge the California Nevada Applications Program (CNAP), one of the NOAA RISA centers, and the Southwest Climate Science Center (SWCSC), sponsored by the Department of Interior, for funding a portion of this work.

Recommended Reading Cayan, D. R., K. T. Redmond, and L. G. Riddle. 1999. ENSO and hydrologic extremes in the Western United States. Journal of Climate 12:2881–​2 893. Dettinger, M. D., F. M. Ralph, T. Das, P. J. Neiman, and D. R. Cayan. 2011. Atmospheric rivers, floods, and the water resources of California. Water 3:445–​478. Gershunov, A., D. R. Cayan, and S. F. Iacobellis. 2009. The Great 2006 heat wave over California and Nevada: Signal of an increasing trend. Journal of Climate 22:6181–​6203. Iacobellis, S. F., and D. R. Cayan. 2013. The variability of California summertime marine stratus: Impacts on surface air temperatures. Journal of Geophysical Research: Atmospheres 118:9105–​9122. Luers, A. L., D. Cayan, G. Franco, M. Hanemann, and B. Croes. 2006. Our changing climate—​assessing the risks to California: A summary report from the California Climate Change Center, California Energy Commission Report CEC-500-2006-077. . Accessed March 12, 2015.

Climate  23

Minnich, R. A. 2006. California climate and fire weather. Pages 13–​ 37 in N. G. Sugihara, J. W. VanWagtendonk, K. E. Shaffer, J. FitesKaufman, and A. E. Thode, editors. Fire in California’s ecosystems. University of California Press, Berkeley, California. Pierce, D. W., editor. 2012. California climate extremes workshop report. Scripps Institution of Oceanography. . Accessed March 12, 2015. Westerling, A. L., D. R. Cayan, T. J. Brown, B. L. Hall, and L. G. Riddle. 2004. Climate, Santa Ana winds, and autumn wildfires in southern California. EOS, Transactions, American Geophysical Union 85:289–​296.

Glossary Adiabatic  An adiabatic process is one that takes place without heat being exchanged between the system and its surroundings. Advection  The horizontal transfer of any atmospheric property by the wind. Aerosol  Solid particles suspended in atmosphere. These tiny particles are produced from both natural or anthropogenic sources such as fossil fuel burning. Albedo  The percentage of incident solar radiation on a surface that is reflected back into space. Catalina Eddy  A localized circulation feature along the southern California coast. The Catalina Eddy develops when northwesterly flow along the central California coast moves south of Point Conception and develops cyclonic circulation centered near Santa Catalina Island. Convective  Fluid motions that transport and mix properties of the fluid. Convection is often associated with upward vertical motion initiated by warm surface temperatures. Cyclonic  Circulation that moves in a counterclockwise direction in the Northern Hemisphere (clockwise in the Southern Hemisphere) and is associated with low pressure systems and rising air. Foehn  A wind that descends along the leeside of a mountain range. As the air descends, the pressure increases and the temperature rises due to adiabatic compression. Heat capacity  The amount of heat absorbed (or released) by an object divided by the temperature increase (or decrease). Infrared radiation  Electromagnetic radiation with wavelengths between 0.7 and 1000 microns. In general, components of Earth’s climate system are at temperatures that cause them to emit radiation in the infrared range. Jet stream  A concentrated band of strong winds in the atmosphere. While there are several different jet streams in the atmosphere, the term usually refers to the polar jet stream, which is conducive for the formation of mid-latitude cyclonic storm systems. Leeward  Refers to the direction in which the wind is moving. Also known as “downwind.” Marine stratus  These are low sheet-like clouds that form over ocean surfaces. These clouds are generally horizontally expansive with limited vertical growth. Microclimate  The climate of a localized region that differs in some respect from the climate of the surrounding area. Monsoon  A wind circulation system that changes direction seasonally and is caused by the uneven heating rates of land and water regions. North Pacific High  A semipermanent area of high pressure centered around 30°N off the west coast of North America. The location and intensity vary seasonally.

24  Drivers

Northwesterly wind  Wind that originates from the northwest and moves toward the southeast. This convention (direction ending with “-ly”) is also applied to other wind directions. Orographic uplift  The uplift of air as it encounters such terrain features as mountains. Pressure gradient  The change in pressure over a distance. Saturation point  The maximum amount of water vapor that can exist in an air mass for a given temperature and pressure. Synoptic systems  Weather systems with horizontal-length scales from several hundred to several hundred to a few thousand kilometers and time scales of days to a week or more. Hurricanes and mid-latitude cyclones are examples of synoptic weather systems. Temperature inversion  A vertical region of the atmosphere where the temperature increases with height. A temperature inversion represents very stable conditions that can limit the vertical mixing of air beneath. Tradewinds  The prevailing surface winds found in the tropical regions between the equator and approximately 25°. These winds generally are northeasterly in the Northern Hemisphere and southeasterly in the Southern Hemisphere. Upwelling  Ocean water that rises from depth to replace surface water. Upwelled water is generally cold and nutrient rich. Windward  Refers to the direction from which the wind originates. Also known as “upwind.”

References Abatzoglou, J. T., and T. J. Brown. 2009. Influence of the MaddenJulian oscillation on summertime cloud-to-ground lightning activity over the continental United States. Monthly Weather Review 137:3596– ​3601. Abatzoglou, J. T., K. T. Redmond, and L. M. Edwards. 2009. Classification of regional climate variability in the state of California. Journal of Applied Meteorology and Climatology 48:1527–​1541. Abatzoglou, J. T., R. Barbero, and N. J. Nauslar. 2013. Diagnosing Santa Ana winds in southern California with synoptic-scale analysis. Weather and Forecasting 28:704–​710. Adams, D. K., and A. C. Comrie. 1997. The North American monsoon. Bulletin of the American Meteorological Society 78:2197–​2213. Brewer, W. H. 1930. Up and down California in 1860–​1864. Yale University Press, New Haven, Connecticut. Bromirski, P. D., R. E. Flick, and D. R. Cayan. 2003. Storminess variability along the California coast, 1858–​2 000. Journal of Climate 16:982–​993. California Department of Water Resources. 1983. Water conditions in California, Report 4. May 1, 1983. . June 7, 2015. Cayan, D. R., and L. Riddle. 1993. Atmospheric circulation and precipitation in the Sierra Nevada. Pages 771–​720 in R. Herrmann, editor. Managing water resources during global change: An international conference: AWRA 28th annual conference and symposium. American Water Resources Association, Middleburg, Virginia. Cayan, D. R., K. T. Redmond, and L. G. Riddle. 1999. ENSO and hydrologic extremes in the western United States. Journal of Climate 12:2881–​2 893. Das, A. J., N. L. Stephenson, A. Flint, T. Das, and P. J. van Mantgem. 2013. Climatic correlates of tree mortality in water- and energylimited forests. PLOS ONE 8(7), e69917. . Dettinger, M. D. 2013. Atmospheric rivers as drought busters on the US west coast. Journal of Hydrometeorology 14:1721–​1732. Dettinger, M. D., and B. L. Ingram. 2013. The coming megafloods. Scientific American 308:64–​71.

Dettinger, M. D., D. R. Cayan, H. F. Diaz, and D. M. Meko. 1998. North-south precipitation patterns in western North America on interannual-to-decadal timescales. Journal of Climate 11:3095–​3111. Dettinger, M. D., F. M. Ralph, T. Das, P. J. Neiman, and D. R. Cayan. 2011. Atmospheric rivers, floods, and the water resources of California. Water 3:445–​478. Dettinger, M., K. Redmond, and D. Cayan. 2004. Winter orographic precipitation ratios in the Sierra Nevada—​Large-scale atmospheric circulations and hydrologic consequences. Journal of Hydrometeorology 5:1102–​1116. di Castri, F., and H. A. Mooney. 1973. Mediterranean type ecosystems: Origin and structure. Springer Verlag, Berlin, Germany. Flick, R. E. 1998. Comparison of tides, storm surges, and mean sea level during the El Niño winters of 1982–​83 and 1997–​98. Shore and Beach 66:7–​17. Gershunov, A., and T. P. Barnett. 1998. ENSO influence on intraseasonal extreme rainfall and temperature frequencies in the contiguous United States: Observations and model results. Journal of Climate 11:1575–​1586. Gershunov, A., D. R. Cayan, and S. F. Iacobellis. 2009. The great 2006 heat wave over California and Nevada: Signal of an increasing trend. Journal of Climate 22:6181–​6203. Guirguis, K., A. Gershunov, A. Tardy, and R. Basu. 2014. The impact of recent heat waves on human health in California. Journal of Applied Meteorology and Climatology 53:3–​19. Hidalgo, H. G., D. R. Cayan, and M. D. Dettinger. 2005. Sources of variability of evapotranspiration in California. Journal of Hydrometeorology 6:3–​19. Hoerling, M. P., M. Dettinger, K. Wolter, J. Lukas, J. Eischeid, R. Nemani, B. Liebmann, and K. E. Kunkel. 2013. Present weather and climate: Evolving conditions. Pages 74–​100 in G. Garfin, A. Jardine, R. Merideth, M. Black, and S. LeRoy, editors. Assessment of Climate Change in the Southwest United States: A Report Prepared for the National Climate Assessment. Island Press, Washington D.C. Hughes, M., A. Hall, and J. Kim. 2011. Human-induced changes in wind, temperature, and relative humidity during Santa Ana events. Climatic Change 109:S119–​S132. Iacobellis, S. F., and D. R. Cayan. 2013. The variability of California summertime marine stratus: Impacts on surface air temperatures. Journal of Geophysical Research: Atmospheres 118:9105–​9122. Iacobellis, S. F., J. R. Norris, M. Kanamitsu, M. Tyree, and D. R. Cayan. 2009. Climate variability and California low-level temperature inversions. California Climate Change Center Publication CEC-500-2009-020-F. . June 7, 2015. Kanamitsu, M., E. Yulaeva, and H. Li. 2013. Catalina Eddy as revealed by the historical downscaling of reanalysis. Asia-Pacific Journal of Atmospheric Sciences 49:467–​481. Keeley, J., A. Syphard, and C. J. Fotheringham. 2013. The 2003 and 2007 wildfires in southern California. Pages 42–​52 in S. Boulter et al., editors. Natural Disasters and Adaptations to Climate Change. Cambridge University Press, New York. Lebassi, B., J. Gonzalez, D. Fabris, E. Maurer, N. Miller, C. Milesi, P. Switzer, and R. Bornstein. 2009. Observed 1970–​2 005 cooling of summer daytime temperature in coastal California. Journal of Climate 22:3558– ​3573. Lundquist, J. D, and S. P. Loheide. 2011. How evaporative water losses vary between wet and dry water years as a function of elevation in the Sierra Nevada, California, and critical factors for modeling. Water Resources Research 47. . Minnich, R. A. 2006. California climate and fire weather. Pages 13–​

37 in N. G. Sugihara, J. W. Van Wagtendonk, K. E. Shaffer, J. FitesKaufman, and A. E. Thode, editors. Fire in California’s ecosystems. University of California Press, Berkeley, California. Monteverdi, J., and J. Null. 1997. El Niño and California rainfall. NOAA Western Region Technical Attachment, No. 97-37. November 21, 1997. National Research Council. 2012. Sea-level rise for the coasts of California, Oregon, and Washington: Past, present, and future. National Academies Press, Washington, D.C. Pandey, G. R., D. R. Cayan, and K. P. Georgakakos. 1999. Precipitation structure in the Sierra Nevada of California during winter. Journal of Geophysical Research 104:12019–​12030. Philander, S. G. 1990. El Nino, La Nina, and the southern oscillation. Academic Press, Waltham, Massachusetts. Ralph, F. M., P. J. Neiman, G. A. Wick, S. I. Gutman, M. D. Dettinger, D. R. Cayan, and A. B. White. 2006. Flooding on California’s Russian River: Role of atmospheric rivers. Geophysical Research Letters 33. . Rasmussen, E. M., and J. M. Wallace. 1983. Meteorological aspects of the El Nino/southern oscillation. Science 222:1195–​1202. Rasmussen, E. M., and T. H. Carpenter. 1982. Variations in tropical sea surface temperature and surface wind fields associated with the southern oscillation/El Nino. Monthly Weather Review 110:354–​384. Redmond, K. T., and R. W. Koch. 1991. Surface climate and streamflow variability in the western United States and their relationship to large-scale circulation indices. Water Resources Research 27:2381–​2399. Roden, G. I. 1967. On river discharge into northeastern Pacific Ocean and Bering Sea. Journal of Geophysical Research 72:5613–​5629. Sette, O. E., and J. D. Isaacs, editors. 1960. Symposium on “The changing Pacific ocean in 1957 and 1958.” CalCOFI Reports 7:13–​217. Stahle, D. W., R. D. Griffin, D. M. Meko, M. D. Therrell, J. R. Edmondson, M. K. Cleaveland, L. N. Stahle, D. J. Burnette, J. T. Abatzoglou, K. T. Redmond, M. D. Dettinger, and D. R. Cayan. 2013. The ancient blue oak woodlands of California: Longevity and hydroclimate history. Earth Interactions 17. . Stewart, I. T., D. R. Cayan, and M. D. Dettinger. 2005. Changes toward earlier streamflow timing across Western North America. Journal of Climate 18:1136–​1155. Thorne, J. H., J. H. Viers, J. Price, and D. M. Stoms. 2009. Spatial patterns of endemic plants in California. Natural Areas Journal 29:344– ​366. U.S. Geological Survey (USGS). 2005. Southern California—​w ildfires and debris flows. U.S. Geological Survey Fact Sheet 2005–​3106. van Wagtendonk, J. W., and D. R. Cayan. 2008. Temporal and spatial distribution of lightning strikes in California in relation to largescale weather patterns. Fire Ecology 4:34–​56. Westerling, A.L., D. R. Cayan, T. J. Brown, B. L. Hall, and L. G. Riddle. 2004. Climate, Santa Ana winds and autumn wildfires in southern California. EOS, Transactions, American Geophysical Union 85:289–​296. Westerling, A. L., T. J. Brown, A. Gershunov, D. R. Cayan, and M. D. Dettinger. 2003. Climate and wildfire in the western United States. Bulletin of the American Meteorological Society 84:595–​604. Wildfire Today. 2013. Wildfire today, California: Rim fire at Yosemite NP. . Accessed November 13, 2014. Wolter, K., and M. S. Timlin. 2011. El Niño/southern oscillation behaviour since 1871 as diagnosed in an extended multivariate ENSO index (MEI.ext). International Journal of Climatology 31:1074–​1087.

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THREE

Fire as an Ecosystem Process JON E . KEELE Y and HUGH D. SAFFORD

Introduction Any land surface with sufficient precipitation to produce moderate levels of plant biomass and a seasonal climate that dries the vegetation is likely to be fire-prone. California’s Mediterranean-type climate makes the landscape particularly predisposed to fires: rains occur during winter and because of the ocean influence, temperatures are relatively mild, leading to abundant plant growth that can produce densely vegetated landscapes of potential fuels. The warm, annual summer drought makes this vegetation highly flammable. In combination, these factors contribute to frequent wildfires that can burn extensive parts of the landscape. Droughts in California are an annual event that result in a more predictable fire regime than in most other parts of the U.S., where severe fire weather might occur only every few years, decades, or centuries due to long-period oscillations in synoptic weather conditions (Minnich 2006, Keeley et al. 2012). Fire is an important ecosystem process in many California ecosystems (Sugihara et al. 2006). Prior to Euro-American settlement, more than 40% of the state supported high fire frequencies (fire return intervals less than thirty-five years on average), and another 15%–​20% supported moderate fire frequencies (fire return intervals of thirty-five to one hun-

dred years) (Table 3.1). Ecosystems that burned infrequently include subalpine forests, moist forests in northwesternmost California, and desert vegetation. Fire influences ecosystem composition, structure, and function in many ways, and ecosystems themselves strongly influence fire. As a disturbance, fire is unique in that its intensity and frequency depend on the growth rate of the medium (vegetation) it destroys. Because of this, there is a broadly inverse relationship between fire frequency and intensity, with the strength of that relationship varying by ecosystem type (Huston 2003). Worldwide, fire is a keystone process in Mediterranean-type climate ecosystems, determining structural and distributional patterns of both flora and fauna and influencing biodiversity on both ecological and evolutionary time scales (Keeley et al. 2012). In many ecosystems fire is a principal consumer of plant biomass, and it removes, recycles, and renews various nutrients in plants and soils (Sugihara et al. 2006). In many ways it is an important component of the trophic pyramid in ecosystems in that it competes with other herbivores (Bond and Keeley 2005). Although fire is a natural process integral to the long-term sustainability of many California ecosystems, it is mislead27

TA B L E 3 .1 Major fire regime groupings for California and their percentage of California area and mean pre-Euro-American settlement fire return intervals (FRI)

Fire regime (frequency and severity)

Vegetation-type groups

High frequency/low to moderate severity

Yellow pine, mixed conifer

Percentage of Presettlement FRIB California area A (years)

Oak woodland

14.7

13.5C

9.3

12

12.5

100–​ 150 meter radius), conifer and hardwood recruitment can be delayed by many years and shrubs can dominate much of the site, further delaying forest succession. Assuming they do not reburn, very large areas of stand-replacing fire can remain free of forest for many decades.

Fire Effects on Animal Communities Animals exhibit a diversity of strategies for dealing with fire. Some invertebrates can persist as dormant diaspores in the soil, similar to the bulbs of many herbaceous perennial geophytes. Smaller mammals “shelter-in-place” and survive by seeking refugia such as rock outcrops, moist ravines, and burrows within the burn perimeter. Others, including birds and larger mammals, flee the fire and must subsequently recolonize from the unburned landscape. As a consequence, animals can exhibit different sensitivities to fire regime attributes than do plants within the same ecosystem types. For example, recovery of chaparral vegetation generally is not greatly affected by fire size. However, for chaparral animals that must flee the fire and recolonize afterward, fire size and size of high-severity patches can be immensely important. In contrast to chaparral plants, for some animals the metapopulation dynamics of source populations in the surrounding landscape, including their sizes and dispersion as well as their

Disturbance event

Period of high abundance or diversity

Relative population size

connectivity to each other and the burned site, are critically important to recovery rate after large fire events. The degree of habitat fragmentation outside the burn perimeter can also strongly affect animal recolonization (Sauvajot et al. 1998). For example, habitat of the federally endangered California gnatcatcher (Polioptila californica) in mature sage scrub has been severely impacted by urban growth, diminishing source populations, and thus its potential for recolonizing burned areas (Beyers and Wirtz 1995). Fire behavior can also affect animal recovery. Fast-moving fires can harm animals that must flee the fire but might have little direct effect on species that shelter-in-place in burrows, mesic riparian sites, or other refugia such as rock outcrops. For these species slow-moving fires often generate greater volumes of smoke and longer lasting heat, which can greatly reduce survival (Shaffer and Laudenslayer 2006). Seasonality is another fire regime attribute that can play a greater role in animal than plant response. For example, nesting birds and other animals with new offspring are particularly vulnerable to spring burning in many vegetation types (Smith et al. 2000, Shaffer and Laudenslayer 2006). Seasonality also affects fire behavior; spring fires are likely to burn vegetation with greater moisture and potentially greater smoke production. Changes in fire seasonality can threaten some animal species. For example, the sage grouse (Centrocercus urophasianus), once a very common bird in the western U.S. and Canada, is now a candidate for listing under the Endangered Species Act. This bird depends on Great Basin shrublands dominated by sagebrush (Artemisia), which historically experienced summer and fall burns. With the invasion of the annual cheatgrass (Bromus tectorum), which produces contiguous, fine fuels, the fire season begins earlier and fires lack historical patchiness (see Chapter 30, “Deserts”). The new fire regime disadvantages both the sagebrush and the grouse that depends on it (Baker 2006). Some components of the fauna are well-insulated from fire effects. For example, some flightless insects like the wingless walking-stick (Timema cristinae) enter diapause and survive fires as a tough egg stage buried in the soil (Sandoval 2000); many other arthropods likely survive fire in a similar manner. In this respect, these invertebrates are like plants such as geophytes that persist through the dry season (and consequently the fire season) as dormant bulbs buried in the soil. Each year they might emerge during the wet season, regardless of whether or not a fire occurred. Once fire has passed, animal recovery is influenced by the magnitude of changes in vegetation structure (Smith et al. 2000). Ground-dwelling herbivores and granivores are often food-limited for many months following fire and can be forced to temporarily migrate outside the burn perimeter for forage. As a consequence, animal recovery can often parallel plant recovery (e.g., Rochester et al. 2010). In some animals such as deer the enhanced nutritional value of plant regrowth leads to significant population increases (Hiehle 1961). In crown fire ecosystems the loss of vegetation cover can have profound impacts on visibility of foragers to predators and can enhance mammalian carnivore populations (Schuette et al. 2014) as well as hawks and owls (Lyon et al. 2000). Indeed, postfire habitats are the preferred habitat for a diversity of organisms. Noteworthy are bark beetles like Dendroctonus and Ips species (Furniss and Carolin 1977) and fire beetles (Melanophila species), which are drawn to forest fires even while they are still burning and seek out sites for ovipositing on burned logs (Hart 1998). Other forest species attracted to

0

Period of transition to canopy users Understory and old growth species

Edge species Open site species

Time since disturbance

FIGURE 3.6 Hypothetical patterns of change in animal populations relative to postfire changes in forest structure. Source: Huff and Smith 2000.

high-severity crown fires include several woodpecker species adapted to forage on recently charred trees (Smucker et al. 2005). Certain species groups benefit from the habitats and food sources (e.g., bark beetles) that result when most of the dominant vegetation is killed. Examples include many rodents, woodpeckers, and other birds that forage on the ground or in brush (Raphael et al. 1987, Smith et al. 2000). However, in forests with mixtures of surface and crown fire, increasing proportions of the latter can greatly slow recolonization by many animal species (Smith et al. 2000). Some of the most contentious resource management issues in California revolve around the increasing severity of fire in low- and middle-elevation conifer forests caused by the combined effects of fire suppression and climate warming (Miller et al. 2009, Van Mantgem et al. 2013). Management actions are increasingly necessary to ameliorate the effects of severe fire on populations of rare mesocarnivores that require older, longunburned forest for part of their life cycle (e.g., spotted owl [Strix occidentalis], Pacific fisher [Martes pennant], and goshawk [Accipiter gentilis]). Though few long-term studies have examined animal responses to fire in California’s ecosystems, work in other regions shows that species-specific animal traits result in changing peak abundances with time since fire. For example, long-term studies of postfire animal responses in the Mediterranean-type shrublands in Australia found that fire influences reptile occurrence over century-long time scales (Nimmo et al. 2012). Preliminary postfire studies in California chaparral show that some species dominate early, while others depend more on later successional stages (Rochester et al. 2010). Huff and Smith (2000) have pointed to the diversity of vertebrates’ responses to fire; some initially increase, some decrease, and others remain unchanged (Figure 3.6). Fire can also have important, mostly indirect effects on organisms living in streams and lakes. Postfire increases in turbidity and nutrients change water quality, and channel scouring from debris flows and heavy sediment loads can greatly modify aquatic habitat. Major sediment and nutrient pulses after fire can kill aquatic organisms, but most aquatic effects of fires are ephemeral. Studies of stream impacts of fire find relatively little in the way of long-term damage to aquatic Fire as an Ecosystem Process   37

ecosystems, except where burned landscapes are already compromised by other stressors such as human land use (Gresswell 1999, Minshall 2003).

Fire Effects on Soils, Hydrology, and Carbon Storage Depending on goals and on the variables and time period of interest, fire can have both positive and negative effects on soil, water, and carbon resources. Here we summarize some of the most salient fire effects, but for a more comprehensive treatment we direct the reader to excellent summaries in Debano et al. (1998) and Wohlgemuth et al. (2006).

Soils Fire effects on soil depend on the amount and duration of heat output in conjunction with the physical and chemical properties of the soil itself. Temperatures of 100°C are lethal to most organisms, and these temperatures are readily exceeded by fires at the soil surface. However, temperature pulses are ephemeral, and soil depth strongly attenuates heating as well. At 5 centimeter depth, wildfire-caused temperatures rarely exceed the 100°C threshold, except where fuels are heavy or large logs induce long-term smoldering. Low-intensity fires burning in surface litter and/or herbaceous vegetation, as were historically common in yellow pine and mixed-conifer ecosystems, have little soil impact (except in the surface litter), whereas today’s fires of typically higher intensity have deeper, more widespread effects on soils. High-intensity fires common in chaparral and in conifer forest areas with heavy fuels can drive surface temperatures above 500°C for periods of ten to twenty minutes and temperatures at 2.5 centimeter depth to 100°C to 200°C (Debano 1981). At these higher surface temperatures most nitrogen and organic phosphorus volatilize, and soil physical and chemical structures are greatly altered. However, recent studies have shown that the relationship between fuel load and soil temperatures during fire are much more complicated than previously assumed (Stoof et al. 2013). Different soils and soil conditions vary in sensitivity to the heating effects of fire. Coarse textured soils with relatively low carbon content, such as are common in soils weathered from granitic rocks in many of California’s major mountain ranges, generally show little fire-induced change in their bulk density and relatively little change in porosity. Finer-textured soils (e.g., those with higher clay content) with higher levels of carbon are much more sensitive to high temperatures, and their soil structure can be greatly altered, leading to reduced water infiltration and postfire erosion under sustained rainfall (Wohlgemuth et al. 2006). Soil water content is also an important variable, as moist soils are greatly protected from the downward transmission of heat from fire (Busse et al. 2005). Fire effects on soil water repellency, known as soil hydrophobicity, depend on complex interactions among fire heating, soil and plant chemistry, soil texture and moisture, and characteristics of the litter layer. Fire’s most significant impacts to soil chemistry revolve around soil organic matter and macronutrients. Fire consumes soil litter and carbon compounds in the upper soil, causing important changes in soil structure (loss of aggregation), reduced soil water-holding capacity, and reduced cat38  Drivers

ion exchange capacity. Severely burned soils with high loss

of organics can be largely transformed to ash, and windy conditions during and after fire can lead to local soil losses. The loss of carbon also produces major changes in the soil microflora, which in turn typically causes a release of soluble nitrogen compounds. Nitrogen and phosphorus are largely volatilized by fire, but the net loss of both minerals is often offset by an increase in their soluble, bio-available forms. A (usually short-term) flush of important nutrients is thus typically available to those plants that can use them quickly, such as resprouters and seedlings that germinate in the first year after fire. Much of the nutrient pulse is lost to water transport offsite during the rainy season (Wohlgemuth et al. 2006).

Water Fire’s effects on hydrology depend to a great extent on the soil effects described earlier and on vegetation responses. Since fire removes biomass, it can greatly increase the proportion of precipitation that directly hits the soil. Decreased canopy interception of rain and loss of litter and soil organic matter, which reduce soil infiltration, increase the erosional actions of rainsplash and the transport of water by overland flow. Lower plant cover leads to greater ground accumulation of snow (depending also on wind patterns) but also speeds melting during warm, sunny weather. Burning changes the balance between evaporation and transpiration, increasing the former while reducing the latter. Dry springs commonly begin to flow again after fire, and stream levels are usually enhanced, especially during storm events. Extensive scientific study of this effect has documented over one-thousand-fold increases in stream flow depending on factors including burn severity and precipitation patterns (Tiedemann et al. 1979). One of the major impacts of fire is that it often initiates sediment transport processes including dry ravel, erosion, and debris flows (Cannon et al. 2010, Moody and Martin 2001, Lamb et al. 2013). Models to predict these processes include many soil characteristics such as particle size; organic matter content; soil thickness; underlying rock type; slope incline and aspect; rainfall frequency, duration, and intensity; prefire vegetation and fire severity impacts. Contributing to sediment loss is hydrophobicity. In high-intensity chaparral fires, water-repellent chemicals on the surface can be destroyed and reformed as a hydrophobic layer at deeper levels, leading to major surface soil erosion when subsequent rainfall saturates the upper soil and is then detoured horizontally by the hydrophobic lens. The importance of this layer is greatly influenced by subsequent rainfall patterns and in some cases might not contribute significantly to soil loss (Hubbert and Oriol 2005). Although hillslope erosion is greatly increased by fire, it can return to prefire levels within a few years. The timing of extreme precipitation events in relation to the fire is an important variable (Oliver et al. 2012). Since stream flows are so sensitive to fire, stream channel erosion and sediment transport are both greatly affected by fire as well. A typical pattern is one where hillslope erosion fills smaller channels with sediment, while later storms promote downcutting through these deposits, transporting the accumulated sediment to lower stream reaches. Sometimes channel scour can occur catastrophically, in the form of fast-moving and very destructive debris flows (Wohlgemuth et al. 2006). Finally, fire affects water quality. Suspended solid particles, mostly ash and soil, increase stream turbidity and often carry high levels of soil

nutrients into the stream system. This erosional pulse is usually short-lived, however. Longer-term effects to streams leaving the fire area include elevated temperatures due to loss of vegetation shading and increases in dissolved concentrations of various nutrients such as nitrogen, phosphorus, and calcium (Tiedemann et al. 1979, Oliver et al. 2012).

Carbon Since fire burns biomass, it always results in a short-term loss of carbon. The actual amount lost to burning is highly variable, and overall carbon balances for a given ecosystem are generally expected to equilibrate over time (due to vegetation regrowth) as long as the fire regime is not changing (Kashian et al. 2006). Carbon losses from the soil can strongly affect ecosystem processes. For example, lost organic compounds reduce the cation exchange capacity of the soil and the ability of water to infiltrate and be retained within the soil. Carbon loss also affects nutrient availability for plants, as less carbon substrate for microbes causes bacterial and fungal mortality and subsequent release of nitrogen into the soil (Wohlgemuth et al. 2006). Carbon losses due to fire include both direct losses to combustion and longer-term losses to decomposition. Combustion of biomass averages 20% to 30% for tree boles and large branches in temperate forests affected by high-severity fires, while greater proportions of needles/leaves, cones, and fine branches are consumed (Tinker and Knight 2000, Schuur et al. 2003, Stephens, Boerner, Moghaddas et al. 2012). Carbon combustion also affects the organic layers in soil; these losses can exceed 60% of the prefire carbon. Over the longer term, vegetation killed by fire will decompose, and over decades the overall carbon emissions from decomposition can be five times greater or more than the carbon lost to combustion (Auclair and Carter 1993). Attempts to reduce fire-related carbon loss via forest thinning and other fuel treatments are most likely to realize actual long-term carbon balance gains in ecosystems that have ecological and evolutionary associations with frequent, low-intensity fire, such as ponderosa pine or mixedconifer forests. Such practices are more likely to negatively affect long-term carbon sequestration in ecosystems where fire is rare and/or combustion typically less complete, such as wet forests of the northwest coast and coastal mountains (Hurteau et al. 2008, Mitchell et al. 2009). Changes towards increased summer drought and fire frequency in these wet forests—​as are generally projected by future climate and fire models—​could increase the amount of biomass held in young successional forests, which would reduce carbon sequestration on the ground and increase emissions to the atmosphere.

Fire in Social Context Fires pose an important hazard for people in California, in a state prone to many hazards such as earthquakes, floods, and landslides. Based on written records from the late nineteenth century, the largest fire in the state’s history occurred in Orange County in 1889, burning more than 100,000 hectares. No one died, and no homes were lost, from that fire (Keeley and Zedler 2009). Today, such large wildfires still occur, but due to explosive population growth during the twentieth century (see Chapter 5, “Population and Land Use”), they now account for substantial losses of property and lives. Since the

middle of the twentieth century, the state has averaged more than five hundred homes lost per year from fires, and in the past decade the rate has doubled (Keeley et al. 2013). Most of California’s population growth has occurred in the lower-elevation foothills, valleys, and coastal plains, where humans are responsible for igniting nearly all fires (see Table 3.2). Thus one of the important means of reducing fire losses has been through public campaigns aimed at fire prevention. Since the early twentieth century, county, state, and federal agencies have taken on the responsibility of fighting fires, and suppression tactics have grown increasingly sophisticated over this time. Thus prevention and suppression are the primary responses to fires that threaten human developments. In past several decades there has been increasing emphasis on trying to alter fire outcomes through vegetation management that seeks to change fuel patterns on the landscape. More recently, emphasis has been placed on greater personal responsibility of homeowners to increase the fire resistance of their homes through construction improvements and landscape maintenance. The last avenue for managing fire risk to urban developments has been through changes in local community planning decisions. Just as the history of flood hazards in the state has been partially solved through flood zoning regulations, the future may hold greater fire zoning restrictions as one part of the fire solution.

Fire Management Because of the prevalence of fires in California and mounting losses of resources, property, and lives, fire management has unavoidably become a central focus of federal, state, and local agencies across California. Together, these agencies spend over three billion dollars annually in wildfire prevention and suppression, and the costs of fighting fire are rising rapidly (Safford 2007).

Prefire Prefire management essentially involves manipulating fuels on the landscape with the goals of preventing ignitions from becoming significant fires, allowing for quicker suppression of fires that do occur and reducing the amount of resource damage caused by fires. Prefire fuel treatments can be very effective in slowing or stopping fire and protecting human infrastructure, depending on location, method, and ease and safety of firefighter access (Syphard et al. 2011, Martinson and Omi 2013). In frequent-fire forest ecosystems (e.g., yellow pine, mixed conifer), typical fuel treatment practices include mechanized thinning of medium and smaller trees (usually focused on shade-tolerant species), hand or mechanical reduction of surface fuels, and prescribed fire or fuel pile burning. Data from both northern and southern California show that treatments that incorporate all or most of these steps are effective at reducing fire severity and carbon loss to fire (North and Hurteau 2011, Carlson et al. 2012, Safford et al. 2012). Since fire regimes, forest structure and composition, and fuel loadings in these forest types have been greatly modified by human intervention, fuel reduction treatments—​especially where they incorporate some form of fire reintroduction—​a re generally benign and even restorative in terms of their ecological and environmental effects (Stephens, McIver, Boerner et al. 2012). Various meta-analyses Fire as an Ecosystem Process   39

have found mostly positive or neutral effects of these types of treatments on a wide variety of taxonomic groups (Kalies et al. 2010, Verschuyl et al. 2011). On the other hand, fuel treatments in shrubland landscapes that are burning much more frequently today than before Euro-American settlement—​the general case in sage scrub and chaparral landscapes in southern California—​contribute little to ecological restoration and in some cases may cause environmental damage. For example, invasive species generally increase with reduced native shrub cover (Merriam et al. 2006, Keeley and Brennan 2012), and areas with reduced shrub cover, whether by fire or treatment, are sources of hillslope erosion, stream sedimentation, and stream channel debris flows in the wet season (Wohlgemuth et al. 2006). In such highly flammable and often densely populated landscapes, fuel treatments are a management necessity to protect human lives and property, but strategic thinking is necessary to properly balance the positive features of treatment networks with their potential negative environmental impacts. Fuel treatments in places like southern California can also be conducted for ecological purposes—​for example, to protect older stands of chaparral or coastal sage scrub or to shield regenerating populations of species threatened by overly frequent fire, such as cypresses (e.g., Tecate cypress, Hesperocyparis forbesii) or rare, fire-sensitive habitats. In these cases the treatments themselves would be viewed as environmental sacrifices made on a local scale to realize landscape-level ecological benefits.

Postfire Human management of postfire landscapes can have major effects on ecosystem succession and can generate both positive and negative effects on the native biota. Federal management agencies have tended to focus on short-term issues arising from severe wildfires. Immediately after fire, burned area emergency response (BAER) teams work to identify areas of high erosion hazard and non-native plant invasion (among other things). Treatments implemented based on BAER reports can notably reduce soil loss and decrease exotic invasion of postfire landscapes (Robichaud et al. 2009). In the longer term, tree planting efforts can stabilize populations of rare species, accelerate successional processes, and begin to resequester carbon lost to fire. However, it is critically important that trees appropriate for the site are chosen for such projects. At the same time, postfire management can have negative effects on ecosystems. For example, timber harvest of dead trees from severely burned areas has become controversial because it often disturbs the soil and reduces the density of forest structures known to be important for a number of animal species (Peterson et al. 2009). In some cases such harvest may reduce the amount of fuel available for future combustion, but depending on the method used, fine fuels may actually increase. Many calls have been made for longer-term perspectives in postfire resource management, especially as fire activity and burned area increase in certain ecosystems (Robichaud et al. 2009, Cerdá and Robichaud 2009), but funds are usually lacking to effect such recommendations. A primary concern after fire is soil loss and excessive water flow off recently burned slopes that can lead to floods and debris flows. Approaches to address this problem seek to stabilize slopes to reduce erosion and runoff. One approach includes seeding of fast-growing grasses to stabilize slopes 40  Drivers

prior to winter rains. Another tactic is to distribute hay or other organic matter on burned slopes to attenuate the impact of torrential rainfall. Another approach, rather than to reduce erosion, is to construct barriers that capture debris and prevent runoff from entering drainages that lead to urban environments or other values at risk. Postfire aerial seeding as a management practice has its roots in southern California as a flood control measure. It arose partly from incomplete understanding of the natural capacity for rapid recovery in chaparral ecosystems. For example, one document (Los Angeles River Watershed 1941) stated: “Severe burning so depletes the chaparral cover that artificial measures are necessary to hasten its re-establishment.” However, we know from countless studies over the past fifty years that this is not true; chaparral possesses an extraordinary capacity for regeneration from resprouting of rootstocks and dormant seed banks. In addition, many studies over this time have shown that seeding both exotic and native species is a precarious undertaking that fails more often than not. There are several reasons why seeding is not practical on southern California landscapes, but the primary one is that these seeds require gentle and continuous autumn rains to establish root systems capable of holding soil back from winter rains. The first rains of the year commonly occur as intense torrents in late autumn and winter that wash seeds off the surface of steep slopes before they have an opportunity to establish. Such is not the fate of native seeds that are buried and better protected from transport by these rains. Perhaps a more important reason for not depending on seeding is that other methods for reducing slope erosion, such as mulch or hay bales, have proven to be more effective and far more predictable than seeding. Seeding also has the potential for negative impacts on the conservation of naturally functioning chaparral ecosystems. On those occasions where the rains do cooperate and exotic seeded species establish, they can outcompete native species and sometimes escape to become aggressive invaders. Black mustard (Brassica nigra), which was the favored exotic species used to seed after fires in southern California during the first half of the twentieth century, today is a widespread pest throughout the region. Physical barriers created by mulch and hay bales also can introduce exotic species; as a consequence, more and more such projects are requiring “weed-free” hay.

Global Changes Climate Change Current and projected future climatic trends appear likely to increase the potential for fire in most California wildlands. This reflects the interactions of various factors, including increasing temperatures, changing precipitation patterns, decreasing snowpack, higher probability of drought events, increasing forest fuels, greater fuel continuity (due to both fire suppression and faster growth under warmer climates), and increasing human ignitions (Miller and Urban 1999; Westerling et al. 2006; Lenihan et al. 2008; Flannigan et al. 2009; Miller et al. 2009; National Research Council 2011; Safford, North, and Meyer 2012). Projections based on dynamic vegetation models linked to down-scaled general circulation models (GCMs) suggest that the geographic distributions of major ecosystem types in California could change substantially by the end of the twenty-first century, with much of

the change mediated by changes in fire activity and severity (Hayhoe et al. 2004, Lenihan et al. 2008; see Chapter 14, “Climate Change Impacts”). Lenihan et al.’s (2008) projections for California’s Sierra Nevada and southern coast vary notably among climate scenarios, but some notable commonalities emerge. Under all scenarios the Sierra Nevada section is projected to experience a reduction in conifer forest (evergreen conifer and subalpine) and shrubland and an increase in hardwood-dominated forest (mixed forest and woodland) and grassland. In the south coast section all scenarios predict a loss of shrubland and an increase in grassland. In both sections these projected changes are largely due to increased fire activity, with grassland replacing shrubland and forest due to the inability of woody life forms to regenerate under greatly increased fire frequencies, and hardwoods replacing conifers due to the resprouting ability of hardwoods after fire and increased fitness of hardwoods under warmer conditions (especially where precipitation does not decrease substantially and if anthropogenic nitrogenous inputs are maintained or increased) (Lenihan et al. 2008). Notably, these projections extend trends already apparent in both ecological sections. For example, hardwood density has already increased in the Sierra Nevada section over the past seventy-five years (Bolsinger 1988, Dolanc et al. 2014), and grassland is already replacing shrubland across large areas of lowland southern California (Keeley 2006).

Restoration Ecological restoration typically requires identification of some reference state to provide a target for management and to allow measurement of restoration progress. Because anthropogenic alterations to ecosystems and the global environment have been so ubiquitous, contemporary, unaltered reference ecosystems are difficult to identify. Thus restorationists are often forced to rely on historical information to construct reference states (Safford, Hayward et al. 2012). Historical ecology has provided invaluable evidence of conditions before human degradation occurred, but directional changes in—​ among other things—​population, air and water pollution, and climate have led many to suggest that past conditions might no longer represent reasonable management targets for restoration planning (Millar et al. 2007, Wiens et al. 2012). Long-lived, woody structures in semiarid and arid environments can store centuries of fire occurrence information in the form of fire scars. This has allowed development of historically based management guidance for many western U.S. forest types, but rapid global change requires some reconsideration of management strategies that rely implicitly on an assumption of environmental stasis. Since most western U.S. forests will likely experience even more potential for fire than during the centuries before Euro-American settlement (the usual historical reference period, which coincided with the Little Ice Age), current focus on reintroducing fire and creating forest structures resistant or resilient to fire seems entirely justified. Our current efforts seem paltry when measured against the scale of changes taking place (North et al. 2012). The current focus on preserving “pristine” conditions (i.e., with high fidelity to historical environments) in many ecosystems will likely need to be amended to embrace both the inherent dynamism of ecological systems and the directional changes now well under way (Cole and Yung 2010; Safford, Hayward et al. 2012; Wiens et al. 2012).

Summary Fire is a natural ecosystem process throughout much of California. Historically, humans have played a substantial role in perturbing natural fire regimes. These impacts have differed substantially between montane, coniferous forests and lowerelevation, nonforested environments. Although humans today are responsible for over half of all fires ignited in forested landscapes, lightning has historically been an abundant source of ignition. Many California forests have had a long history of frequent, low-, and moderate-intensity surface fires, and the primary human impact has been suppression of the natural fire regime. These surface fires have been amenable to fire attack, rendering the history of fire suppression largely equivalent to fire exclusion. One consequence has been an anomalous accumulation of surface fuels and ingrowth of young trees, both of which have contributed to the potential for a fire regime shift to high-intensity crown fires. Shrublands and other nonforested landscapes in the state have historically burned in high-intensity crown fires; as a result, fire suppression activities have been unable to exclude fire as in conifer forests. These landscapes have historically low lightning frequency; since human occupation, people have been the dominant source of ignitions. With increasing population growth, fire suppression efforts have worked hard to keep up with increasing numbers of fires. However, twentieth-century fires have been more abundant in central and southern coastal California than historically was the case. This increase in fire frequency has had negative ecosystem impacts by type-converting native shrublands to non-native grasslands throughout many parts of the region. Future global changes are likely to have very different impacts on these two landscapes, with global warming playing a significant role in forests and demographic growth and urban development playing larger roles in coastal plains and foothills.

Acknowledgments This research was supported by the U.S. Geological Survey, Fire Risk Scenario project. Any use of trade names is for descriptive purposes only and does not imply endorsement by the U.S. government.

Recommended Reading Bond, W. J., and B. W. van Wilgen. 2005. Fire and plants. Chapman and Hall, New York, New York. Ford, R., Jr. 1991. Santa Barbara wildfires: Fire on the hills. McNally & Loftin, Santa Barbara, California. Keeley, J. E., W. J. Bond, R. A. Bradstock, J. G. Pausas, and W. Rundel. 2012. Fire in Mediterranean climate ecosystems: Ecology, evolution, and management. Cambridge University Press, Cambridge, UK. Kennedy, R. G. 2006. Wildfire and Americans: How to save lives, property, and your tax dollars. Hill and Wang, New York, New York. Sugihara, N. G., J. W. van Wagtendonk, K. E. Shaffer, J. FitesKaufman, and A. E. Thode. 2006. Fire in California’s ecosystems. University of California Press, Los Angeles, California. Wuerthner, G., editor. 2006. The wildfire fire reader: A century of failed forest policy. Foundation for Deep Ecology by arrangement with Island Press, Sausalito, California. Fire as an Ecosystem Process   41

Glossary

Rhizomes  Underground stems that spread laterally.

Burned area emergency response (BAER)  A federal program that evaluates and makes recommendations for postfire management.

Santa Ana winds  Known by meteorologists as foehn winds, they develop from a high-pressure system in the interior of western North America that drives high-velocity offshore winds towards a coastal low pressure cell.

Cation exchange capacity  Used as a measure of soil fertility it is the total level of nutrients a soil is capable of holding and available for exchange with the soil solution.

Surface fires  Fires that burn fuels near the ground, including herbaceous materials and dead leaves and other litter and are typically of low intensity.

Coarse-grained versus fine-grained patchiness  Coarsegrained environments are composed of large patches of a particular type, and fine-grained are small patches.

Taxa  A taxonomic unit such as species, genus, family, and so on.

Corms  Modified stems used for underground storage of carbohydrates, inorganic nutrients, and buds. Crown fires  Fires that burn the canopy of woody vegetation and are typically high intensity. Differenced normalized burn ratio (dNBR)  A remote sensing index that provides a measure of biomass before and after fire and is a measure of fire severity, or as is sometimes used, burn severity. Dry ravel  Also known as dry creep, describes the dry gravitational movement of soil particles, often driven forward by wetting/drying or freeze-thaw cycles and common especially in southern California. Endogenous regeneration  Plants that reproduce by seeds or resprouts from within the burn perimeter and do not depend on colonization from outside the burned area. Euro-Americans  Americans with ancestry in Europe. Facultative seeders  Plants that regenerate following fire by resprouting and from germination of dormant seed banks after canopies are killed by crown fire. Fire return interval  Time between fires for a spatially explicit area. Fire return interval departure (FRID)  The difference between current and presettlement fire frequencies. General circulation models (GCMs)  A type of mathematical climate of the general circulation of planetary atmosphere used to forecast future climates. Granivores  Animals that feed largely on seeds. Herbivores  Animals that feed largely on herbaceous material. Holocene  Along with the Pleistocene, one of the two epics of the Quaternary geological period; it is the current epoch that began ten thousand years ago. Lignotuber  Woody swelling that develops at the base of a stem, which has buds capable of initiating new shoots, typically after fire. Masting  A boom-and-bust pattern of seed production, where a population has synchronous seed production in one year followed by years of little or no seed production. Mediterranean-type climate  Moderate precipitation during winter when temperatures are mild alternating with hot summer droughts, typical of regions between 30–​38o latitude and on the western side of five continents. These include southern Europe and northern Africa, California, central Chile, the Cape of South Africa, western Australia, and south Australia. Metapopulation  Populations connected by colonization, which are important sources of recovery when a population is extirpated. Obligate resprouters  Plants that regenerate following fire only by resprouting after canopies are killed by crown fire. Obligate seeders  Plants are completely killed by crown fires and that regenerate from germination of formerly dormant seed banks. 42  Drivers

Tertiary  The geological period lasting from 65 million to 1.8 million years, including the epochs Paleocene, Eocene, Oligocene, Miocene, and Pliocene.

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Fire as an Ecosystem Process   43

seedlings. Pages 255–​273 in M. A. Leak, V. T. Parker, and R. L. Simpson, editors. Seedling ecology and evolution. Cambridge University Press, Cambridge, UK. Keeley, J. E., and P. H. Zedler. 1998. Evolution of life histories in Pinus. Pages 219–​251 in D. Richardson, editor. Ecology and biogeography of pines. Cambridge University Press, Cambridge, UK. Keeley, J. E., and P. H. Zedler. 2009. Large, high intensity fire events in southern California shrublands: Debunking the fine-grained age-patch model. Ecological Applications 19:69–​94. Lamb, M. P., M. Levina, R. A. DiBiase, and B. M. Fuller. 2013. Sediment storage by vegetation in steep bedrock landscapes: Theory, experiments, and implications for postfire sediment yield. Journal of Geophysical Research: Earth Surface 118:1147–​1160. Lenihan, J. H., D. Bachelet, R. P. Neilson, and R. Drapek. 2008. Response of vegetation distribution, ecosystem productivity, and fire to climate change scenarios for California. Climatic Change 87 (suppl. 1): S215–​S230. Lewis, H., 1973. Patterns of burning in California: Ecology and ethnohistory. Ballena Press, Menlo Park, California. Lightfoot K. G., R. Q. Cuthrell, C. M. Boone, R. Byrne, A. S. Chavez, L. Collins, A. Cowart, R. R. Evett, P. V. A. Fine, D. Gifford-Gonzalez, M. G. Hylkema, V. Lopez, T. M. Misiewicz, and R. E. B. Reid. 2013. Anthropogenic burning on the central California coast in late Holocene and early historical times: Findings, implications, and future directions. California Archaeology 5:371–​390. Littell, J. S., D. McKenzie, D. L. Peterson, and A. L. Westerling, 2009. Climate and wildfire area burned in western U.S. ecoprovinces, 1916–​2 003. Ecological Applications 19:1003–​1021. Los Angeles River Watershed. 1941. Survey report for the Los Angeles River watershed. U.S. Government Printing Office. 67 pp. Lyon, L. J., E. S. Telfer, and D. S. Schreiner. 2000. Direct effects of fire and animals responses. Pages 17–​23 in J. K. Smith, L. J. Lyon, M. H. Huff, R. G. Hooper, E. S. Telfer, and D. S. Schreiner, editors. Wildland fire in ecosystems: Effects of fire on fauna. General Technical Report RMRS-GTR-42, volume 1. USDA Forest Service, Rocky Mountain Research Station, Ogden, Utah. Mallek, C. R., H. D. Safford, J. H. Viers, and J. Miller. 2013. Modern departures in fire severity and area vary by forest type, Sierra Nevada and southern Cascades, California. Ecosphere 4(12): 1–28. Martinson, E. J., and P. N. Omi. 2013. Fuel treatments and fire severity: A meta-analysis. Research Paper RMRS-RP-103. USDA Forest Service, Rocky Mountain Research Station, Fort Collins, Colorado. Merriam, K. E., J. E. Keeley, and J. L. Beyers. 2006. Fuel breaks affect nonnative species abundance in California plant communities. Ecological Applications 16:515–​527. Miles, S. R., and C. B. Goudey. 1997. Ecological subregions of California: Section and subsection descriptions. Technical Paper R5-EMTP-005. USDA Forest Service, San Francisco, California. Millar, C. I., N. L. Stephenson, S. L. Stephens. 2007. Climate change and forests of the future: Managing in the face of uncertainty. Ecological Applications 17:2145–​2151. Miller, C., and D. L. Urban. 1999. Forest pattern, fire, and climatic change in the Sierra Nevada. Ecosystems 2:76–​87. Miller, J. D., H. D. Safford, M. Crimmins, and A. E. Thode. 2009. Quantitative evidence for increasing forest fire severity in the Sierra Nevada and southern Cascade Mountains, California and Nevada, USA. Ecosystems 12:16–​32. Miller, J. D., C. N. Skinner, H. D. Safford, E. E. Knapp, and C. M. Ramirez. 2012. Trends and causes of severity, size, and number of fires in northwestern California, USA. Ecological Applications 22:184–​2 03. Miller, J. D., and A. E. Thode. 2007. Quantifying burn severity in a heterogeneous landscape with a relative version of the delta Normalized Burn Ratio (dNBR). Remote Sensing of Environment 109:66– ​8 0. Minnich, R. A. 2006. California climate and fire weather. Pages 13–​ 37 in N. G. Sugihara, J. W. van Wagtendonk, K. E. Shaffer, J. FitesKaufman, and A. E. Thode, editors. Fire in California’s ecosystems. University of California Press, Los Angeles, California. Minshall, G. W. 2003. Responses of stream benthic macroinvertebrates to fire. Forest Ecology and Management 178:155–​161. Mitchell, S. R., M. E. Harmon, and K. E. O’Connell. 2009. Forest fuel reduction alters fire severity and long-term carbon storage in three Pacific Northwest ecosystems. Ecological Applications 19:643– ​655.

44  Drivers

Miyanishi, K. 2001. Duff consumption. Pages 437–​475 in E. A. Johnson and K. Miyanishi, editors. Forest fires: Behavior and ecological effects. Academic Press, San Diego, California. Moody, J. A., and D. A. Martin. 2001. Initial hydrologic and geomorphic response following a wildfire in the Colorado Front Range. Earth Surface Processes on Land 26:1049–​1070. National Research Council. 2011. Climate stabilization targets: Emissions, concentrations, and impacts over decades to millennia. National Academies Press, Washington, D.C. Nimmo, D. G., L. T. Kelly, L. M. Spence-Bailey, S. J. Watson, A. Haslem, J. G. White, M. F. Clarke, and A. F. Bennett. 2012. Predicting the century-long post-fire response of reptiles. Global Ecology and Biogeography 21:1062–​1073. North, M. P., B. M. Collins, and S. Stephens. 2012. Using fire to increase the scale, benefits, and future maintenance of fuels treatments. Journal of Forestry 110:392–​401. North, M. P., and M. D. Hurteau. 2011. High-severity wildfire effects on carbon stocks and emissions in fuels treated and untreated forest. Forest Ecology and Management 261:1115–​1120. Noss, R. F., J. F. Franklin, W. L. Baker, T. Schoennagel, and P. B. Moyle. 2006. Managing fire-prone forests in the western United States. Frontiers in Ecology and the Environment 4:481–​487. Oliver, A. A., J. E. Reuter, A. C. Heyvaert, and R. A. Dahlgren. 2012. Water quality response to the Angora fire, Lake Tahoe, California. Biogeochemistry 111:361–​376. Paleoindian migrations. Quaternary Science Reviews 30:269–​272. Pausas, J. G., and J. E. Keeley. 2009. A burning story: The role of fire in the history of life. BioScience 59:593–​601. Peterson, D. L., J. K. Agee, G. H. Aplet, D. P. Dykstra, R. T. Graham, J. F. Lehmkuhl, D. S. Pilloid, D. F. Potts, R. F. Powers, and J. D. Stuart. 2009. Effects of timber harvest following wildfire in western North America. General Technical Report PNW-GTR-776. USDA Forest Service, Pacific Northwest Research Station, Portland, Oregon. Pinter, N., S. Fiedel, and J. E. Keeley. 2011. Fire and vegetation shifts at the vanguard of Pyne, S. J. 1995. World fire. The culture of fire on earth. Henry Holt and Company, New York, New York. Raphael, M. G., M. L. Morrison, and M. P. Yoder-Williams. 1987. Breeding bird populations during twenty-five years of postfire succession in the Sierra Nevada. Condor 89:614–​626. Raven, P. H., and D. I. Axelrod. 1972. Plate tectonics and Australasian paleobiogeography. Science 176:1379–​1386. Robichaud, P. R., S. A. Lewis, R. E. Brown, and L. E. Ashmun. 2009. Emergency post-fire rehabilitation treatment effects on burned area ecology and long-term restoration. Fire Ecology 5(1):115–​128. Robock, A. 1988. Enhancement of surface cooling due to forest fire smoke. Science 242:911–​913. Rochester, C. J., C. S. Brehme, D. R. Clark, D. C. Stokes, S. A. Hathaway, and R. N. Fisher. 2010. Reptile and amphibian responses to large-scale wildfires in southern California. Journal of Herpetology 44:333– ​351. Safford, H. D. 2007. Man and fire in southern California: Doing the math. Fremontia 35(4):25–​29. Safford, H. D., G. Hayward, N. Heller, and J. A. Wiens. 2012. Climate change and historical ecology: Can the past still inform the future? Pages 46–​62 in J. A. Wiens, G. Hayward, H. D. Safford, and C. M. Giffen, editors. Historical environmental variation in conservation and natural resource management. John Wiley and Sons, New York, New York. Safford, H. D., J. Miller, D. Schmidt, B. Roath, and A. Parsons. 2008. BAER soil burn severity maps do not measure fire effects to vegetation: A comment on Odion and Hanson (2006). Ecosystems 11:1–​11. Safford, H. D., M. North, and M. D. Meyer. 2012. Climate change and the relevance of historical forest conditions. Pages 23–​46 in M. P. North, editor. Managing Sierra Nevada forests. General Technical Report PSW-GTR-237. USDA Forest Service, Pacific Southwest Research Station, Albany, California. Safford, H. D., J. T. Stevens, K. Merriam, M. D. Meyer, and A. M. Latimer. 2012. Fuel treatment effectiveness in California yellow pine and mixed conifer forests. Forest Ecology and Management 274:17–​2 8. Safford, H. D., and K. M. Van de Water. 2014. Using fire return interval departure (FRID) analysis to map spatial and temporal changes

in fire frequency on National Forest lands in California. Research Paper PSW-RP-266. USDA Forest Service, Pacific Southwest Research Station, Albany, California. Sandoval, C. 2000. Persistence of a walking-stick population (Phasmatoptera: Timematodea) after a wildfire. Southwestern Naturalist 45:123–​127. Sauvajot, R. M., M. Buechner, D. A. Kamradt, and C. M. Schonewald. 1998. Patterns of human disturbance and response by small mammals and birds in chaparral near urban development. Urban Ecosystems 2:279–​297. Sawyer, J. O., T. Keeler-Wolf, and J. M. Evens. 2009. A manual of California vegetation. Second edition. California Native Plant Society, Sacramento, California. Schuette, P., J. Diffendorfer, D. Deutschman, S. Tremor, and W. Spencer. 2014. Carnivore distributions across chaparral habitats exposed to wildfire and rural housing in southern California. International Journal of Wildland Fire 23 (4), 591–600. Schuur, E. A., S. E. Trumbore, M. C. Mack, and J. W. Harden. 2003. Isotopic composition of carbon dioxide from a boreal forest fire: Inferring carbon loss from measurements and modeling. Global Biogeochemical Cycles 17(1):1–​9. Schwilk, D. W., J. E. Keeley, E. E. Knapp, J. McIver, J. D. Bailey, C. J. Fettig, C. E. Fiedler, R. J. Harrod, J. J. Moghaddas, K. W. Outcalt, C. N. Skinner, S. L. Stephens, T. A. Waldrop, D. A. Yassey, and A. Youngblood. 2009. The National Fire and Fire Surrogate Study: Effects of alternative fuel reduction methods on forest vegetation structure and fuels. Ecological Applications 19:285–​304. Shaffer, K. E., and W. F. Laudenslayer. 2006. Fire and animal interactions. Pages 118–​144 in N. G. Sugihara, J. W. van Wagtendonk, K. E. Shaffer, J. Fites-Kaufman, and A. E. Thode, editors. Fire in California’s ecosystems. University of California Press, Los Angeles, California. Skinner, C. N., J. H. Burk, M. G. Barbour, E. Franco-Vizcaíno, and S. L. Stephens. 2008. Influences of climate on fire regimes in montane forests of north-western Mexico. Journal of Biogeography 35:1436–​1451. Skinner, C. N., A. H. Taylor, and J. K. Agee. 2006. Klamath Mountain Bioregion. Pages 170–​194 in N. G. Sugihara, J. W. van Wagtendonk, K. E. Shaffer, J. Fites-Kaufman, and A. E. Thode, editors. Fire in California’s ecosystems. University of California Press, Los Angeles, California. Smith, J. K., L. J. Lyon, M. H. Huff, R. G. Hooper, E. S. Telfer, and D. S. Schreiner, editors. 2000. Wildland fire in ecosystems: Effects of fire on fauna. General Technical Report RMRS-GTR-42. USDA Forest Service, Rocky Mountain Research Station, Fort Collins, Colorado. Smucker, K. M., R. L. Hutto, and B. M. Steele. 2005. Changes in bird abundance after wildfire: Importance of fire severity and time since fire. Ecological Applications 15:1535–​1549. Stephens, S. L., R. E. J. Boerner, J. J. Moghaddas, E. E. Y. Moghaddas, C. Edminster, C. E. Fiedler, B. R. Hartsough, J. E. Keeley, E. E. Knapp, J. D. McIver, C. N. Skinner, and A. Youngblood. 2012. Fuel treatment impacts on estimated wildfire carbon loss from forests in Montana, Oregon, California, and Arizona. Ecosphere 3(5): article 38. . Stephens, S. L., and B. M. Collins. 2004. Fire regimes of mixed conifer forests in the north-central Sierra Nevada at multiple spatial scales. Northwest Science 78:12–​23. Stephens, S. L., J. D. McIver, R. E. J. Boerner, C. J. Fettig, J. B. Fontaine, B. R. Hartshough, P. L. Kennedy, and D. W. Schwilk. 2012. The effects of forest fuel-reduction treatments in the United States. BioScience 62:549–​560. Stoof, C. R., D. Moore, P. M. Fernandes, J. J. Stoorvogel, R. E. S. Fernandes, A. J. D. Ferreira, and C. J. Ritsema. 2013. Hot fire, cool soil. Geophysical Research Letters 40:1–​6. . Sudworth, G. B. 1900. Stanislaus and Lake Tahoe Forest Reserves, California, and adjacent territories. Pages 505–​561 in Twenty-

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FOUR

Geomorphology and Soils ROBERT C. GR AHAM and A . TOBY O’GEEN

Overview of Geologic Processes California’s highly diverse geologic setting is the result of the interplay of tectonics, sediment accumulation, emplacement and alteration of oceanic rocks, intrusion of magmas, extrusion of volcanic rocks, glaciation, and lateral displacement of rocks along faults. These conditions—​coupled with dramatic differences in climate, biota, topography, and landscape age—​g ive rise to a diversity of soils. Nevertheless, the state can be divided into distinct physiographic provinces (Figure 4.1) to discuss the relationships between geology and climate that have interacted to produce distinct soils and ecosystems. The majority of the chapter is organized by physiographic province, beginning with the Sierra Nevada, moving west to the Central Valley and Coast Ranges, then north to the Klamath and Cascade Mountains and the Modoc Plateau, then down and around through the desert Basin and Range, Transverse and Peninsular Ranges, and back up California’s coast. Although we present an overview of the kinds of soils present in the state, our emphasis is on soil processes and interactions with ecosystems. California’s landscapes and lithologies are inextricably related to the convergence of two tectonic plates: the Pacific plate pushes and scrapes northwest along the North Ameri-

can plate at a rate of about five centimeters per year. The San Andreas fault along California’s western edge expresses the boundary between the two plates, but numerous other faults result from the tectonic forces. Lateral and vertical movements along the faults create valleys, uplift mountain ranges, and shape stream courses. Erosion of continually uplifted terrain has kept mountain landscapes rugged and unstable and has produced huge amounts of sediment that fill valleys (Harden 2004). The types of rocks that form soils to support California’s ecosystems owe their formation, or at least their exposure, to tectonic movement. As the Pacific plate has been subducted beneath the North American plate in the area of northern California, oceanic sediments have been accreted against the plate margin. In some places slabs of ultramafic mantle rock are lodged into these sediments. Both the sediments and ultramafic rocks have been uplifted and exposed—​in the northern Coast Ranges, the western Sierra Nevada, and the Klamath Mountains. These rocks have been metamorphosed by the high pressures and temperatures caused by deep burial. As the Pacific plate is further subducted, its rocks melt and magma rises up through the North American plate causing 47

Oregon

Oregon

Idaho

Nevada

Nevada

Utah

Nevada

N

ev ad a

Arizon

a

Arizona Nevada Mexico

FIGURE 4.1 Physiographic regions of California. Map: Parker Welch (UC Santa Cruz).

volcanic extrusions, mostly in the eastern part of the state. Granites in the Sierra Nevada and elsewhere in the state are also the result of melted subducted Pacific plate rocks that intruded upward into the North American plate but cooled before reaching the surface, and that faulting later uplifted and erosion exposed (Harden 2004). Sediments have accumulated in the deep trough of the Great Valley as they have been shed off the surrounding mountains. More recently, the higher elevations of the Klamath Mountains, the Cascade Range, and especially the Sierra Nevada were shaped by glaciation that exposed fresh bedrock and deposited moraines and glacial outwash material at lower elevations.

Photo on previous page: Coastal landscape at Cambria showing two elevations of marine terraces: a lower one with grass and a higher one with trees on the skyline. The soil on the lower terrace (center of photo) has a thick, dark horizon. Photo: Tom Rice. 48  Drivers

Sierra Nevada Geology and Geomorphology The Sierra Nevada occupies approximately 70,000 square kilometers, extending from its southern end east of Bakersfield to just south of Lassen Peak in the north. The mountain range is asymmetric, consisting of a gradually sloping western face created by a west-tilted fault block and a steep eastern escarpment, which represents the normal faulting that delineates the western edge of the Basin and Range province. Elevation at the crest is highest in the south due to greater uplift, with peaks over 4,000 meters, and decreases to between 1,800 and 2,450 meters at the northern end. The topographic and altitudinal trend of the southern Sierra is more complex compared to the rest of the west slope, with a westward-stepped topography that dips toward the south from the latitude of Mount Whitney (Clark et al. 2005).

The southern Sierra and much of the high country across the mountain range is an exposed granitic batholith. The Sierran Batholith, emplaced 210 million to 80 million years ago, consists of granitic rocks, interspersed with small plutons of more mafic rocks. Metamorphic rocks are also found in the Sierra as roof pendants, which are remnants of the rock that once covered the batholith. The metamorphic cap has been removed by erosion at high elevations and throughout the southern Sierra but remains in parts of the central and northern Sierra and throughout the northern half of the foothill region where uplift has been less. Uplift and faulting was accompanied by volcanic activity. Volcanism was common in the northern Sierra between twenty million and five million years ago. Volcanic activity that originated east of what is now the Sierra crest extruded massive basaltic lava flows and andesitic mudflows that followed ancient drainage systems all the way to the Sacramento Valley floor. The remnants of these deposits can be seen as table mountains throughout the region. This inverted topography was formed where ancient river valleys were filled with volcanic flows and the surrounding older terrain was removed by erosion. Topographic trends are governed by lithology and the general asymmetry of the mountain range. The east slope landscapes are characterized by very steep mountains drained by steep V-shaped canyons at elevations below where glaciation has occurred. Slopes become moderate to gently sloping near the base of the east slope, where glacial till, outwash, and Quaternary alluvial fan deposits exist. At high elevations (1,830 to 3,050 meters) within the exposed batholith, terrain consists of gently to moderately sloping plateaus, steep mountains, and U-shaped valleys sculpted by glaciers. At lower elevations (920 to 1,830 meters) of the west slope, granitic terrain gives rise to steep slopes drained by very steep-sided canyons, particularly in the southern Sierra. The metamorphic terrain of the west slope and foothill region consists of a moderately steep plateau dissected by major rivers. The granitic foothill landscapes are typically more hilly, with rounded summits, convex and concave hillslopes, and a network of small interconnected lowlands.

Soil Formation Influences Along the west slope of the Sierra, the orographic effect of air masses cooling with increasing elevation creates a strong climatic gradient that produces systematic patterns in soil development. The mean annual precipitation (MAP) in the foothills ranges from 200 millimeters in the south to 1,020 millimeters in the central and northern foothills. As elevation increases, MAP increases to a maximum of 2,290 millimeters at the highest northern and central peaks. MAP decreases dramatically on the steep east side, ranging from 200 to 1,270 millimeters, depending on elevation. Precipitation generally increases in a northerly direction. Mean annual air temperatures decrease from 18°C in the foothill region to 4°C at the crest. A bioclimatic gradient exists across the Sierra. Vegetation progresses from blue oak and live oak woodlands at low elevations through ponderosa pine, mixed conifer, and red fir forests at mid-elevations and alpine lodgepole pine systems at highest elevations, as described in Chapter 12, “Vegetation,” and Chapter 27, “Montane Forests.” Soil development trends correspond to this altitudinal gradient, which imposes a weathering environment that is limited by moisture at low elevations and by low temperature at high elevations. A zone

of high-weathering intensity exists along the entire Sierra Nevada at mid-elevations (1,000 to 2,500 meters). This belt of intense soil development occurs in all parent materials and reflects the combined influence of mild temperatures and high precipitation, predominantly as rain (Dahlgren et al. 1997, Rasmussen et al. 2007). At high elevation (above approximately 3,000 meters), landscape age is generally constrained to the end of the last glacial maximum. Here soil cover and weathered bedrock were scoured by advancing ice sheets, resetting the pedologic clock at approximately ten thousand years before present. It is likely that soils formed at lower elevations (e.g., below approximately 3,000 meters) are significantly older. Moreover, soils formed from glacial material such as outwash and till can have a wide range of ages as a result of multiple glaciations. Soils formed on older glacial deposits (older than eighty thousand years) contain reddish Bt horizons (Birkeland and Janda 1971). In some of these older soils, subsoil horizons are cemented by pedogenic silica. Soils formed in till deposited fourteen thousand to twenty-one thousand years ago are less developed, with only weak B horizons. Soils younger than this are only slightly weathered and have colors that are close to those of the parent material (olive brown and yellowish brown) (Birkeland and Janda 1971).

Soils of Granitic Terrain Elevation-controlled climate strongly influences chemical and physical properties of soils on granitic terrain in the Sierra Nevada (Dahlgren et al. 1997). For example, soil pH decreases approximately two units from around neutral (pH = 6.5–​7.5) in the foothills to strongly acidic (pH = 4.7–​5.5) at elevations greater than 2,000 meters, where base cations are removed by greater leaching. Carbon:nitrogen ratios in the upper 18 centimeters of soil increase with elevation from a low of 12.4 in the oak woodlands of the foothill region to a maximum of 32.0 at elevations above 2,000 meters, where coniferous plant residues are more resistant to microbial decomposition due to lower N levels and higher lignin and polyphenol content. Soil organic carbon (SOC) content summed for the A and B horizons, shows a bell-shaped relationship with elevation (Figure 4.2). Relatively warm soil temperatures throughout the year in the low foothills accelerate microbial decomposition of plant residues, resulting in low SOC (~5 kg m-2). SOC is highest at mid-elevations (~14 kg m-2) due to the tremendous biological productivity there. SOC decreases at upper elevations (~10.5 kg m-2 ) because cold temperatures limit primary productivity. Soil depth, clay content, and pedogenic iron oxides follow a trend similar to that of SOC, with highest levels at mid-elevations. This is due to the intense weathering facilitated by high precipitation and mild temperatures in that elevation range (Dahlgren et al. 1997). Soil morphologies reflect differences in soil forming processes. At high elevations (higher than 2,500 meters), cold temperatures limit pedogenesis to the extent that litter accumulation and darkening of the soil profile by soil organic carbon are the dominant processes. Soils are coarse textured (sands and loamy sands) and often have an abrupt transition to hard bedrock in areas where glaciation occurred. These soils are Entisols. At elevations between 1,800 and 2,500 meters, Bw horizons have slight illuvial clay accumulation, more intense colors, and development of soil structure. At somewhat lower elevations, soils are deep and more highly Geomorphology and Soils   49

dependence on elevation. At mid-elevations soil textures are typically loams and sandy loams in surface horizons, gradually transitioning to clay loams and clays in the subsoil. Clay mineralogy of these soils is dominated by kaolin and amorphous iron and aluminum oxides. Many soils in volcanic terrain of the Sierra are Ultisols. Alfisols are common where precipitation is lower and leaching is less. At upper elevations Andisols and Inceptisols are prevalent (Rasmussen et al. 2007). Soils with andic properties are found in winter-snow–​ dominated landscapes between 1,700 and 2,150 meters. Mineral transformations are minimal at higher elevations.

Solum organic C (kg m-2)

20 16 12

8

4 0

Soils of the Foothills 0

500

1000

1500

2000

2500

3000

Elevation (m) FIGURE 4.2 Organic carbon content in the solum (A + B horizons) of soils in granitic terrain along an elevation transect in the Sierra Nevada. Black dots represent data generated by Dahlgren et al. (1997), open circles are from soil survey characterization data, and the line is generated from the means at the different elevations. Source: Adapted from Dahlgren et al. 1997.

weathered and are underlain by thick zones of ­weathered bedrock. Soil textures are commonly sandy loams in surface horizons, abruptly transitioning to sandy clay loams in Bt horizons. High concentrations of pedogenic iron give rise to well-aggregated soils, which promote rapid infiltration and drainage. A fundamental role of soil is to regulate water and nutrients to the ecosystem. Deeply weathered, crumbly granite bedrock is a characteristic of Sierran landscapes that contributes to the fertility and productivity of this ecosystem. This weathered bedrock is an important source of plant-available water (Graham, Rossi, and Hubbert 2010) and nutrients such as P, Ca, Mg, and K that would otherwise be unavailable. The stored water that weathered bedrock supplies during the dry season governs the productivity and survival of coniferous forests (Arkley 1981, Witty et al. 2003, Bales et al. 2011). Weathered bedrock is a more important source of water for plant transpiration during the summer than are overlying soils (Anderson et al. 1995; Hubbert, Beyers, and Graham 2001; Hubbert, Graham, and Anderson 2001; Rose et al. 2003).

Soils of Volcanic Terrain Many soils in the region are formed from volcanic materials and have what are known as andic properties. Some of these properties, such as low bulk density, high plant-available water retention, and high soil organic carbon contents, promote forest productivity (Shoji et al. 1993, Rasmussen et al. 2007). But these soils also require careful consideration relative to some of their characteristics, including pH-dependent charge, low cation exchange capacity, phosphate fixation, and high susceptibility to erosion (Dahlgren et al. 2004). Compared to soils derived from granite and many other parent materials, soils formed from volcanic materials tend to be more developed due to the abundance of highly weatherable components such as glass, pyroxenes, biotite, and amphiboles. Nevertheless, soils derived from andesite, a common volcanic rock in the Sierra, show weathering trends similar to those outlined for granite, with the same strong 50  Drivers

The oak-grassland foothill region (Figure 4.3) has a northsouth trending sequence of parent material lithology. The northern foothill region consists mainly of metavolcanic and volcanic rocks (greenstone, basalt, and andesite). The central foothill region consists of a complex mixture of rock types including metasedimentary, metavolcanic, sedimentary, volcanic, and intrusive igneous. The southern foothill region is largely granitic with some volcanic rocks interspersed. Soils developed from marble, serpentinite, andesite, and metavolcanic rocks are most developed. These soils typically have subsurface clay contents that exceed 30% and strong red colors throughout much of the profile. Soils derived from greenstone (metavolcanic) often have a claypan (see Figure 4.3). These horizons create seasonal perched water tables that, in sloping terrain, result in subsurface lateral flow that supplies a significant component of streamflow in ephemeral streams (O’Geen et al. 2010, Swarowsky et al. 2012). Foothill soils derived from metasedimentary and granitic rocks tend to show less morphologic development. They are less red, indicating lower pedogenic iron content, and their clay contents tend to gradually increase with depth, typically to a maximum of 30%. With the exception of soils derived from granite, most foothill soils contain abundant rock fragments, with median values exceeding 20%, particularly in the subsoil. Soil depth is perhaps the most important soil property governing the quantity of water and nutrients in semiarid landscapes. Soils derived from granite are most deeply weathered, often exceeding 160 centimeters in thickness. Soils derived from marble and metasedimentary rocks also tend to be very deep. Soils derived from metavolcanic rock tend to reach 1 meter in thickness. Soils derived from serpentinite and slate are commonly less than 1 meter deep and in many instances are shallow (less than 50 centimeters deep).

Soils of Montane Valleys A little less than 10% of the Sierra Nevada consists of upper montane meadows (Ratliff 1985). These meadowlands perform many important ecosystem services such as flood water retention, maintenance of summer streamflows, forage production for grazing animals, carbon sinks, and hotspots of biodiversity (Norton et al. 2011). Soil properties of the meadows are influenced by water table dynamics. In meadows that are saturated with water during most of the growing season, the limited supply of oxygen slows microbial decomposition of plant residues resulting in a buildup of organic residues. The soils are Histosols. Drier meadows have mineral soils, typically Mollisols. Meadow landscapes are vulnera-

FIGURE 4.3 The oak grasslands in the foothills of the Sierra Nevada (above) have soils with claypans (very clayey Bt horizons) when formed from metavolcanic rock. Here, the claypan is the bottom zone where the knife is propped up. Note the vertical cracks in the dry clay.

ble to encroachment by surrounding coniferous forest if they dry out. Meadow drying can occur due to long-term climatic changes or by channel incision that lowers the groundwater table of the entire system. There are a variety of other types of valley landscape positions in the Sierra. Rapid uplift has resulted in steeply incised canyons with narrow valley floors in settings where glaciation has not occurred. These soils are highly variable in terms of chemical and physical composition because they are a product of the depositional environment. For example, landforms that experience high-energy flood events have coarse-textured soils with high rock fragment content and areas that experience low-energy flooding have fine-textured soils and may be rich in organic matter because imperfect drainage impedes its decomposition. Many soils in these landscapes are stratified with abrupt transitions between horizons that have very different textures. In the central and northern foothills many valley soils have formed in tailings and debris deposits produced by placer mining operations during the Gold Rush. These landscapes can have a wide range of soil properties but typically consist of stratified deposits of stony, cobbly, and gravelly material with enough fine sand or silt to

support annual grasses. Soil thickness can be highly variable in these landscapes.

The Central Valley Geology and Geomorphology The Central Valley is one of the most productive regions in the world. The valley is approximately 58,000 square kilometers stretching around 650 kilometers in length. It is often considered as two regions separated by the San Joaquin Delta, the Sacramento Valley to the north and the San Joaquin Valley to the south. The Sacramento and San Joaquin river systems converge at the delta, which is at sea level. The valley gently slopes to an elevation of 120 meters at its southern end and approximately 245 meters at its northernmost reach. The region consists of an elongate asymmetric syncline, once a large inland sea that began filling with sediment approximately 150 million years ago. The ancient sediment has been estimated to be as much as 13 to 15 kilometers thick and is capped by Pleistocene alluvium (Alt and Hyndman Geomorphology and Soils   51

2000). Gently sloping, nearly level alluvial fans exist across the valley, derived from granitic, metamorphic, sedimentary, and volcanic rock sources. Fans created by the Mokelumne, Stanislaus, Tuolumne, Merced, San Joaquin, Kings, Kaweah, and Cosumnes Rivers have drainage basins that connect to glaciated portions of the Sierra Nevada, whereas the fans created by the Calaveras, Chowchilla, and Fresno Rivers, and minor fans produced by many smaller rivers, have drainage basins that are limited to nonglaciated regions (Weissmann et al. 2005). Gently sloping floodplains and terraces are found along streams and rivers. Both gently and steeply undulating Pleistocene terraces occupy the valley margins. These terraces are remnants of fans that document depositional and erosional response to Quaternary climate change and regional uplift (Weissmann et al. 2005). During the initial stages of Sierran uplift, the mountains were capped with metamorphic rock. As uplift continued, the mountain caps were removed by erosion, exposing the underlying granitic batholith. Thus a great deal of the early alluvium from metamorphic rock was buried by subsequent alluvium from granitic sources. However, uplift and erosion has exposed older portions of the alluvial fan by stripping more recent alluvial cover at valley margins (O’Geen, Pettygrove et al. 2008). This alluvial sequence grades entirely into granitic alluvium in the southern part of the valley. The agriculturally rich landscapes of the Central Valley contain an incredibly diverse array of soils, perhaps more so than any other agricultural area of similar size in the United States. This soil variability reflects the influence of multiple parent materials spanning a range of ages, from modern-day stream deposits to ancient geomorphic surfaces that are among the oldest in the country. Soils of the Central Valley have formed from thick alluvium that has accumulated over thousands to millions of years. Most of the sediment has come from the Sierra because of its large rivers and the vast volume of sediment supplied by glacial outwash. The absence of glaciation and the smaller watersheds of the Coast Range have limited its sediment deposition to the western margin of the Valley. Within the basin floor, meandering rivers have mixed Coast Range and Sierran alluvium. Terraces and alluvial fans exist as a sequence of buried alluvial deposits, where younger alluvium overlies older deposits. Older alluvium remains exposed at the valley margins, isolated from younger deposition by uplift of the Sierra and Coast Range. These old terraces have soils with very different properties compared to recent alluvium.

Soils of the Basin and East Side On the east side of the Central Valley, soils have formed from alluvium of mixed sources that has accumulated over thousands of years from erosion of the Sierra Nevada. Systematic spatial patterns in soil exist because the parent materials have been deposited by water over time (Figure 4.4). A typical sequence from the basin floor (center of the valley) east to the foothill region includes four alluvial soil landscapes. 1. Weakly developed soils in basin alluvium and recent floodplain deposits. These formed between the present and about ten thousand years ago. Floodplains along major river systems contain sandy soils with little pedogenic alteration. A typical soil profile of this landscape position is layered, reflecting the energetics of the deposi52  Drivers

tional environment, where high-energy flood events deposit coarse-grained materials and low-energy events deposit fine-grained material. Since deposition is often event-based, soil profiles commonly have large and abrupt changes in texture with depth. Much of the basin experienced moderate to low energy sediment deposition, resulting in fine-textured soils with poor drainage conditions and smectite clays, many of them Vertisols, which are difficult to manage. Their change in volume in response to moisture status destroys roads, foundations, structures, and sidewalks. These soils also damage tree roots, making them more susceptible to disease. Vertisols are commonly flooded for rice production because they are slowly permeable when wet. Many basin soils are poorly drained due to a seasonally high water table that extends to the soil surface or within the root zone. The poor drainage slows organic matter decomposition, resulting in soils with high soil organic matter content (Mollisols). 2. Weakly developed soils on basin margins and broad alluvial fans. These were emplaced from approximately ten thousand to seventy thousand years ago. Large expanses of this landscape exist, derived from rivers (Mokelumne, Cosumnes, Stanislaus, Tuolumne, Kern, and Kings Rivers) that deposited glacial sediment as massive alluvial fans in the valley floor. Soils on these fans are typically coarse-textured and show little evidence of soil formation; the main process being accumulation of organic carbon in an A horizon. As a result, soils are typically Entisols and Mollisols. 3. Highly developed soils on low terraces. These deposited approximately 130,000 to 330,000 years ago. Large expanses of this soilscape rise above more recent alluvial fan deposits. Much of the landscape has microtopographic highs and lows (patterned ground) and the soils, formed in granitic alluvium, often contain duripans and claypans that restrict water and root penetration. The dominant soil of this region is the San Joaquin series, California’s state soil. 4. Highly developed soils on dissected high terraces (deposited more than 600,000 years ago) on the eastern edge of the valley. The landscape consists of alluvial fans that have been dissected by erosion into remnant “islands” (high terraces) with ancient soils. The soils have duripans and/or claypans and abundant resistant gravels and cobbles composed primarily of quartzite and chert. Clay content often exceeds 40% in the Bt horizons. The pH decreases with depth from moderately acidic in the A horizon to strongly acid in horizons below. The clay mineralogy is dominantly kaolinitic, indicating that these soils are highly weathered (O’Geen, Pettygrove et al. 2008). In places, erosion and terrace dissection have removed much of the upper parts of these very old soils. In the past, soils with restrictive horizons, such as duripans and claypans, limited the agricultural possibilities to annual crops. As prices of fruit, nuts, and grapes increased, these soils were developed for production through the destruction of restrictive horizons by deep tillage and mixing. Currently, soils with restrictive horizons are rare in the San Joaquin Valley to the extent that they have become endangered (Amundson et al. 2003). Land leveling is another common land management practice that has altered the Central Valley

0

20

50

100 km

Central Valley Landforms Weakly developed soils in basin alluvium and recent floodplains Weakly developed soils on basin margins and recent fan deposits Weakly developed soils on semi-consolidated volcanic lahar Highly developed soils on low terraces Highly developed soils on high terraces Weakly developed soils derived from bedrock of the foothills Miscellaneous landforms FIGURE 4.4 the spatial extents of soil landscapes of the central Valley. Map developed by V. Bullard, M. Walkinshaw, t. Harter, and A. t. o’Geen using the u.S. department of Agriculture–​natural resources conservation Service SSurGo (Soil Survey Geographic) database.

landscapes for irrigation purposes. These agricultural practices, coupled with urban expansion, mean that soils in the Central Valley have been, and continue to be, subject to significant changes.

Soils of the West Side The west side of the valley displays a sequence of soils that increases in age and soil development up fan (see Figure 4.4). Alluvium from the Coast Range tends to be finer-textured compared to east side alluvium, resulting in slower drainage. Geologic conditions give rise to an extensive area of saltaffected soils, particularly in the San Joaquin Valley. The Coast Ranges consist of oceanic sediments that are high in sodium and other soluble salts. The flushing of salts via deep percolation is impeded by a slowly permeable clay lens, creating a regionally high groundwater table. Coast Range parent materials also weather to relatively fine-textured soils, which drain slowly, so soil water is held close to the surface, where it is subject to evaporation and salts are left behind. Saline, sodic, and saline-sodic soils are present under these conditions, especially at the interface of the eastern edge of alluvial fans and the basin margin. Saline soils have soluble salt concentrations with electrical conductivity (EC) values that exceed 4 ds m-1, a level that many standard crops cannot tolerate. Sodic soils have low EC but high exchangeable sodium percentage (ESP), >15% on cation exchange sites. Sodicity disperses clays, thereby clogging pores and drastically decreasing permeability. The pH of sodic soils is also extremely high, often greater than 9. Saline-sodic soils have EC above 4 ds m-1 and ESP above 15%. These soils do not have restricted permeability, but do cause osmotic stress in plants. In the past, saline soils were reclaimed by leaching, while sodic and saline-sodic soils were treated with gypsum and then leached. This practice has been largely discontinued because there is no safe place to dispose of the leachate, which can be high in selenium. Many innovative practices are now employed to manage salt-affected soils on the west side. One example involves segregating land for high-value salt-­intolerant crops and reusing drain waters on adjacent fields for salt-tolerant crops.

Soils of the Delta The delta is a triangular-shaped inland marsh. The landscape is not a true delta in that it has expanded in an eastern direction rather than westward toward the ocean (Shlemon and Begg 1975). This is because converging Sacramento and San Joaquin Rivers encounter a topographic constriction caused by the Montezuma Hills and the Coast Ranges on the downstream side. Water is impounded in an eastern direction and constrained by alluvial fans at its eastern edge. The California Delta Region occupies almost 300,000 hectares, much of which consists of peat (Histosols). Peat deposits were formed during the Holocene as melted continental glaciers created sea level rise, which flooded inland channels within the region. The growth of dense wetland vegetation, primarily tules, provided the organic residues that have aggraded through time because microbial decomposition is inhibited by the water-saturated conditions. Peat accumulations reach depths of 9 to 15 meters in the thickest part and thin to the east and along the delta margins. The delta does not consist 54  Drivers

solely of peat. The network of sloughs, rivers, and tributaries creates a series of natural levee deposits. These formed before human modifications to the delta landscape through the deposition of coarse-textured sediment during flood events along floodplains of stream/river channels. This process created microtopographic highs that grade down to near sea level with increasing distance from the channel. Most levee deposits have been leveled for agriculture or during anthropogenic levee construction, but the sand patches remain interstratified with poorly drained thick peat deposits. Before the arrival of settlers of European descent, the delta was one of California’s most dynamic landscapes, subject to change from daily tides, annual floods, droughts, shifts in climate, and sea level rise. A network of anthropogenic levees and drainage systems were completed by the late 1930s that “reclaimed” approximately 3,000 square kilometers for agricultural use (Ingebritsen et al. 2000). Drainage of peat soils has resulted in land subsidence due to microbial oxidation and compaction by dewatering. A significant portion of the delta has subsided as much as 15 meters, reducing its elevation in places to 3 meters below sea level. Over one billion cubic yards of soil are projected to be lost by 2050 from microbial oxidation alone. As a result, the delta has become a very unstable manufactured landscape where levees are prone to failure due to earthquakes and flooding. The subsidence-induced susceptibly to levee failures threatens not only local communities and agriculture, but the drinking water for more than twenty-five million people as well. A levee breach in the wrong location could cause a saltwater intrusion that would destroy water quality.

Vernal Pool Soils California vernal pools are seasonal, freshwater wetlands commonly formed by perched water above a low-permeability layer, such as a duripan, claypan, or bedrock. Vernal pools are found in low areas of gently undulating topography. Vernal pool landscapes consist of integrated and unintegrated drainages composed of a series of basins and mounds. Seasonal pools in the basins are typically small (tens of square meters), with maximum water depth seldom greater than 50 centimeters. They rarely remain saturated for more than sixty to ninety consecutive days. Vernal pools experience hydrologic extremes in California’s Mediterranean climate, where they become inundated with standing water, experience episodes of saturated conditions during winter and spring, then become completely desiccated during summer and fall. In many instances seasonally submerged soils of the basins meet hydric soil criteria (and jurisdictional wetland definitions) but may represent just 3% to 5% of the vernal pool landscape area. While the surrounding microtopographic highs (mounds) do not meet current hydric soil criteria (O’Geen et al. 2007), subsurface hydraulic connectivity links water and nutrient transport in the soils across vernal pool landscapes (Rains et al. 2006). Disturbance of up-gradient vernal pool landscapes may have appreciable impacts on hydrological and biogeochemical processes in all down-gradient vernal pools. Thus it is important that these landscapes are not unnaturally divided into well-drained (mound) and poorly drained (basin) areas for land-use purposes. Due to complex spatial and temporal patterns associated with vernal pools, delineating, conserving, and managing these landscapes result in many challenges and conflicts.

Coast Ranges

TA B LE 4 .1 Geographic distribution of the dominant soils inventoried by soil survey in the Coast Ranges

Geology and Geomorphology The Coast Range Province extends from the Transverse Ranges to the Oregon border for approximately 1,000 kilometers. From the Pacific Ocean, the Range extends roughly 130 kilometers to the Central Valley. The east-west extension of the Coast Range is considerably less in northern California, where it borders the Klamath Mountains. The Ranges were created by transform motion and compression along the Pacific and North American plate boundary approximately 3.5 million to 5 million years ago. Thus the Coast Ranges are quite young and the associated topography reflects this, with steeply sloping dissected uplands separated by ephemeral stream valleys (Harden 2004). The Coast Ranges are derived from rocks of the Franciscan Assemblage, Great Valley sequence, and Salinian block. The Franciscan Assemblage and Great Valley sequence consist primarily of interbedded sandstone and shale. Chert, conglomerates, schists, and serpentinite occur to a lesser extent. Uplift, faulting, compression, and folding of the interbedded marine deposits has resulted in a highly complex sequence of contrasting lithologies throughout the Coast Ranges. The Salinian block occurs west of the San Andreas fault, primarily in the southern half of the Coast Ranges. It is mainly granite. Volcanic rocks consisting of rhyolite, basalt, pyroclastic flows, and andesite are exposed along the central coast and east of the San Andreas fault. The age of these rocks becomes progressively younger from the south where rocks are up to fifteen million years old, to the north in the Clear Lake volcanic fields where they are as young as ten thousand years (Harden 2004). A common geomorphic feature, particularly in hillslopes formed from sandstone and shale, are spoon-shaped hollows, which are concave hillslopes mantled by thick colluvium. They are formed by landslides caused by rapid uplift and slope failure along contact zones of different rock types and along joints and fractures. Over time, these hollows fill with sediment. Periodically, during winter storm events, soils toward the base of these hollows become saturated with water, and high pore water pressures cause debris flows (Dietrich and Dorn 1984). The hummocky landscape throughout the Franciscan terrain reflects the dynamic nature of this landscape.

Soils Despite the complex lithology and topography of the region, soil variability is relatively low from a taxonomic standpoint (Table 4.1). Mollisols are common throughout the west side of the Coast Range, where temperatures are mild relative to warmer inland soils. The milder temperatures may encourage soil organic matter accumulation by slowing microbial activity and the decomposition of plant residues. Mollisols are also common in soils derived from shale, because this rock weathers rapidly to smectite clays and soil organic matter accumulates preferentially on smectite (Gonzalez and Laird 2003). Alfisols, with thinner A horizons and less soil organic matter than Mollisols, are common throughout the inland areas of the Coast Ranges, where precipitation is moderate and temperatures are high. Weakly developed soils (Entisols and Inceptisols) occupy steep slopes, convex landforms, and steep, south-facing slopes in semiarid regions of the Coast Ranges

Soil order

Percentage of region Geographic extent

Entisols

19

Steep slopes where erosion outpaces the rate of soil development, or along floodplains

Inceptisols

15

Convex ridge tops and backslopes throughout the region

Mollisols

35

Valley floors footslopes, toeslopes, and concave hillslopes

Alfisols

19

Throughout the Coast Ranges

Ultisols

0.8

Found on older stable landforms in areas of high precipitation

Vertisols

8

Lowlands and lower slope angles of hillslopes across the region

Aridisols

3

Southeastern side of the Coast Ranges and on some south-facing slopes throughout the southern half

Histosols

0.03

Found in wetlands

Other

0.4

Typically rock outcrop, water, and urban landscapes

source: USDA-NRCS Soil Survey Geographic Database

(Beaudette and O’Geen 2008). They are common on granitic parent materials. Desert soils (Aridisols and Entisols with an aridic soil moisture regime) are found in the southeastern Coast Ranges, particularly on south-facing slopes. Highly weathered Ultisols are on old, stable landscape positions in the northern Coast Ranges, where precipitation is high. Soils derived from volcanic rocks, such as the Pinnacles Formation, are mostly Mollisols, and some Entisols on steep south-facing slopes. Vertisols are common in lowland positions, where fine-textured alluvium has been deposited and downslope drainage yields high silica and base cation concentrations favorable for smectite clay formation. The eastern zone of the southern and central Coast Ranges is dominated by the Moreno shale, which is old sea floor material containing selenium-rich pyrite (iron sulfide). As the shale weathers, the sulfides are oxidized to sulfuric acid, creating an extremely acidic soil environment (pH 30% slopes) at the source and a bench (≤15% slopes) where the material is deposited. Groundwater flow exposed by the landslide produces springs, so stabilized landslide benches often have wet meadows with a dense growth of rushes, sedges, forbs, and grasses. Topographic convergence of water in the wet meadows gives the soils an aquic moisture regime and predominantly reducing conditions that inhibit organic matter decomposition. Consequently, soils are organic-rich Mollisols and in some places they are Histosols. Soils on the surrounding well-drained landscape, including the landslide scarp and flanks, have a xeric moisture regime reflective of the climate (Lee et al. 2004). Weathering in the xeric soils produces smectite clay that accumulates as clayey argillic horizons. These soils are Mollisols. Erosion from the surrounding landscape, particularly after forest fires, delivers fine sediments to the wet meadows. These fines, together with smectite precipitated from soil solutions that drain from upslope, yield thick, clayey sediments in the wet meadows. Soil development is minimal at high elevations, on young deposits, and on steep, eroded slopes. Soils on these sites are generally Inceptisols. Aluminosilicate-rich rock within ultramafic terrain gives rise to soils and vegetation that stand in contrast to those of the ultramafic areas. This is particularly

true if the rock contains Ca-bearing minerals (Lee et al. 2001). Volcanic ash from the Cascade volcanoes to the east, present in small amounts in most A horizons of this region, is another exogenous source of potassium, calcium, and aluminum in these ultramafic soils (Lee et al. 2004).

Soils of Metamorphic Terrain At low and mid-elevations, most soils derived from metasedimentary and metavolcanic rock have argillic horizons. Maximum clay contents range between 15% and 45% and aluminum-rich clay minerals dominate (e.g., kaolin; Graham et al. 1990), regardless of mean annual precipitation (MAP). In the wetter, western part of the Klamath Mountains, leaching has depleted base cations so the soils are Ultisols. In the drier, eastern part of the region, leaching is less and the soils are Alfisols, with Mollisols at the driest margins. Inceptisols are on the younger, less stable, or higher elevation sites. At MAPs of 500 to 1,800 millimeters, soil pH is mostly between 5.5 and 7, but at higher MAP soil pH drops as low as 4.5 due to accumulation of aluminum on cation exchange sites. Nevertheless, the ratio of exchangeable Ca:Mg is always above 0.3, mostly above 1, and as high as 12. In general, forest vegetation on nonultramafic soils does not experience a harsh geochemical environment, but at high elevations, such as on the Siskiyou crest, the dense red fir forest is interrupted in places by barren areas. These areas are predominantly on quartz muscovite schist. Their origin is unclear but may be related to effects of wildfire and overgrazing (Atzet and Wheeler 1982). Severe herbivory by gophers keeps these barrens devoid of tree seedlings. Only sparse, low-growing herbaceous plants inhabit the barrens, leaving the soil open to rainsplash and sheetwash erosion. Nevertheless, the organic matter that is produced by the herbaceous plants is mixed deeply by the gophers, resulting in 70-centimeter-thick A horizons, while those in the red fir forests, where there are no gophers, are only half as thick (Laurent et al. 1994). In some cases extreme acidity results when ammonium released from the mica structure undergoes nitrification (Dahlgren 2005). Levels of exchangeable calcium in the upper 30 centimeters of soil are four times higher under red fir compared to the barren areas (Laurent et al. 1994). Landslide topography is prevalent in metasedimentary terrain, resulting in scarps with poorly developed soils and benches with deep, productive colluvial soils. As in serpentinitic terrain, landslide benches may contain springs and small wetlands.

Other Mountain Soils Soils derived from granitic rock tend to be light yellow–​brown and are coarser-textured than those from metamorphic or ultramafic rock. Granite contains a substantial amount of quartz, which is resistant to weathering and persists as sand grains in the soils. Soil textures are generally loamy sand to loam. Weathering of feldspars and biotite produces clays, which, under conditions of sufficient geomorphic stability, are translocated to the subsoil to form an argillic horizon. In the wetter western side of the province, soils are more leached of base cations and are Ultisols. In the drier eastern side, soils are Alfisols. At high elevations, or on less stable geomorphic positions, soil development is less and the soils are Inceptisols. Soil pH is typically in the range of 6 to 7 and increases Geomorphology and Soils   57

upward in the soil profile. In the western higher rainfall areas, soil pH is about one unit lower. When the parent rock contains more mafic minerals, as in diorite or gabbro, the soils have more clay (sandy clay loam textures) and are redder. At mid- and low elevations both metamorphic and granitic bedrocks are fractured to such an extent that they provide access to woody roots, and weathering has made them porous enough to hold appreciable plant available water. At higher elevations soils are more likely to be underlain by hard rock because cold temperatures have impeded weathering and glaciation has scoured to hard bedrock within the past ten thousand years. In some of the high-elevation basins, with seasonally perched or high water tables, Spodosols have formed because the coarse textures allow thorough leaching by the abundant snowmelt. Organic acids from conifers, shrubs, and litter layers are dissolved in the soil solution. They act as chelates to keep iron and aluminum soluble so they can be translocated to the subsoil to form the characteristic Spodosol morphology. Soils on marble and limestone are less acidic than soils on metamorphic and granitic parent materials, with pH values as high as 8.2 in the lowest horizons and decreasing upward to about 6.5 in A horizons. These soils are thin on exceptionally steep, rocky slopes and are thick in deep colluvium on lower slopes. They are mostly loamy Mollisols.

the land area in this region consists of volcanic plateaus and lava flows that extend for great distances. For example, the Tuscan Volcanic Plateau extends from near Lassen Peak to the Sacramento Valley. The Medicine Lake Highland region, east of Mount Shasta, consists of more than one hundred cinder cones, domes, and lava flows superimposed on a basalt shield volcano (Southard and Southard 1989). Many of the steeply sloping mountains are prone to slope failure, particularly if their rocks have been hydrothermally altered such that primary minerals have been converted to clay, thereby destabilizing the bedrock. High-elevation peaks have been shaped by water erosion and powerful glacial activity. Intermediate elevations experienced thinner sheetlike glacial ice advances. Plateaus are nearly level in the Sacramento Valley and become steeper toward the source. Dissection by streams and rivers has produced deep canyons, but the intervening uplands tend to undulate gently (Alexander et al. 1993). Glacial outwash and recent alluvium occupy valley landscape positions. Much of the region has experienced recent volcanic activity, and as a result, rhyolitic ash blankets a significant part of the terrain. The complex terrain and sequence of volcanic and glacial events has produced soils on a variety of parent materials such as lava, volcanic ash, mudflows, alluvium, colluvium, and glacial till.

Valley Soils

Soil Landscape Relationships

Soils in the intramontane valleys are formed in alluvium, including some glacial outwash, such as on the west side of Scott Valley. Soils of floodplains and alluvial fans of the larger tributaries to the valleys are Entisols. They are very coarsetextured and have very low water-holding capacities. In the basins of large valleys, such as Scott Valley, soils have a finer texture (e.g., loam, sandy clay loam, silty clay loam), a seasonally high water table that results in reducing conditions and gray subsoil colors, and accumulation of humified organic matter resulting in black A horizons. These soils are Mollisols with an aquic moisture regime. Some of them may be alkaline, with pH values ranging from 8.2 to 9. Soils in wide parts of river canyons, such as at Happy Camp and Seiad Valley, are formed in alluvium of floodplains and successively older, higher terraces. The floodplain soils are very coarse-textured Entisols with low water-holding capacities. On successively higher terraces, soils are progressively older, more weathered (Inceptisols-Alfisols-Ultisols), contain more clay, and have higher water-holding capacities. While some floodplain soils are in riparian areas, the older terrace soils are elevated and may be some distance from the current stream channel.

The degree of soil mineral weathering dictates the spatial distribution of soil orders across the region. Soil development in volcanic materials evolves over the trajectory from EntisolsAndisols-Inceptisols-Alfisols and/or Ultisols. Entisols are found in relatively recent deposits of volcanic ejecta where soil formation has been limited to the accumulation of soil organic matter (Figure 4.5). Over time, volcanic parent materials are altered to poorly crystalline minerals that impart unique chemical and physical behaviors and result in soils classified as Andisols. With increasing time and weathering, more crystalline minerals such as kaolin are formed. This intermediate stage of weathering gives rise to Inceptisols. As time progresses, clays are translocated into subsurface layers, creating argillic horizons in Alfisols and Ultisols (Takahashi et al. 1993). Andisols occupy 15% of the California Cascade region and the distribution of these soils is controlled by elevation and proximity to volcanic activity. Andisols are common in areas where volcanic ejecta have been replenished regularly throughout the Holocene. They are also commonly found at higher elevations (approximately 1,700 to 2,050 meters) where weathering is limited by cool temperatures. They are generally not found at highest elevations where cold temperatures inhibit weathering that produces poorly crystalline minerals. Conditions that encourage the evolution of Andisols into Alfisols and Ultisols include a combination of mild to hot temperatures and high precipitation, which promotes the illuvial translocation of clays and leaching of base cations. These conditions tend to exist at intermediate elevations, around 500 to 1,500 meters, where precipitation falls predominantly as rain. Alfisols are the most extensive soil order in the California Cascades, covering 39% of the region. Ultisols form in this region only under optimal conditions, including stable landscapes with gentle slopes, warm temperatures, and high rainfall. Thus the distribution of these soils is limited, occupying about 2% of the region. Inceptisols and Entisols together cover 38% of the region and are found at the

Cascade Range Geology and Geomorphology Cascade volcanoes have produced a spectrum of extrusive igneous rocks (basalt, andesite, rhyolite) that span a range of susceptibility to weathering. Volcanic activity in the Cascade Range is caused by subduction of the Gorda plate beneath the North American plate. Many of the volcanoes in the Cascade region have a classic cone shape created by multiple sequences of eruptions. Mount Shasta and Mount Lassen are the two largest volcanoes of the Californian Cascade region. Much of 58  Drivers

Southern Cascades Soil Orders Alfisols Andisols Entisols Inceptisols Mollisols Ultisols Vertisols

Area shown

FIGURE 4.5 Geographic extent of soil orders in the California Cascade region. The high peak in the upper left is Mount Shasta. Map by California Soil Resource Lab http://casoilresource.lawr.ucdavis.edu/. Source: USDA-NRCS Soil Survey Geographic Database upscaled to 1 km grid.

highest elevations, on very steep slopes, or on recent lava and mud flows. The degree of weathering and distribution of soils is constrained by proximity to recent volcanic activity where intermittent deposition of volcanic ash supplies new material, rejuvenating the profile and resetting the pedogenic time clock. In addition, a slope threshold was recognized for this region where soil development was found to be greatest on slopes less than 30% (Alexander et al. 1993). As is the case throughout the state, climate and the climatemoderating effects of topography have a strong influence on soil development. In an elevation gradient extending from 270 to 2,030 meters in the southwestern Cascades, soil base saturation and pH decrease with increasing elevation due to increased leaching of the soil profile and changes in vegetation (Alexander et al. 1993). Clay content and soil depth are greatest in the lower portion of the mixed conifer zone at elevations between 500 and 1,000 meters, reflecting the favorable weathering conditions associated with mild temperatures and high precipitation. Soil organic carbon content is constant below 1,700 meter elevation where temperatures are warmest, but increase at higher elevations. The relationship between elevation and soil organic matter is likely driven by differences in soil temperature. Cold temperatures slow microbial activity, resulting in a greater stock of soil organic residues. Warm temperatures stimulate microbial activity, promoting rapid decomposition of soil organic matter. The degree of weathering as regulated by the age of volcanic deposits is illustrated by a chronosequence of mudflows on the slopes of Mount Shasta. Soils formed on deposits ranging in age from twenty-seven to twelve hundred years (Dickson and Crocker 1953a, 1953b, 1954). Soil properties—​including pH, base saturation, organic matter, and water-holding capacity—​changed rapidly over this short pedological timeframe. For example, soil organic matter was found to reach steady state within five hundred years. Valley and basin soils are typically very deep. Many basin soils are poorly drained due to a high groundwater table. The mild temperatures and poor drainage give rise to soils with large carbon stocks. Many lowland soils are Mollisols.

The Cascade Range contains active and inactive hydrothermal areas, which have a large impact on soil properties. Soil temperature, acidity, and chemical/mineralogical properties vary dramatically over short distances in hydrothermal areas and impact vegetation establishment and succession. An active fumarole in Lassen Volcanic National Park produced a thermal gradient ranging from 100°C to 15°C, along which soil pH increased from strongly acid (pH 1 vector were apportioned among categories based on relative contributions of each vector to introductions.

As a result of differences in the numbers of introductions and the characteristics of receiving environments, some ecosystems and regions in California have many invasive species while others have relatively few. In California, marine, alpine, and forest ecosystems tend to have fewer exotic alien species, while freshwater and grassland ecosystems are frequently highly invaded (Mooney et al. 1984). These differences likely reflect a combination of ecosystem characteristics, human land use intensity and history, and patterns of introductions. In particular, the number of introductions (or propagule pressure), and especially the number of intentional introductions, is increasingly identified as an important predictor of the number of established invaders (Leprieur et al. 2008, Keller et al. 2011, Kempel et al. 2013). Marine ecosystems generally have fewer invasive species than many other ecosystems in California, a pattern that is true for marine systems worldwide (Lockwood et al. 2013). This may be due to a relative lack of introductions or limited human ability to alter natural conditions. Early detection of an invasion of the Mediterranean marine weed Caulerpa taxifolia in nearshore southern California sea-bottom environments, most likely from improper aquarium dumping, allowed it to be successfully eradicated in 2006. Compared

to other ecosystems, there also have been few intentional introductions of marine fishes (Schroeter and Moyle 2006). The unconfined nature of most marine ecosystems limits the ability of invaders or people to strongly alter conditions in marine environments to favor invasive species (Schroeter and Moyle 2006, Lockwood et al. 2013). A critical exception is San Francisco Bay—​a relatively confined, estuarine system that by some measures is the most invaded ecosystem in the world (see Chapter 19, “Estuaries: Life on the Edge”). In contrast, freshwater fish communities in California are highly invaded. This is likely due to a long history of fish stocking, high volumes of traffic in systems such as the Sacramento–​San Joaquin River Delta, and other sources of introductions such as release from aquaria and dispersal through canals (Fuller 2003, Leprieur et al. 2008). In addition, people have widely altered conditions in freshwater ecosystems, including temperature, water quality, and hydrologic regime. In California almost two-thirds of non-native freshwater fishes (both established and not) are from within the U.S.—​ either from another state or from another part of California (Fuller 2003). Finally, California grasslands are highly invaded (Minnich Biological In vasions   235

2008). Of the eleven hundred established non-native plant species in California, three hundred occur in grasslands, and their dominance is nearly complete in many areas (D’Antonio et al. 2007). Much research has attempted to establish why grasslands became so invaded so soon after European arrival in California. Likely this was due to a combination of disturbance through intensive grazing and cultivation (Stromberg and Griffin 1996, Minnich 2008), severe drought at a critical time early in the invasion process, and many intentional and unintentional introductions of exotic grasses for sheep and cattle forage (D’Antonio et al. 2007). Grassland invaders extend into oak savannas and woodlands and into the highly invaded coastal sage scrub and somewhat less invaded chaparral ecosystems at lower elevations and at the coast. Despite marked variation in the degree of exotic species dominance and impact across California’s diverse ecosystem types, no ecosystem type entirely escapes the effects of biological invasions. Every ecosystem chapter in this volume describes a broad range of introduced species, ranging from pathogens such as the fungus responsible for sudden oak death (SOD) that affects redwood forests and oaks

(see Box 13.2; Chapter 25, “Oak Woodlands”) to introduced trout in high alpine lakes (see Chapter 32, “Lakes”) to introduced mammals offshore in the Channel Islands (see Chapter 34, “Managed Island Ecosystems”) and exotic insect pests in agricultural areas (see Chapter 38, “Agriculture”). Although these introduced species vary widely in their effects and are welcomed in some human-dominated systems (Coates 2006; see Chapter 39, “Urban Ecosystems”), many exotic invaders have negative or undesirable effects on California’s biological diversity and on ecosystem functions and services valued by California's human communities.

Impacts of Invasive Species in California Impacts of invasive species on their co-occurring native species and host ecosystems are far ranging—​from genetic impacts on individuals, to extinctions from a community, to altered nutrient cycles and disturbance regimes at an ecosystem scale, sometimes with direct impacts on human well-being (Perrings 2011, Schierenbeck 2011, Lockwood et

BOX 13.2  SUDDEN OAK DEATH In the mid-1990s, forest managers and scientists across California were alerted to damaged and dying tan­ oaks (Lithocarpus densiflorus) and coast live oaks (Quercus agrifolia) infected by an unknown pathogen. The first cases appeared in 1994, in tanoaks in Marin and Santa Cruz Counties. Shortly thereafter, the pathogen was identified as Phytophthora ramorum, which had also been recently found in rhododendron (Rhododendron sp.) and viburnum (Viburnum sp.) in Western Europe (Lane et al. 2003, Rizzo et al. 2002). P. ramorum soon acquired widespread, epidemic proportions along a 300 kilometer strip of the central California coast (Garbelotto et al. 2001), infecting a large number of trees in redwood and oak woodland ecosystems. Because of its unexpected appearance and aggressiveness, the disease came to be known as sudden oak death (SOD).

B O X 1 3 . 2 F I G U R E 1   Die-off of tanbark oak (Lithocarpus densiflorus) resulting from the spread of sudden oak death (SOD) syndrome through coastal California forests. Photo: U.S. Forest Service.

236  Biota

Phytophthora ramorum is an oomycete, which are commonly known as wet molds due to their similarity with fungi. As pathogens, oomycetes can have both ecologically and socially devastating impacts. For example, the potato blight that contributed to the mid-nineteenthcentury Irish famine was caused by the aggressive oomycete Phytophtora infestans (Haas et al. 2011). Until very recently, details about the origins of P. ramorum invasion in California were mostly unknown. However, new computational and genetic techniques, with intensive efforts in spatial sampling coverage, have allowed the precise reconstruction of SOD history and pathways of invasion in California. Croucher et al. (2013) conclude that the pathogen arrived into California through infected nursery plants starting in Santa Cruz and Marin Counties, and that the epidemic in California stems from only three or four individuals that evolved from a single genotype. In addition, Croucher and his colleagues found initial evidence that some of the evolved genotypes identified in California seem better adapted to forest environments than others. Further research on this particular topic might inform forest management strategies against the spread of the pathogen. As for its global historical distribution, specific differences in the mating types between European and North American populations, along with a single dominant genotype and a clonal population structure, have led to the conclusion that it originated in a previous, third location. China has been suggested as the most likely origin because of its native abundance of Phytophthora species, particularly in Yunnan province (Goheen et al. 2005). SOD is able to infect a wide array of species and to use some of them as reservoirs (Grünwald et al. 2008, Garbelotto et al. 2003); soon after the first cases of affected tanoaks and coast live oaks were detected, it was found that P. ramorum was in fact able to infect many of the dominant trees of the northern and cen-

al. 2013). At the smallest scale, researchers are beginning to examine the impacts of invasive species on the genetics of native populations. When introduced species spread into an area, there may be some exchange of genetic material between native and invasive populations. This often happens via hybridization, as is the case with the highly invasive smooth cordgrass (Spartina alterniflora), originally introduced to salt marshes along the San Francisco Bay. Smooth cordgrass can hybridize with its native congener, California cordgrass (Spartina foliosa), and the resulting hybrid has spread to colonize previously open mudflats (Daehler and Strong 1997). Similarly, rainbow trout (Oncorhynchus mykiss), a California native introduced as a game fish to many of California’s lakes and streams, can hybridize with locally endemic trout subspecies in California (California golden trout [O. mykiss aguabonita] and Paiute cutthroat trout [O. clarki seleniris]), substantially altering the genetic structure of the native populations (Busack and Gall 1986, Cordes et al. 2006). At the population scale, native species can be influenced by invasive species in a wide variety of ways, and many native populations in California are threatened by invasive species.

tral California coast. Despite its name, the disease also infects coastal redwoods (Sequoia sempervirens), big leaf maple (Acer macrophylla), bay laurel (Umbellularia californica), Douglas-fir (Pseudotsuga menziesii), and Pacific madrone (Arbutus menziesii) (Grünwald et al. 2008). Infected species differ in their susceptibility to SOD, influencing the spread of the disease. For example, bay laurel seems to be almost unaffected by infection. This allows P. ramorum to persist in bay laurel throughout the dry summer until the wet season, when the pathogen can spread and infect other individuals and species. Thus bay laurel appears to serve as a reservoir for transmission, facilitating the infectious cycle. This was demonstrated in an eight-year study by McPherson et al. (2010), which found that tanoak and coast live oak infection and mortality rates increased with the presence and abundance of bay laurel. SOD’s transport and dispersal mechanisms provide it with high mobility. Once spores are produced, they are easily splashed or carried down to the soil by rain or water from sprinklers. If spores reach streams or agricultural runoff, they can potentially be carried over long distances. Human shoes, bicycle wheels, and the use of outdoor or camping equipment may also facilitate transmission among distant sites (Davidson and Shaw 2003, Fisher et al. 2012). Specific native and introduced beetle species also facilitate the spread and impacts of SOD. Beetles such as the oak ambrosia beetle (Monarthrum scutellare) and oak bark beetles (Pseudityophthorus pubipennis) predispose coast live oak and black oak (Quercus kelloggii) to infection by attacking and wounding trees, increasing wood decay, and interrupting sap flow (Švihra and Kelly 2004). Moreover, beetle species that have symbiotic relationships with specific tree fungal pathogens (e.g., Monarthrum dentigerum, Xyleborinus saxeseni, and Xyleborus californicus) can colonize bleeding cankers caused by SOD, further

Data for plant invasions are especially good. Over one thousand naturalized plant species are thought to occur in California (Rejmánek and Randall 1994, Rejmánek 2003, Stohlgren et al. 2003). According to some analyses, these exotic species have already contributed to the extinctions of 14 plant species and to declines in another 709 plant species that are either listed or proposed for listing as endangered by the threat of invasive species (Seabloom et al. 2006, Stein et al. 2000). The highest numbers of invasive plant species in California occur in coastal areas, which also harbor the highest number of threatened native plant species (Seabloom et al. 2006). In other taxonomic groups, several introduced diseases have had major impacts on their host populations. These include sudden oak death, which is decimating populations of susceptible oak trees around the state (Rizzo and Garbelotto 2003); chytrid, a fungal disease that targets amphibians (Briggs et al. 2010); and the relatively newly arrived West Nile virus, which has caused population declines in corvids (crows and jays) in addition to presenting a threat to human health (Koenig et al. 2007). Invasive animals also often displace native species; for example, the invasive bullfrog (Litho-

infecting trees with their associated fungal pathogens and speeding mortality by up to 65%. At the community and ecosystem scales, changes in the composition of several coastal forests in California are already occurring due to SOD-caused mortality (Box 13.2 Figure 1). These changes in turn cause alterations in fire regimes, chiefly by increasing fuel material at different forest strata (Metz et al. 2012). Alterations of nutrient dynamics occurs through changes in forest composition, which in turn affect litterfall quality and quantity (Cobb et al. 2013) as well as the ectomycorrhizal fungi associated with tanoaks (Bergemann et al. 2013). Although SOD is causing tremendous damage to California’s coastal forests, some ecophysiological and management mechanisms are emerging as options to fight the epidemic. Scientists have found that some coast live oaks naturally produce high concentrations of phenolic compounds and tyrosols in their phloem. These biochemical defenses against the pathogen can be effective at preventing infection or inhibiting it once the tree is infected (Nagle et al. 2011). Management efforts have been growing based on new understanding of the disease and its hosts. Since there is no known cure once a tree is infected, most management strategies are aimed at preventing the spread of the disease to new areas and protecting susceptible individuals. These include inspections of nursery plants prior to purchases, removal of infected oaks, removal of nonoak host trees, disposal of plant debris, and monitoring activities (Alexander and Lee 2010). More recently, application of phosphonate, an environmentally benign, narrow-spectrum fungicide, has been shown to temporarily deter SOD infection and spread within an individual. Forest managers and scientists are looking for ways to improve and expand the use of phosphonates as an effective management tool (Garbelotto et al. 2013, Schmidt and Garbelotto 2010).

Biological In vasions   237

bates catesbeianus) reduces populations of native frogs, such as the mountain yellow-legged frog (Rana muscosa) and the California red-legged frog (R. draytonii) (Kupferberg 1997). California harbors many dramatic examples of invasive species altering the areas they invade at the community scale (the composition and relative abundances of constituent species). In many cases, even if individual native species are not directly threatened with extirpation by the arrival of an exotic species, native species abundances plummet such that previously common species become rare. Invasive grasses (including Avena barbata, A. fatua, Bromus diandrus, B. hordeaceous, and Hordeum murinum) have replaced native (and largely perennial) grasses in over 9 million hectares of California grasslands. Cape ivy (Delairea odorata) invasion in California coastal areas can cause a 30% reduction in native species diversity (Alvarez and Cushman 2002). Invasive Argentine ants (Linepithema humile) displace most native ant species, leading to the homogenization of ant communities in invaded areas (Box 13.3; Holway and Suarez 2006). The composition of the aquatic flora and fauna of the San Francisco Bay has been transformed by invasive species, such that the majority of species as well as over 90% of the biomass within the Bay are non-native (Cohen and Carlton 1998). Islands are particularly invasible, and because of their often-endemic species communities, the impacts of invaders on island communities can be especially strong. Feral sheep and pigs on Santa Cruz Island, for example, consumed large amounts of native vegetation before their removal, causing the near-extinction of several endemic species, transforming forests and shrublands into invasive-dominated grasslands and leaving behind so much bare ground that they significantly increased the frequency and size of landslides (Van Vuren and Coblentz 1987). In many cases, invasive species can also have large-scale impacts on ecosystem properties. These include influences on nutrient cycling (Ehrenfeld 2003), hydrology (Levine et al. 2003), and disturbance regimes (Mack and D’Antonio 1998). Generalities about the influence of invasive species on ecosystem properties have been hard to come by—​impacts tend to be related to traits of the individual invasive species and can be variable and unpredictable. Salt cedar (Tamarix spp.), intentionally introduced in the intermountain West, now grows along arid rivers and waterways in parts of California, transpiring at a higher rate than native vegetation in the upland areas that it invades. By doing so, salt cedar can reduce streamflow, dry ephemeral desert pools, and increase aridity in areas it has invaded (Zavaleta 2000). Disturbance regimes with which native species have coevolved, including fire frequency, erosion, and biotic disturbances, can also be influenced by invaders. For instance, Great Basin desert areas dominated by annual invasive grasses are prone to increased fire frequency, while areas of coastal California with invasive pigs (Sus scrofa) have increased levels of soil turnover and disturbance (Mack and D’Antonio 1998). In addition to the many ecological effects of invasive species are their great economic costs and impacts on ecosystem services. Environmental, economic, and health costs of invasive species are associated with increased wildfire potential, reduced water resources, erosion, flooding, threats to wildlife, losses in crops, forests, fisheries, and grazing capacities, diminished outdoor recreation opportunities and spread of diseases, among other costs (Elton 1958, Mack et al. 2000, Puth and Post 2005, Perrings 2011). Invasive plants can cause problems as weeds in agricultural fields or can be toxic to domestic and stock animals; invasive pathogens cause dis238  Biota

eases in citrus crops; and a variety of aquatic invaders can cause ship fouling. While it is often very difficult to attach specific values to the impacts of invasive species, a few efforts have been made to understand the true economic impact of various invaders (Duncan et al. 2004, Pimentel et al. 2005). Approximately six hundred species of introduced insect and mite pests affect crops in California; these species collectively account for about two-thirds of crop losses in California (Dowell and Krass 1992). The value of field crops in California in 2010 was approximately $5.5 billion (CDFAb), so losses due to invasive pests are likely on the order of hundreds of millions of dollars annually. For example, to avoid potential yearly losses of $228 million in food crops to the establishment of exotic fruit flies in the state, the California Department of Food and Agriculture has spent approximately $3.9 million in projects aimed at the eradication of fruit fly outbreaks (CDFA 2009). Each feral pig is conservatively estimated to cause $200 in crop and environmental damage yearly, though this estimate does not include many of the indirect effects caused by feral pigs (Pimentel et al. 2005). Unifying ideas for predicting the impacts, and thus the environmental and economic costs, of invasive species have been hard to come by. However, on a case-by-case basis, the risk posed by a single invasive species can be conceptualized as a function of both the probabilities of introduction, establishment, and spread of a species and the magnitude of potential impacts of that species. Some researchers have argued that non-native species should be judged problematic and removed or prevented from spreading only if harmful impacts are apparent (Davis et al. 2011). Others argue that given the great difficulty in predicting potential impacts, especially as both climate and the distributions of other species change, the precautionary principle should guide our approach to invasive species (Simberloff et al. 2005).

Invader Management: Challenges and Success Stories Prevention Prevention is both the most successful and cost-effective strategy for the management of potential invasive species, but it is only possible when exotic species are prevented from entering a pathway or when transported organisms are prevented from release or escape (Lodge et al. 2006). Prevention policies vary in their geographic scale of focus, with a number of local, state, and federal agencies, NGOs, business interests and research institutions working together to prevent exotic species introductions in California. Because many of the worst invasive species arrive via international transport, many policies addressing introductions are in the form of international agreements. For example, in 2005 the International Maritime Organization passed guidelines for preventing introductions via ship ballast water and sediments. A challenge to preventive policies is that the World Trade Organization, which promotes international trade, does not allow member nations to block entry of trade items without a strong case that they are harmful. At the national level, U.S. policy towards introductions has been mostly ad hoc; for example, the 1900 Lacey Act allows the federal government to prevent importation of blacklisted animal species but does not specify how animals should be

BOX 13.3  THE ARGENTINE ANT IN CALIFORNIA First documented in the United States in 1891, the Argentine ant (Linepithema humile) made its way to California in 1907 (Woodworth 1910) and continues to spread (Holway 1995, Sanders et al. 2003). Many people living in coastal California are familiar with the Argentine ant as a pest that can swarm homes during the wet months of the year. The ant is native to northern Argentina, southern Brazil, Uruguay, and Paraguay. While its introduced distribution is global, the Argentine ant most commonly occurs in areas similar to its Mediterranean-climate region of origin (Roura-Pascual et al. 2006). Human-mediated dispersal, via cargo shipments and other forms of transportation, is largely responsible for the spread of the Argentine ant (Holway 1995). Without human interference these ants disperse slowly overland; along with transportation, high colony fertility and an ability to nest in a wide variety of environments have facilitated their spread (Holway 1995, Suarez et al. 2001, Silverman and Brightwell 2008). Behavioral changes within their introduced range have also allowed Argentine ants to succeed as invaders (Buczkowski et al. 2004). In their native range, Argentine ants coexist with other ant species and compete with other colonies of Argentine ants (Tsutsui and ­Suarez 2003). However, in California these ants behave as supercolonies; competition among Argentine ant colonies disappears. The high level of cooperation observed within the introduced range of this species is likely due to high levels of genetic similarity in the invasive population (Tsutsui and Suarez 2003). Low genetic variation prevents ants from distinguishing close kin (Tsutsui et al. 2001, Tsutsui and Suarez 2003). In turn, cooperation allows Argentine ants to reach very high densities and competitively displace native ants (Holway 1999, Tsutsui et al. 2000), leading to a host of ecological problems. Argentine ants are used as a model system to study the mechanisms behind community disassembly and other community processes because of the frequently disruptive effect they have on native communities (e.g., Sanders et al. 2003). In habitats invaded by Argentine ants, native ants decline in both abundance and diversity (Human and Gordon 1997). Argentine ants appear to have a high tolerance for fragmented and disturbed landscapes and tend to be more abundant near development and exotic vegetation (Suarez et al. 1998, Laakkonen et al. 2001), making these habitats even more hostile for native ant species. Argentine ants have also been implicated in the declines of both coast horned lizards (Phrynosoma coronatum) and shrews (Notiosorex crawfordi) in southern California. These declines appear to be caused by the invader’s displacement of native harvester ants, which are an important food source for both declining species (Laakkonen et al. 2001). Non-ant invertebrates also show marked declines in diversity and abundances in

the presence of Argentine ants (Human and Gordon 1997). In agricultural settings the Argentine ant contributes to high densities of insect pests such as the California red scale (Aonidiella aurantii), mealybugs (e.g., Pseudococcus viburni, P. mauritimus), and aphids (Aphis gossypii) (Lach 2003, Daane et al. 2007, Powell and Silverman 2010) (Box 13.3 Figure 1). These small insect pests suck phloem from crop plants and secrete a sugary, honeydew substance, which is then collected by the ants. In exchange, the ants defend these insects from predation and parasitism. The net effect of the Argentine ant on crops is unknown. Although the ants protect the small sap-sucker pests from predators, they also inadvertently protect the plant from other, potentially more damaging insect herbivores (Lach 2003).

B O X 1 3 . 3 F I G U R E 1   The Argentine ant (L. humile) tending aphids on a coyotebrush (Baccharis pilularis var. consanguinea). Photo: Dan Quinn, Jasper Ridge Docent.

Management of the Argentine ant has proven extremely difficult, and early detection and prevention of spread are critical. The ant is particularly difficult to manage because it has no known natural enemies (Orr et al. 2001). As a result, the primary control method is poisonous chemical baits, but even these are frequently ineffective because the ant can typically draw from other food resources. To date the only documented country eradication has occurred in New Zealand (Harris et al. 2002). Field experiments indicate that several integrated pest management (IPM) strategies might be successful in controlling these ants if their access to habitat, resources, and/or nesting sites is limited (Silverman and Brightwell 2008).

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TA B L E 13 . 3 California codes and agencies involved in the prevention of exotic species introductions.

State code

State agency

Invasive species managed

California Food and Agriculture Code

Department of Food and Agriculture

Animal pests, plant diseases, noxious weeds

California Fish and Game Code

Department of Fish and Wildlife

Live wild animals and plants, aquaculture

California Water Code

State Water Resources Control Board Regional Water Quality Control Boards

Biological pollutants in water

California Harbors and Navigation Code

Department of Boating and Waterways

Aquatic weeds

California Public Resources Code

State Lands Commission Department of Fish and Wildlife

Marine organisms in ballast water from commercial vessels

assessed for blacklisting. Federal agencies involved in preventing introductions include the U.S. Department of Agriculture (USDA) and its Marketing Service, Animal Plant Health Inspection Service, the U.S. Fish and Wildlife Service, the U.S. Department of Commerce, and the U.S. Coast Guard (CDFG Invasive Species Program 2008). At the state level, the California legislature has passed more specific regulatory controls under the Fish and Game, Food and Agriculture, and Public Resources Codes, while the California Water and Harbors and Navigation Codes also address aquatic invasion management (CDFG Invasive Species Program 2008) (Table 13.3). Many state agencies in California are involved in the detection and management of invasive species. Codes established by the state legislature give state agencies regulatory power to manage different categories of invasive species. California’s Department of Food and Agriculture is primarily responsible for operating agricultural inspection stations set up at entry points throughout the state to prevent introductions of exotic agricultural pests. Prevention policies vary as to whether they focus on deliberate versus inadvertent introductions (Simberloff et al. 2005) and whether they target particular species versus introduction pathways based on enforcement at entry points such as ports and state boundaries (CDFG Invasive Species Program 2008). While deliberate introductions account for the naturalization of more than half of all invasive plants (Mack and Erneberg 2002), conflicting interests and unpredictability of invasions impedes stricter limitations, such as on the importation of ornamental plants. Importers and retailers who directly facilitate the transfer of species across borders lack economic incentives to limit introductions, and federal policy does not lay out a coherent approach to limiting introductions. While the impacts of species introductions are inherently unpredictable, current legislation reflects more of an “innocent until proven guilty” approach to invasion prevention. Moreover, even with quantitative risk analysis calculations, risk estimates lack confidence limits and can give a false sense of safety (Simberloff et al. 2005).

Early Detection, Rapid Response, and Eradication Some species will inevitably go undetected at their points of entry despite efforts to monitor potential introduction path240  Biota

ways. When prevention is ineffective, a small portion of introduced species will establish populations, at which point eradication may be a potential management strategy. The success of eradication efforts in most areas hinges on detection during the early stages of invasion, which can be accomplished with standardized monitoring programs, active surveillance, and cooperation with information networks (Lodge et al. 2006, CDFG Invasive Species Program 2008). In the early stages of invasion, introduced species occupy a limited range and may be functionally innocuous. For example, eradication of exotic weeds in California is usually successful when infestations are smaller than one hectare but is often unrealistic when infestations are greater than 1,000 hectares (Rejmánek and Pitcairn 2002). An incipient population may persist in this way for a variable amount of time (weeks to years) (Lodge et al. 2006) before it experiences rapid growth and range expansion. This lag time provides the best window of opportunity to take management action. Eradication is often feasible when incipient invasions are identified early and have a limited range, and response is rapid (Simberloff et al. 2005). Island invasions are especially good candidates for eradication efforts due to their smaller typical size and their isolation. Eradication of mammals from islands is a powerful conservation tool that is gaining popularity as success stories are documented in the literature (Donlan et al. 2003), although eradication of one invasive may facilitate the invasion of other species. The eradication of black rats (Rattus rattus) on Anacapa Island, part of Channel Islands National Park, in 2001–​2002 is a story of success through meticulous planning and collaboration. The presence of the endemic Anacapa deer mouse (Peromyscus maniculatus anacapae) and other protected seabirds required careful planning, anticipation, and mitigation of nontarget effects. Follow-up monitoring suggests that negative impacts were short-lived, but that the benefits to native seabirds resulting from the removal of rats from the island are ongoing (Whitworth et al. 2005, 2013). Like islands, freshwater systems have been devastated by invasive species and are relatively isolated and well delineated, often making eradication of invasive species feasible. The eradication of invasive pike from Lake Davis and of invasive trout from many high Sierra lakes to protect threatened salmon, other native fishes, and frogs are the best known freshwater eradication successes (see Chapter 11, “Biodiversity”). A host of opportunities persist to prevent species

extinctions and protect ecosystem properties through welldesigned freshwater invasive species eradications in California. Successful eradications have also taken place outside of the confines of islands and fresh water systems in California. “Killer alga” (Caulerpa taxifolia), a fast-growing marine alga listed as one of the world’s one hundred worst invaders by the International Union for the Conservation of Nature, was introduced at two sites in southern California in 2001–​2002. The San Diego Regional Water Quality Control Board recognized the potential cost of the invasion without immediate action, comparing the infestation to an oil spill, and was able to mobilize emergency funding to begin eradication efforts in just seventeen days (L.W.J. Anderson 2005). The eradication was declared successful in 2005. Although federal regulations prohibit the importation, interstate sale, or transport of the aquarium strain of Caulerpa taxifolia (Southern California Caulerpa Action Team 2003), many species of Caulerpa, including C. taxifolia, are available online (Walters et al. 2006). After successful eradication, preventing new infestations is an important way to ensure long-term success. Cost-effective management must include monitoring for early detection, which should be weighed against the estimated costs of eradication and the even more costly alternatives of long-term control. Surveys to detect small populations are expensive, but if a small invasive population goes undetected, eradication can quickly become more costly (Lodge et al. 2006). Similarly, post-eradication monitoring is important to ensure that eradication has been achieved or to identify the need for additional control before the invasion reexpands significantly and inflates follow-up costs. Eradication of most invasive species requires a combination of culling and follow-up control methods (Parkes et al. 2010). The initial (one- to several-year) cost of aggressive eradication programs can be high, but the alternative—​long-term (but incomplete) control of invasive species—​requires ongoing effort, expense, and an indefinite amount of time and resources (Zavaleta et al. 2001, Ewel et al. 1999).

Long-term Control and Management Ongoing control efforts for harmful invasive species require coordination across scales and reassessment over time to consider population changes and to incorporate learning about exotic species effects and their responses to implemented controls. In general, long-term control strategies are costly and appropriate mainly for established species with significant impacts on biodiversity, human health, infrastructure, and other values. Other established species with less severe impacts are in practice accepted and managed as ecosystem components, in some cases even incorporated into providing services such as habitat for native species (e.g., Shapiro 2002). At the state level, the California Department of Fish and Wildlife’s (CDFW) Invasive Species Program uses specific criteria and expert input to determine whether or not to take action to control an invasive species (CDFG Invasive Species Program 2008). At the county level, Cooperative Weed Management Areas (CWMAs) typically organized by agricultural commissioners’ offices and consisting of  local stakeholder groups, work together on weed management.  Such organizations have increased from fewer than twenty in the state in 2000 to over forty today, covering all counties (California Invasive Plant Council 2014). CWMAs can foster coordination among managers, increasing effectiveness and reduc-

ing costs of invasive control (Epanchin-Niell et al. 2009). A more recent trend has led to increased multicounty, landscape-scale coordination of invader mapping and prioritization efforts. Both county- and multicounty-scale efforts can ultimately lead to more successful and cost-effective control and eradication than ad hoc, local activities, which are more likely to lead to reinvasion (Doug Johnson, pers. comm.). Effective control should limit the dispersal and reduce densities or extents of invasive species as well as consider effects of control activities on native species and ecosystems of special concern and on human health and well-being. Many different tactics can be used in long-term control, such as hunting to control feral pigs, timed grazing to control invasive grassland weeds, and large-scale efforts to disrupt the mating of exotic agricultural pests. Wild pigs were intentionally introduced as livestock and for hunting and are now found throughout most of California (Waithman 2001). While most pigs in California are managed through hunting, pigs on Santa Cruz Island were eradicated in 2006 using a combination of pig-proof fencing, aerial and on-the-ground hunting, and Judas pigs (sterilized, radio-collared females) tracked to locate remaining individuals (Parkes et al. 2010). The invasive northern pike (Esox lucius) was similarly introduced illegally in California to improve recreational fishing. In a series of attempts to eradicate pike in northern California, lakes and tributaries were treated with piscicides to kill all fish then restocked with rainbow trout (Lee 2001, Vasquez et al. 2012). Northern pike reintroductions are prohibited in California, but illicit fish stocking is responsible for the introduction of many exotic fish species (CDFW 2015). Strategic education and interagency coordination can reduce reintroduction risks (Johnson et al. 2009). Both control and eradication can, and often must, be complemented with additional restoration measures to successfully recover target ecological values and reduce the likelihood of reinvasion (Zavaleta et al. 2001) and to ensure that the removal of one invasive does not facilitate invasion of another species (Erskine Ogden and Rejmánek, 2005). Moreover, restoration efforts themselves can reduce invasion, such as when replanting and vegetation restoration reduce the availability of open, disturbed sites for reinvasion. Various state and national government agencies manage the spread of some invasive species through restoration. Restoration of invaded areas often involves the removal of invasive species and replanting with native species but can also involve measures such as restoration of flood flows in rivers, temporary erosion prevention structures, and reintroduction of animals such as seabirds. Ongoing control and restoration efforts can be labor-intensive and often rely on large numbers of community volunteers. This underscores the value of prioritization focused on harmful invaders and incorporating the feasibility of successful removal and restoration, especially when underlying changes that facilitate invasion (such as altered flows in dammed and diverted streams and rivers) cannot be remedied.

Biocontrol Biocontrol is a special type of long-term management in which natural predators, parasites, pathogens, or competitors of invasive species are intentionally introduced as control agents (UC IPM 2007). Classical biological control relies on the introduction of one or more additional exotic species when the benefit of introducing the control agent is expected Biological In vasions   241

to outweigh the inherent risk of introducing additional exotic species. Effective biocontrol species are generally invertebrates or pathogens that have tightly coevolved with an invasive species in its native range and that play a significant role in regulating its population there. Generalist vertebrates are poor biocontrol agents because of their propensity to switch to native prey species. The U.S. Department of Agriculture’s Animal and Plant Health Inspection Service, Plant Protection and Quarantine must authorize any import of a new biological control agent into the country (Scoles et al. 2012), and in California the state Department of Agriculture maintains guidelines for biocontrol introductions (CDFA 2013a, Aslan et al. 2014). International organizations also provide standards and guidelines for biological control (Tanaka and Larson 2006). Biocontrol introductions should use cost-benefit analyses that incorporate both economic and ecological consequences, weighing ecological damage against the consequences of inaction. An evolutionary and landscape perspective would also improve cost-benefit analysis, as introduced species can acquire new hosts, adapt to new environmental conditions, and disperse to new areas. Ecological damage has probably occurred more often than recorded due to minimal monitoring of post-introduction, nontarget effects of biocontrol efforts (Simberloff and Stiling 1996). Biocontrol has rarely been used to manage invasive species that impact primarily natural ecosystems. However, it has been widely and successfully used in agricultural systems to manage crop pests and noxious weeds in pasture lands. A classic example of biological control that originated in California is the control of the cottony cushion scale (Icerya purchasi). The scale insect was established in the Los Angeles area and developed as a major citrus pest, causing massive destruction of citrus trees. In its native range in Australia, the scale insect was not known to be a pest as populations were controlled by the Australian vedalia ladybeetle (Rodalia cardinalis) and a parasitic fly (Cryptochaetum iceryae). Vedalia beetlesand the parasitic fly were captured in Australia and released in California in 1888. Within a year and a half, control was achieved and has been maintained with the establishment of wild populations of both biocontrol agents (Pedigo and Rice 2009). Biocontrol has also been a successful strategy for controlling some noxious weeds in California grasslands. St. John’s wort, or Klamath weed (Hypericum perforatum), was first identified in California near the Klamath River and became an invasive weed by the 1920s (DeBach 1974). Klamath weed, which is toxic to cattle, invaded grasslands and became the target of control efforts. Klamath weed was successfully controlled by the introduction of Chrysolina beetles that suppress the weed in its native range (Huffaker and Kennett 1959). In contrast to the successful control of Klamath weed, numerous biocontrol efforts have not successfully controlled yellow starthistle, one of the most economically and ecologically damaging exotic invasive plants in California (see Box 13.1; CDFA 2013d).

The Future of Biological Invasions in California The future of invasions in California will be determined by both environmental changes and changes in how we respond to invasive species. The interacting effects of international trade, land use change, atmospheric pollution, and climate change will alter the suite of new species introduced, iden-

242  Biota

tities of new and formerly benign non-native species that become invasive, habitat susceptibility to invasion, and invader impacts. However, changes in public understanding of the problems posed by invasive species, prioritization by managers of invasive species removal as both critical and possible, and capitalization on past successes and new techniques and technologies could dramatically reduce the future impacts of harmful invasions on California’s ecology and society. Increasing transport, maritime trade, international air travel, and online commerce have the potential to increase the number of species introduced to California and the likelihood and rates of establishment and spread (Perrings et al. 2009). For example, the top ten U.S. container ports—​which include both Los Angeles and Oakland in California—​experienced a 54% increase in container movements from 2001 to 2006, and projections indicate that U.S. port container traffic will double by 2020 and triple by 2030 (U.S. Maritime Administration 2009); all else being equal, the arrival of potentially invasive species in California is expected to increase. The conditions that determine whether newly or previously introduced exotic species will establish, spread, and become problematic are also changing due to atmospheric pollution, global climate change, and land use change in California. Specific impacts are often difficult to predict. For example, elevated concentrations of CO2 (due to increasing greenhouse gas emissions) can preferentially benefit some introduced species, but its impacts are less clear in the context of responses by other native species and of other, ongoing global changes (Dukes and Mooney 1999). Nevertheless, some average trends indicate likely directions for biological invasions under changing conditions. The traits that make many species successful invaders, such as high plasticity (Davidson et al. 2011), rapid dispersal, and generalist habit (Dukes and Mooney 1999), will also likely serve them well under changing conditions in California. In general, disturbances from fire to land use change are expected to increase in California and to continue to facilitate invasions as climate change continues and the state’s population approaches fifteen million by midcentury (California Department of Finance Demographic Research Unit 2013; see Chapters 5, “Population and Land Use,” and 14, “Climate Change Impacts”). Extreme events such as floods, droughts, and heat waves are also expected to increase in magnitude and frequency, which could increase invasions and their impacts through a variety of mechanisms (Diez et al. 2012). For example, when the marine epibenthic fouling community of Bodega Harbor was exposed to a simulated heat wave, nonnative species were better able to tolerate the initial disturbance than native species and to maintain their dominance in the months following the event (Sorte et al. 2010). Where invasive species are currently limited by cool temperatures, warming could facilitate spread. For example, winter frosts currently limit the pink bollworm (Pectinophora gossypiella), a major cotton pest, to southern California’s desert valleys. As winter frosts decrease in the San Joaquin Valley, the survival and spread of the pink bollworm is expected to increase (Gutierrez et al. 2006). In aquatic systems non-native animal species have, on average, a performance advantage over native species associated with increases in temperatures and CO2 levels (Sorte et al. 2013). Warming temperatures could also favor the expansion of warm-water freshwater species native to the eastern portion of the U.S., such as largemouth bass, green

sunfish, and bluegill that were originally introduced for sport fishing (Moyle et al. 2013). Though several factors point to a worsening biological invasions problem in California, innovative policies and responses to prevent introductions and control spread have the potential to counter this scenario. Invasive species management efforts in California increasingly include formal prioritization: evaluating which invasive species are worth aggressively managing versus tolerating or embracing, balancing negative impacts with limited resources, and recognizing that some invasive species could provide ecosystem services that partially compensate for the damage they cause (Doug Johnson, pers. comm.). Technology innovations could also aid with prevention and detection of invasive species in the future. For example, smartphone apps allow an increasingly connected public to submit geo-tagged photos of suspected weeds and pests, facilitating early detection and control. Higher-resolution remote sensing and imaging spectrometers (Simberloff 2013) offer promising approaches to detecting and monitoring invasions, providing repeated cover over wide and less-accessible areas (Vitousek et al. 2011). Improvements in DNA sequencing will enable better detection of introduced species, allowing quick species identification as well as detection of particular species at very low densities, thereby aiding in prevention efforts (Simberloff 2013, Cross et al. 2011). Beyond techniques and technological innovations, the potential exists for attitudes towards invasions among both the public and decision makers to shift. Increased attention to invasive species management successes, ranging from eradications in freshwater and island systems to successful biocontrol efforts in agricultural systems, could supplant a sense of hopelessness about highly invaded systems in the state and could encourage more use of successful approaches towards a wider range of targets. A continued trend towards distinguishing clearly harmful invaders, of which there are many, from relatively innocuous and even positively valued nonnatives (e.g., Coates 2006), allows both managers and the public to concentrate effort and attention on the highest priorities for action. Finally, our responses to global change and invasions interact, meaning that priorities, policies, and management practices will need to be revised to take advantage of synergies and avoid unintended consequences of management practices. For example, climate change may undermine the assumptions used in established risk-assessment protocols for preventing introductions based on climate matching (Pyke et al. 2008), or could cause biocontrol organisms to become less effective (Gutierrez et al. 2008) or to become invasive themselves (Pyke et al. 2008). On the other hand, land use policies intended to reduce greenhouse gas emissions might also reduce rates of landscape fragmentation that accelerate invasions (Pyke et al. 2008); California’s Sustainable Communities and Climate Protection Act of 2008, SB 375, could play such as role. Invasive species might even deliberately be used to maintain key ecosystem functions in the face of threats like sea level rise to coastal ecosystems (Hershner and Havens 2008). Exotic biological invaders have dramatically influenced California’s ecosystems. However, California has also been—​a nd hopefully will continue to be—​an important global stage for continual innovations and evolving understanding about how to both harness established exotics for good and effectively prevent and reverse harmful invasions.

Summary Invasive species have had tremendous impacts on the ecosystems and economy of California. Both ecological and human factors govern the dynamics of invasions in California. California’s equable Mediterranean climate, diversity of geologic and climatic conditions, and widespread natural disturbances such as fires and floods contribute to surplus resources and favorable habitats for potential invaders. But patterns of invasions in California are most strongly dictated by both historical and modern human impacts on the introduction, establishment, and spread of invasive species. Major phases of terrestrial biological invasions in California have been defined by shifts in human land use following the first European settlements and exacerbated by an increasing human population and improvements in transportation infrastructure. California’s unique human history has driven patterns of both intentional introductions for food, pets, sport, and horticulture as well as unintentional (though preventable) introductions via high international and interstate trade and traffic. Humans have also increased the process of establishment and spread for invasive species by altering existing disturbance regimes, patterns of connectivity, and abiotic conditions. In general, densely populated areas of California, especially along the coast, harbor the highest numbers and associated impacts of introduced species. California’s open ocean, alpine, and forest ecosystems tend to have fewer invaders, while its freshwater, grassland, and estuarine ecosystems are generally highly invaded. Documented impacts of invasive species on native species in California include genetic impacts, local or species-level extinctions via disease and displacement, changes in community composition and native species diversity, and altered ecosystem processes such as nutrient cycling and disturbance regimes. Economic costs include direct losses to crops or managed resources and investments in management and control efforts. Costs also accrue from decreased ecosystem services, such as decreased climate and flood regulation and reduced water resources. Preventing introductions of potentially invasive species in the first place is the most successful and cost-effective management strategy. For species that establish populations in California, management options include eradication targeted at high-priority, early-stage invaders; long-term control efforts such as weeding, hunting, timed grazing, and mating disruption; and biological control, which has been used to manage invasive agricultural pests and noxious weeds. The future of invasions in California will be shaped by increases in international trade and transport, the state’s population, climate change and disturbances, and innovations and investments in management. A great many successful eradication and control efforts illustrate that invaded areas can be restored, while a great many invasions have also become irreversible parts of California’s diversity and ecological dynamics.

Acknowledgments We thank Peter Moyle, Marcel Rejmánek, and Richard Hobbs for greatly improving the chapter.

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Recommended Reading Bossard, C. C., J. M. Randall, and M. C. Hoshovsky. 2000. Invasive plants of California wildlands. University of California Press, Berkeley, California. Minnich, R. A. 2008. California’s fading wildflowers: Lost legacy and biological invasions. University of California Press, Berkeley, California. Richardson, D. M., editor. 2011. Fifty years of invasion ecology: The legacy of Charles Elton. Wiley-Blackwell, Hoboken, New Jersey. Thompson, L., G. A. Guisti, K. L. Weber, and R. G. Keiffer. 2013. The native and introduced fishes of Clear Lake: A review of the past to assist with decisions of the future. California Fish and Game 99:7–​41. Walters, L. J., K. R. Brown, W. T. Stam, and J. L. Olsen. 2006. ­E -commerce and Caulerpa: Unregulated dispersal of invasive ­species. Frontiers in Ecology and the Environment 4:75–​79.

Glossary Alien  A species or taxon that is not native to a given location (i.e., that has been introduced by human activity). Alien taxa include beneficial (e.g., domestic), benign, and harmful species and also include both naturalized and unestablished taxa. Synonyms include exotic, non-native, nonindigenous, and introduced. Biocontrol  A special type of long-term management in which natural predators, parasites, pathogens, or competitors of invasive species are intentionally introduced as control agents (UC IPM 2007). Also known as biological control. Congener  A species within the same genus as another species. Disturbance regime  This concept describes the dominant patterns and sources of biophysical change in a given ecosystem that trigger processes of ecological succession at different spatial and temporal scales; or the pattern of disturbance (for example, fire or flooding) that shapes an ecosystem over time. Endemic  Describes a taxon that occurs only in a particular geographic region. Exotic  See Alien. Feral  Describes an organism or taxon that has escaped or been released from captivity or domestication. Hybridization  The spread of genes of one species into the genes of a different species. Intentional introduction  Purposeful introduction of an exotic species to a new environment by people. Intentional introductions include introductions of ornamental plants, fish and game species, and biocontrol agents. Introduced species  See Alien; however, “introduced” can refer to that subset of alien species that reach unconfined habitats—​for example, captive pets can be considered alien but not introduced unless they are released from captivity. Introduction  The process of introducing an alien species into a new environment or locale. Invasibility  A characteristic of an ecological community describing the ease with which exotic species can become established and invasive in that community. Invasive  Describes a non-native species that establishes selfsustaining populations and spreads outside of its native range; a subset, harmful invasive species cause environmental and/or economic harm in their introduced ranges. Naturalized species  The term has been applied to describe a variety of non-native species conditions. It is considered an imprecise terminology but most recently refers to established, consistently reproducing species outside their native range, 244  Biota

whether or not they are harmful or spreading (Richardson et al. 2011). Nonindigenous  See Alien. Non-native  See Alien. Oomycete  A relatively large taxon of eukaryotic organisms belonging to Heterokontophyta phylum along with brown and golden algae. They resemble fungi by presenting mycelial growth and in their nutrition habits. Pathway  The route through which an exotic species arrives in a new location, such as a particular shipping route. Propagule  A seed, a spore, or any part of an organism that is capable of producing a new individual. Riparian  The zone surrounding a river or stream on either side that is strongly influenced by the presence of the river. The riparian zone also directly influences the river through shading or inputs of organic material.

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FOURTEEN

Climate Change Impacts CHR ISTOPHER B. FIELD, NONA R . CH IAR IELLO, and NOAH S. D IFFENBAUGH

Introduction Throughout this volume, climate emerges as a fundamental force shaping the distribution and dynamics of every ecosystem type in California. Climate changes thus implicitly are expected to affect all of California’s ecosystems, and indeed they are identified in many chapters of this book as central forces of change in the present and future. While ecosystemspecific effects and projected changes are elaborated by chapter, we provide an overview of expected climate changes and their general impacts in California. We highlight examples from throughout the state of the types of direct and indirect climate change effects that now influence biodiversity and processes within and across ecosystem types. This chapter explores the possible trajectories and consequences of continued climate change during the coming century. Projections of future impacts depend on a variety of unknowns and assumptions, foremost what trajectory of ongoing greenhouse emissions the world pursues in the coming century. For this reason, impact projections are generally referenced to specific, future emissions scenarios that range from no mitigation (continuation of recent trends in emis-

sions growth) to ambitious mitigation. Future impacts are almost universally far greater under scenarios of no mitigation and smaller under aggressive mitigation scenarios. However, past emissions already ensure a degree of committed climate changes that will continue to unfold over the coming centuries regardless of future emissions changes.

Climate Change in California Since the beginning of the twentieth century, California has warmed by approximately 0.9°C, close to the global average. The pattern across most of the state was one of rapid warming from about 1910 to 1935, cooling through the 1940s and 1950s, and rapid warming since about 1970 (Kadir et al. 2013). All regions of the state have warmed, though warming over the past forty years has been smaller in the North Coast region than in the rest of the state (Kadir et al. 2013), perhaps as a result of increased persistence of summer fog in the most coastal stations (Cordero et al. 2011). 251

Photo on previous page: The south fork of the Feather River feeding Lake Oroville, a reservoir formed by the Oroville Dam, on September 5, 2014. Photo: Kelly M. Grow, California Department of Water Resources. 252  Biota

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Annual average precipitation over California has not changed consistently over the past century (Kadir et al. 2013), but the warmer temperatures tend to increase the fraction of precipitation falling as rain rather than snow. This is reflected in a decreasing fraction of annual flow that occurs during the spring snowmelt season in most of the rivers that drain the Sierra Nevada range and in decreased spring snowpack in the Northern Sierra. In the Southern Sierra, where elevations are higher, spring snowpack has increased by about 10% over the past sixty years, approximately the same magnitude as the decrease in the Northern Sierra. Sea level along much of California’s coast has risen by 15 to 20 centimeters, approximately the global mean increase, over the past century. At the upper end of this range is the Golden Gate in San Francisco, which has the longest continuous sea level record for North America (Flick et al. 1999). Due to land subsidence, uplift, and other processes that change the land surface elevation, sea level change as recorded by long-term tide gauges has varied from +2.08 mm yr-1 at La Jolla to –​0.73 mm yr-1 (decreasing sea level) at Crescent City (National Research Council 2012). Over the twenty-first century, climate change can be characterized as falling into two eras. The next few decades will be an era of mostly committed climate changes that are largely independent of investments in mitigation, including nearterm investments. During this era, mitigation will be important, but its consequences will emerge only gradually. This is largely because cumulative emissions, which control the magnitude of climate change (IPCC 2013), separate gradually over time among even sharply contrasting emissions scenarios (Figure 14.1). In the latter decades of the twenty-first century, investments in mitigation become dominant influences on climate, leading to an era of climate options. Substantial inertia in both physical and socioeconomic aspects of the climate system means that the set of achievable options in the future depends on mitigation efforts to reduce greenhouse gas emissions soon, during the era of near-term committed climate change, as well as during the era of climate options. By midcentury (2041–​ 2 060), California will begin to emerge from the era of committed climate change, with warming in the range of 1°C to 2°C above the late twentieth century for a scenario with aggressive mitigation and warming of 2°C to 3°C for one that continues recent trends in emissions growth (Figure 14.2). By the end of the century (2081–​ 2100), this difference is dramatic, with warming of 1°C to 2°C above the late twentieth century in a scenario with aggressive mitigation and 4°C to 5°C in a scenario that continues recent trends. For many kinds of climate change impacts, the difference between a world that is approximately 2°C warmer than in preindustrial times and one that is more than 4°C warmer is likely to be dramatic, underscoring the importance of considering a range of possible futures. Nonetheless, even with 1°C warming, the occurrence of hot extremes is projected to increase over much of California (Diffenbaugh and Ashfaq 2010, Scherer and Diffenbaugh 2014). In contrast to the clear trends in temperatures, projected California precipitation does not change consistently. The global projections, with generally wetter polar regions and drier subtropics, especially for scenarios with large amounts

Year FIGURE 14.1 Time trend of observed and projected global annual temperature relative to 1986–​2 005. Projections and a measure of uncertainty (shading, representing ± 1.64 standard deviations) are for two scenarios or Representative Concentration Pathways (RCPs) with anthropogenic radiative forcing in year 2100 of 2.6 W m-2 (RCP2.6) or 8.5 W m-2 (RCP8.5). Projections for the two scenarios are from Phase 5 of the Coupled Model Intercomparison Project (CMIP5) of the World Climate Research Programme, representing two dozen modeling centers. Source: IPCC 2013.

of warming, place California in a zone on the border between wetter and drier. Without clear changes in projected annual precipitation, other aspects of the water cycle—​including changes in snow versus rain, snowpack melting date, interannual variability and extremes—​could be very important (Diffenbaugh et al. 2015). Warmer temperatures in California’s future will have a range of direct impacts on California ecosystems. Some of these are described in this chapter. Many of the impacts of a changing climate will be indirect. Some of these are physical. For example, warmer temperatures in the Sierra Nevada will increase the fraction of precipitation that falls as rain, further decreasing spring snowpack and probably lengthening the period during summer when mountain ecosystems experience drought. Climate model simulations indicate that Sierra Nevada spring snowpack at the end of the twenty-first century could be in the range of 27% of current values for a scenario with little warming to 11% of current levels in a high-emissions, business-as-usual scenario (Hayhoe et al. 2004). Other examples of physical indirect effects could include changes in coastal fog (Snyder et al. 2003) and risk of wildfire (Westerling et al. 2011).

Climate Change in a Multistressor Context Another class of potential indirect effects is biologically mediated. Changes in the abundance or distribution of competitors, herbivores, pathogens, pollinators (Gilman et al. 2010, Aslan et al. 2013), and invasive species (Clements et al. 2004, Hellmann et al. 2008, Sandel and Dangremond 2012, Bellard et al. 2013) can have profound effects on the success of individual species and the functioning of ecosystems. In general, impacts of direct effects are better understood than are impacts of indirect effects, but a large body of literature documents the potential importance of indirect effects. Some of this comes from research focused on climate change. Much more comes from basic research in population, community, and ecosystem ecology. In most parts of the world, but especially in California, climate change impacts on ecosystems occur in the context of a wide range of interacting anthropogenic impacts. The most important of these for ecosystems currently are changes in the composition of the atmosphere, extensive changes in

Observations

Projections

Annual temperature

1901-2012

RCP8.5

Mid-21st century Late 21st century

But over coming decades, the relative roles of direct impacts of climate change, indirect impacts of climate change, nonclimate anthropogenic factors, and interactions among all of these will shift toward an increasing role for climate change, especially in a high-emissions world with large amounts of warming. The relative roles of climate change and other factors could also vary among ecosystem types.

RCP2.6

Observed Impacts of Climate Change to Date

0 2 4 6 Trend (˚C over period) 0

2 4 6 Difference (˚C)

Mid-21st century Late 21st century

RCP2.6

Annual precipitation

RCP8.5

1951-2010

Trend (mm yr -1decade-1) -20 0 20 40

Difference (%)

Solid color White

Significant trend Insufficient data

Diagonal lines Trend not significant

Solid color

Very strong agreement

White dots Strong agreement Diagonal lines Little or no change Divergent Gray changes

FIGURE 14.2 Observed changes in annual temperature and precipitation in California during the twentieth century and projections for the twenty-first century. Projections are for two scenarios: a low greenhouse gas emissions scenario (“peak and decline”), RCP2.6, and a high greenhouse gas emissions scenario, RCP8.5. Projections show differences in the mid-twenty-first-century period of the CMIP5 RCP8.5 model ensemble, calculated as 2046–​ 2065 minus 1986–​2 005, as well as differences in the late-twentyfirst-century period of the CMIP5 RCP8.5 ensemble, calculated as 2081–​2100 minus 1986–​2 005. Colors represent multimodel means. Source: Diffenbaugh et al. 2014.

land cover and land use, and the presence of large numbers of non-native plants and animals. All these factors can have profound effects on biodiversity (Sala et al. 2000) and ecosystem function (see Chapters 11, “Biodiversity,” and 15, “Introduction to Concepts of Biodiversity, Ecosystem Functioning, Ecosystem Services, and Natural Capital”). Over the next few decades, effects of these other factors may well be more important than effects of climate change (Settele et al. 2014).

Climate change impacts on California ecosystems are evident from many case studies, most of which have focused on species or species assemblages whose historical distribution, abundance, or other properties are well documented in surveys or collections. The historical material provides baseline information for comparison with contemporary conditions. Through careful resurveys, matched as closely as possible to the original measurements and sites, these studies generally have examined ecological changes in relation to recorded climate change in circumstances where other causal explanations could be ruled out. In some cases, the ecological history is embedded in the current environment in forms such as plant architecture, tree stumps, or ocean sediment, providing site constancy and a continuous or near-continuous record of ecological change. A study of twentieth-century climate warming and expansion of subalpine conifers in the central Sierra Nevada exemplifies this approach (Millar et al. 2004). At treeline sites across eastern Yosemite National Park and National Forest lands, krummholz whitebark pine (Pinus albicaulis) doubled its yearly branch elongation during the twentieth century and the flat-topped krummholz developed a more conical treelike form due to release of vertical branches. On northfacing slopes on the eastern slope of the Sierra Nevada, historically barren snowfields below treeline were invaded by lodgepole pine (Pinus contorta), western white pine (Pinus monticola), and whitebark pine from surrounding closed forest. Invasion of subalpine meadows by lodgepole pine changed meadows from sharply defined communities dominated by grasses, sedges, and forbs to mixed herbaceous-tree communities. All four of these trends—​branch elongation, vertical branch release, snowfield invasion, and subalpine meadow invasion—​were derived from field measurements and correlate with the regional minimum monthly temperature, which rose by 3.7°C during the twentieth century (Millar et al. 2004). This study illustrates the rigor with which ecosystem change can be attributed to climate change. First, the four ecological indicators were examined in separate sites over an extensive area, yielding four independent correlations with regional temperature change. Second, the ecological responses are consistent with the role of severe climate in the spatial distribution of subalpine communities (see Chapter 28, “Subalpine Forests”). Third, other potential explanations such as land-use history and substrate type could be ruled out. Fourth, both the temperature record and the temporal detail in the ecological indicators showed a multidecadal signal that aligned with phases of the Pacific Decadal Oscillation (PDO), strengthening a climate-based explanation (Millar et al. 2004). By far the best explanation for the observed changes is that trees began to establish and change growth form in areas that previously were too cold or had snowpack for too much of the growing season. Climate Change Impacts   253

254  Biota

0.7 0.6

Normalized cover

In mountain regions a species’ distribution can be expected to shift upslope to where the amount of climate warming is compensated by the lapse rate —​that is, the elevation-driven temperature decrease—​which on average is 6.5°C for each kilometer of elevation increase (Lundquist and Cayan 2007). By this reasoning, each 1°C of warming should lead to about 154 meters of upward shift in elevation. Initially, distributional changes might be more subtle than a range shift. One example comes from the Deep Canyon Transect, a 2,314 meter elevation gradient in the Santa Rosa Mountains that was surveyed in 1977 and again in 2007 (Kelly and Goulden 2008). Ten dominant species characterize the transect as it rises from desert scrub through pinyon-juniper woodland, chaparral shrubland, and conifer forest. Comparing the two surveys, no shifts occurred in range limits during the intervening thirty years, but for every species except desert agave (Agave deserti), cover increased toward the upper range limit and decreased toward the lower (Figure 14.3). In 2007 the cover-weighted distribution of each species was centered 65 meters higher on average than in 1977, an amount consistent with the lapse rate and observed increases in mean and minimum temperature (Kelly and Goulden 2008). Lower range limits can be set by high-temperature stress or by interactions between temperature and other factors. Increased aridity is generally expected as climate warms, primarily due to greater warming of land than oceans and, consequently, greater evaporative demand (Sherwood and Fu 2014). In mountain regions this effect is compounded by reduced snowpack, which not only lengthens the snowfree season but also reduces the water bank available to meet the more prolonged and intensified evaporative demand of a warming climate. In the Sierra Nevada, evidence is mounting that warming-mediated drought stress is pervasive in forests at lower elevations and is driving upslope retreat of ponderosa pine (Pinus ponderosa). Over a period of six decades, its lower range limit has moved upslope by about 180 meters, retreating from areas where monthly minimum temperatures in winter previously dropped below freezing but now stay consistently above freezing. The finding is based on a comparison between historical vegetation maps from the 1928–​1940 Wieslander Vegetation Type Map (VTM) Survey and a 1996 survey by the U.S. Forest Service (Thorne et al. 2006). Although the loss of mature pines in the area of retreat was due primarily to timber harvest or other disturbances, new ponderosa pine stands failed to recruit even where fire, urbanization, and conversion to grassland could be ruled out as causal factors. Ponderosa pine forest has been replaced mostly by montane hardwood forest and annual grasslands (Thorne et al. 2006). Although there are geographic uncertainties associated with the VTM Survey (Keeley 2004, Kelly et al. 2007), the retreat of ponderosa pine is on a scale much greater than these spatial uncertainties (Kadir et al. 2013). Mechanisms that drive plant mortality have tremendous leverage on ecosystem structure and function in a time of climate change, but they remain poorly understood (Anderegg et al. 2012). Mortality in old-growth forest stands in Sequoia and Yosemite National Parks has been analyzed with physically based models to determine whether temperature or water deficit plays a bigger role (Das et al. 2013). In low-elevation forests, mortality is best accounted for by water deficit, whereas at higher elevations, temperature alone can account. The results suggest that climate change can be viewed as

0.8

1977 2007

Elevation Shifts

0.5 0.4 0.3 0.2 0.1 -400

-200

0

200

400

Elevation (m) FIGURE 14.3 Asymmetrical shift to greater upslope cover of domi­ nant plant species along the Deep Canyon Transect in southern California’s Santa Rosa Mountains. Points represent the mean total normalized vegetation cover for the ten most widespread species surveyed. Surveys were conducted at fixed elevation intervals along the transect in 1977 and 2008. Elevation is referenced to the central point along the elevation gradient for each of the ten most widespread species. Source: Kelly and Goulden 2008.

shifting the transition from water-limited to energy-limited forests along elevation gradients (see Chapter 27, “Montane Forests”). Temperature effects on pathogens and insect populations could be the proximate causes for warming-driven forest mortality at high elevation (Das et al. 2013). Large trees are expected to be more tolerant of climate change, but this expectation may not be correct. In Yosemite National Park the density of large-diameter trees in plots surveyed during the 1990s was three-fourths of what the Wieslander surveys recorded in the 1930s (Lutz et al. 2009). Sierran shifts have also occurred in butterflies and small mammals. Of twenty-eight species of small mammals surveyed by Joseph Grinnell in Yosemite National Park between 1911 and 1921, half showed an increase in the midpoint of their elevation range a century later (Moritz et al. 2008). Species at lower elevations expanded their range upward, and species of higher elevations showed an upward retreat of their lower limit (Figure 14.4). Between the two surveys, mean minimum temperature increased approximately 3°C (Moritz et al. 2008). Among 127 species of butterflies surveyed for three decades across an elevation range of 25 meters to 2,775 meters, mean elevation shifted upward by 93 meters (see Figure 14.4) (Forister et al. 2010). Extensive population surveys of Edith’s checkerspot butterfly (Euphydryas editha) by Parmesan (1996) throughout its documented historical range found that the butterfly’s range had shifted not only upward in elevation and but also northward, consistent with observed warming.

Marine Latitudinal Shifts Sedimentary assemblages of planktonic foraminifera provide one example of climate-related latitudinal shifts in marine communities. Carried by ocean currents, these protozoans have shells (tests) that are preserved when the organisms die and sink to the seafloor. Layered sediments in the Santa Barbara Basin provide a fourteen-hundred-year record of climate-related conditions in the California Current

8

Small mammals

6

2 1575

1275

975

675

375

75

-75

-375

0 -675

Number of species

4

Shift in midpoint of elevation range (m) 40

Butterflies

30 20 10

1600

1400

1200

1000

800

600

400

200

0

-200

-400

-600

0

Shift in mean elevation (m) FIGURE 14.4 Histograms of elevation change for species on the west slope of the Sierra Nevada.

Top: Shifts in the midpoint of the elevation range for 28 species of small mammals in Yosemite National Park based on a repeat of the Grinnell survey a century later. Survey sites range in elevation from 57 meters to 3871 meters. Seven species showed no change (column marked by an arrow). Source: Moritz et al. 2008, Table 1. Bottom: Shifts in mean elevation for 127 species of butterflies based on surveys during 1998–​2 007 as compared with surveys two decades earlier. Survey sites ranged from 25 meters to 2775 meters. Source: Forister et al. 2010.

because the shells provide a timeline of species with known temperature affinity. As the ocean warmed during the twentieth century, temperate and subpolar species became less abundant in deposited sediments and tropical and subtropical species increased, forming a modern foraminiferan assemblage unlike any found earlier in the record (Field et al. 2006). North Pacific ecosystems have witnessed temperature-related changes in other sea life as well, but the sedimentary foraminifera provide a long enough time series to separate the signal of twentieth-century warming from those of natural climate oscillations. Invertebrates of the rocky intertidal in Monterey Bay also show a northward latitudinal shift in community composition. Repetition of a 1930s survey of invertebrates, six decades later but in precisely the original plots, revealed an increase in abundance for eight of nine southern species and a decrease in five of eight northern species as shoreline ocean temperature increased by 0.75°C (Barry et al. 1995). The pattern was sustained and expanded in an enlarged survey, which also identified some southern species that were unreported in the original study, including a gastropod and the solitary form of an anemone (Sagarin et al. 1999). Natural climate oscillations occurred with similar timing relative to the studies and so could be discounted as an explanation for the changed assemblage.

Edaphic Endemics For species that have very narrow habitat requirements, the odds of successful migration from a climatically unsuitable habitat to a better one are reduced if the habitat is discontinuous. Edaphic endemics are a special case because their sub-

strates often occur as outcrops or “islands” that are small in size, with intervening substrates that disadvantage endemics in competition with soil generalists (Damschen et al. 2012). The fate of edaphically restricted species in a changing climate is especially consequential for California because its degree of edaphic specialization likely exceeds that of any other floristic province in the world (Harrison 2013). Serpentine endemics are associated with the ultramafic rocks serpentinite and peridotite, which weather to form soils that are chemically distinctive, shallow, rocky, and less productive than more typical soils. In the context of climate change, serpentine species are a pivotal case. On the one hand, their response to climate change may serve as a general model for interactions between climate change and highly fragmented habitats. On the other hand, the ability to tolerate serpentine soils may inherently confer greater tolerance of water stress, suggesting that a warmer, drier environment might affect serpentine endemics less than it would soil-generalist species, or might even shift the competitive balance in their favor (Damschen et al. 2012). Evidence to date on serpentine endemics conforms more to the first possibility—​g reater risk from climate change. In the diverse and endemically rich Klamath-Siskiyou Mountains straddling the California-Oregon border, an increase in mean summer temperature of about 2°C has affected serpentine communities even more than plant communities of more normal soils (Damschen et al. 2010). The finding is based on a repeat of a classic study in ecology, Robert Whittaker’s 1949–​ 1950 survey of the Siskiyou vegetation in relation to natural variation in soil type, elevation, and a topographic moisture gradient. In a resurvey six decades later, losses in diversity and cover occurred on both soil types but impacts were greater on serpentine soils and disproportionately affected more endemic species (Damschen et al. 2010).

Migration Timing Migratory species depend on climate and resources across areas as large as entire hemispheres. Migration timing, such as the spring arrival of migratory songbirds, is an example of phenology that is fundamental to species persistence. Using arrival dates recorded by long-term bird observatories and banding stations together with weather data, it is possible to ask whether arrival dates have changed, whether temperatures have changed during the arrival window, and whether any changes in arrival and temperature occur in concert. The answer to all three is yes for some species in northern and central California (MacMynowski et al. 2007). Arrival dates for ten of twenty-one songbird species have trended earlier over the past two to three decades. For eight of them, there is a significant association with temperature and with a large–​ scale climate index such as the El Niño-Southern Oscillation (Table 14.1). Correlations with a climate index could reflect climate changes in the wintering range or migration route (MacMynowski et al. 2007). Collectively, these case studies form a coherent picture. The link between climate change in California and changes in the persistence, distribution, abundance, or activity of species is discernable. At the cold-limited end of their distribution, many species are expanding upward in elevation or north in latitude as climate warms. At lower elevations and lower latitudes, many species are retreating, most likely because of interactions between temperature and other factors. In many, Climate Change Impacts   255

TA B L E 14 .1 Trends in spring arrival timing of migratory songbird species in central and northern California

Degree of association with climate

Arriving earlier

Arriving later

No change

Very likely climate-associated

Barn swallow Black-headed grosbeak Warbling vireo Western kingbird Wilson’s warbler

Likely climate-associated

Black-throated gray warbler Orange-crowned warbler Vaux’s swift

Cliff swallow Swainson’s thrush

Nashville warbler Olive-sided flycatcher Western wood-pewee

Possibly climate-associated

Lazuli bunting Northern rough-winged swallow

House wren

MacGillivray’s warbler Pacific slope flycatcher

No climate association

Blue-gray gnatcatcher Western tanager Yellow warbler

Source: MacMynowski et al. 2007.

if not most cases, range contractions and upslope retreat entail extinction of local populations that might be uncompensated by expansion upslope or northward, resulting in a net loss of populations. At the community level some species are responding more than others, some not at all. This could reflect differences in sensitivity to climate change or differences in response timing. In either case, differential species responses will reshape community assemblages, altering a multitude of biotic interactions. The focus on settings where climate change impacts could be evaluated in isolation from other factors can sharpen awareness of climate change but profoundly constrains a realistic picture of climate interactions that will increasingly alter ecosystems and their services and functions. Many of these case studies correlate ecological change not with mean annual temperature but with seasonal or monthly minima or maxima, consistent with the stresses that establish range limits. This adds urgency to the need for an improved understanding of the consequences of projected increases in extreme events, especially hot extremes.

Impacts of Future Climate Changes Over the next few decades, the era of committed climate changes, the climate change impacts that California will experience will be largely independent of action to address the climate challenge, making it reasonable to pinpoint a suite of impacts. The actual trajectory of climate has some uncertainty, even over the next few decades, because of uncertainty about climate sensitivity to greenhouse gases, background variability, and changes in the frequency or severity of damaging extreme events (IPCC 2012). In the last part of the twenty-first century, the era of climate options, possible futures range from those where climate is still similar to midcentury to those where the difference from preindustrial times approaches that between glacials and interglacials in the past (Diffenbaugh and Field 2013). For the era of climate options, it is useful to think about a broad range of possible ecosystem futures rather than about generic climate change impacts. 256  Biota

The situation is the same for California as for every other part of the world. A late twenty-first century with climate change stabilized at a global average warming of approximately 2°C over preindustrial levels will have ecosystems that look different and function differently from those in the world of the twentieth century. The impacts on California of climate changes to date are already widespread and consequential, and those in a 2°C world will be much larger. But a future with continuing business-as-usual emissions and global average temperatures in 2100 on the order of 4°C over preindustrial levels would be so different that the tools for projecting and describing the conditions become completely inadequate. Many studies, overviewed earlier in this chapter, address ecosystem impacts of the climate changes to date. For these, the traditional research tools of observation, analysis, modeling, and experimentation are appropriate, though with some important limitations, especially related to nonequilibrium conditions. For the +2°C world, the toolkit narrows and becomes more a sketch of possibilities than a generator of predictions. For the +4°C world, all bets are off. Many courageous investigators have conducted studies of ecosystems in +4°C worlds, and the results of those studies provide important insights. These insights need to be understood, however, as barely scratching the surface of the interacting processes that will shape the rapidly changing +4°C world. Most of the literature on climate change and climate change impacts is not this blunt about the limitations of our knowledge. That is appropriate, because most studies are focused on identifying contributions to knowledge rather than gaps. But this is a setting where it is important that we not overinterpret based on the results available. The paucity of information about the future, especially about the +4°C world, is heavily shaped by our almost total lack of even rudimentary ability to characterize effects of exceedingly rapid temporal changes or a wide range of interacting ecological and anthropogenic effects. Is it a waste of time to even consider California ecosystems in a +4°C world? Certainly not. But we should remember that the science to date is in its infancy and could well be missing or misjudging many of the most important processes.

The Velocity of Climate Change We usually think about the rate of climate change as the change in temperature per unit of time, typically in units of degrees per century or degrees per decade. For organisms that can make physiological adjustments (Cleland et al. 2012), for managers contemplating engineered changes in ecosystems (Hobbs et al. 2009, Suding 2011), and for organisms with short enough generation times for evolution to play a role (Parmesan 2006), degrees per unit time might be a reasonable measure. But for most organisms the challenge of adjusting to a changing climate is the challenge of moving to a new location when climate in the original habitat moves outside the acceptable range. As a starting point for understanding the challenge of moving in response to a changing climate, it is useful to think about how fast an organism would need to move to stay in its original climate. This defines a second kind of velocity of climate change, with units of distance per unit of time (Loarie et al. 2009). The spatial velocity of climate change has been calculated two different ways (Diffenbaugh and Field 2013). One method compares climate model output for the present and future and calculates, in the modeled output, the shortest distance at some future time to a model grid cell with the same annual average temperature (or some other climate parameter). The second method starts with the modeled temperature difference but calculates the distance to a location with the original temperature based on a high-resolution view of current spatial gradients of temperature. The first (modeled trend) method (Figure 14.5) provides a coarse-grained picture that is especially meaningful over large temperature changes and distances. The second (current gradient) method captures the important role of topography but tends to be insensitive to potentially important effects like organisms becoming squeezed off the tops of mountains. For California, both approaches reveal important aspects of the challenge of moving in response to climate change. Across most of the U.S., isotherms of temperature run basically east-west, with annual average temperature decreasing roughly 1°C per 100 kilometers northward movement. In areas with little topography, required movement velocities for staying in a climate with a constant annual average temperature are about 2 kilometers yr-1 for a +2°C world and 4 kilometers yr-1 for a +4°C world. In California, however, isotherms run basically north-south, with much steeper gradients (see Chapter 2, “Climate”). This reorientation of isotherms is a result partly of regional atmospheric circulation and partly of the strong effect of elevation on temperature. Roughly speaking, annual average temperature cools about 6.5°C for every 1 kilometer elevation with important local and weather-related variability (Lundquist and Cayan 2007). Because of the state’s generally steep gradients in existing temperature, the velocity of climate change in California will be less than in many other regions. Especially in the Sierras, shifting roughly 300 meters higher (for a +2°C world) or 600 metes higher (for a +4°C world) over the twenty-first century might require moving only a few tens of kilometers instead of a few hundred (Figure 14.6). Because warming on land is likely to be greater than the global average warming, the necessary velocity vertically or poleward will likely be greater than what is calculated based on global mean warming. California’s topographic relief increases the probability that some species will be able to track climate change. While it is likely that many species will be able to capitalize on this

topographic opportunity, many others probably will not, for four main reasons. First, maximum movement rates vary dramatically across taxa, with many rarely moving at rates as rapid as a few kilometers per decade (Settele et al. 2014). Second, many California ecosystems are structured around long-lived plants that do not reach maturity for a century or more. The concept of an ecosystem shifting is not meaningful in a time frame so compressed that the dominants do not provide the structure on which other members of the ecosystem depend. Third, the surface area in the state decreases rapidly with elevation, with suitable habitat for many species decreasing even more rapidly because the highest sites are often rocky and/or steep. Fourth, climate change might literally squeeze many species off the tops of the highest local mountains, especially in a +4°C world. The American pika (Ochotona princeps) is emblematic of the challenges faced by high-elevation species (Moritz et al. 2008). But in addition to the high-elevation specialists, California has many important ecosystems for which a big upward shift, especially a shift of 500 meters or more in a +4°C world, would push them either off the tops of the mountains or into a zone with little or no soil. This includes most places in the Coast Ranges, where total relief is on the order of 1,000 meters or less, as well as many iconic locations in the Sierras. While the velocity of climate change can say a lot about the rates of movement necessary for an organism to remain in a constant climate, it does not speak directly to the question of how far a species needs to move to persist. To answer that question, one needs to consider niche breadth, especially the extent to which species can persist in warmer habitats. Niche breadth can be addressed at the scale of the species, the plant functional type, or the biome. Based on pollen records, vegetation changes have generally tracked rates of climate change. Many studies have addressed this for the period following the last glacial maximum (LGM) (Pitelka et al. 1997). Even when climate zones shifted over thousands of kilometers, biomes tended to keep up. Still, historical velocities of climate change, since the LGM and at other periods in the last fifty-five million years of Earth’s history, have all been very slow in comparison to velocities over the past century and projected for the next (Diffenbaugh and Field 2013). In comparison to past periods at the global scale, anthropogenic warming is one to two orders of magnitude more rapid. Historical rates of range shift do not provide the necessary information to ask whether modern biomes may be able to track future changes in climate, particularly in the modern context of landscapes fragmented by human activity.

Species Ranges A large body of work on present and potential species ranges is based on statistical approaches that define a species niche based on relationships between presence/absence observations and a number of bioclimatic variables. BIOCLIM (Booth et al. 2013) and MaxEnt (Elith et al. 2011) are widely used examples of these bioclimate envelope models (BEMs). When combined with climate model output, these models can be used to estimate the velocity of climate change, the extent of overlap between current and future suitable habitats, or change in the size of suitable area for a species. All of these metrics are relevant not only to the presence/absence of individual species but also to future biodiversity (McGill 2010) and risk of extinction (Thomas et al. 2004). Climate Change Impacts   257

Velocity of climate change based on nearest equivalent temperature

RCP 8.5 2081-2100

0.5

1

2

4

8

16

32 64 128

km yr-1

Velocity of climate change based on present temperature gradients FIGURE 14.5 The velocity of climate change (km yr-1) as determined by two methods. Color scales are different in the two panels.

Top: Climate change velocity calculated by identifying for each grid point the closest location with a future annual temperature similar to the grid point’s baseline annual temperature, based on the CMIP5 RCP8.5 model ensemble and accounting for baseline-period noise. Source: Diffenbaugh and Field 2013. Bottom: Climate change velocity calculated from the ratio of the projected temporal gradient in temperature (from the CMIP3 ensemble in scenario A1B of the Special Report on Emissions Scenarios) and the present spatial gradient in temperature at each location, °C yr-1 ÷ °C km-1 = km yr-1. Source: Loarie et al. 2009.

SRES A1B 2050-2100

0.01

Loarie at al. (2008) combined climate envelope models with global climate model simulations to explore possible changes in the diversity, range size, and movement required of California endemic plants, a group that includes 2,387 taxa. Their conclusions about diversity depend critically on whether taxa can move with climate change. With a +4°C world and species that cannot move, plant biodiversity decreases by 27%. In a +2°C world where species can move, biodiversity increases, especially along the northern coasts. For all scenarios, biodiversity tends to decrease in the foothills of the southern Sierra and to increase toward the north and the coast. Changes in range size also depend on assumptions about whether species can move. In a scenario with a +4°C world and no movement, over 66% of endemics experienced range reductions of 80% or more. With the assumption that species can move, the center of the distribution shifts by an average of up to 151 kilo258  Biota

0.1

1.0

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km yr-1

meters, depending on the climate scenario. While most of the shifts are projected to be higher in elevation and northward, many of the range expansions are southward, reflecting the strong influence of elevation.

Biome Boundaries/Species Ranges Modern tools for predicting biome boundaries as a function of climate trace from at least the observations of von Humboldt and Bonpland (1807), who noted consistent relationships between climate and vegetation structure, even for areas separated by large distances with vegetation of distinct evolutionary histories and in similar climates even when they occurred at different elevations. Von Humboldt’s insights about rule-based biome maps directly support modern biome

10

0.1

Velocity (km yr -1)

1

0.01 100 km

FIGURE 14.6 The velocity of climate change across California, in an enlarged view of the bottom panel of Figure 14.5. Keeping pace with changing temperature entails greater velocities in more level areas such as the Central Valley and lower velocities on mountain slopes. For this region the temporal gradient in mean annual temperature increases with distance from the coast, whereas the spatial gradient is strongly affected by topography. Source: Loarie et al. 2009.

and dynamic global vegetation models (DGVMs), which have added gradually to the core of von Humboldt’s discovery. Most recent biome models add constraints based on carbon or water balance to the basic rules that set vegetation type as a function of climate. DGVMs are conceptually similar but performance is evaluated separately for several plant functional types, such that biomes are emergent results not specified in advance. All these approaches can produce reasonably accurate maps of the large-scale distribution of the world’s biomes, especially for potential natural vegetation, or ecosystems not shaped by human activities. All these approaches can also reproduce biomes of the past, especially the period since the last glacial maximum. For this period, modeled biomes can be validated by comparison with biomes reconstructed from palynological records. One consistent result from these validations is that biomes in the first few millennia following the LGM were similar to modern biomes in some respects but different in others. The biomes from the past tended to have now-absent species combinations that make them “no-analog” ecosystems in the modern biogeography (Williams and Jackson 2007). All these approaches, however, require additional information to generate reasonable biomes for some regions. The best-known example is the prairie peninsula (in the upper midwest of the U.S.), where climate conditions point to forest dominance but the pre-European vegetation was prairie. For these areas the inclusion of a process like grazing or wildfire in the model is critical for generating realistic biomes. California biomes in a +4°C world of the late twenty-first century, simulated with the MAPSS-CENTURY 1 model (MC1), maintain surprising consistency with the present distribution (Figure 14.7). The simulation indicates substantial losses in alpine/subalpine forests, mixed evergreen woodlands, shrublands and deserts, with increases in mixed evergreen forest and grassland (Figure 14.8). The finding that

high-elevation ecosystems tend to be pushed off the tops of the mountains is very widespread. It is one of the reasons that alpine habitats and species tend to be among the state’s and the world’s most endangered. In the simulation of Lenihan et al. (2003), the decrease in shrublands and mixed evergreen forests is compensated largely by an increase in grasslands, mainly as a consequence of greater water limitation in the coastal mountains and Sierra foothills. The replacement of conifer-dominated forest by broadleaved-dominated forest over much of the northern part of the state is a feature of the simulations with wetter climates. MC1 simulates the locations of biomes as units, without considering the role of climate change in shaping the particular species composition. Empirical data from Kelly and Goulden (2008) indicate that this is likely not the general pattern, but that species mixtures change individualistically, with some species adjusting quickly to climate signals while other adjust slowly. Biome models that generate biomes from a mixture of plant functional types potentially have the ability to track some kinds of compositional changes, especially when these changes are regulated by environmental and not biotic controls. A few DGVMs (e.g., Moorcroft et al. 2001) simulate some aspects of biotic interactions but mainly those related to competition for light, water, and nutrients.

Interacting Factors: Climate Change as a Threat Multiplier In general, the tools for estimating the boundaries of biomes or species distributions account for climate and atmospheric CO2 . They sometimes account for soil characteristics and atmospheric pollution. They also typically assume (1) universal availability of all potentially viable taxa, (2) lack of complication from pests and pathogens, and (3) a crisp partitioning of the landscape into human-dominated and natural units. Calibrations and validations are grounded in current or past patterns, where the role of interacting factors including other environmental changes and human impacts is either minimal or poorly known. In a world that is increasingly dominated by these interacting factors, how should approaches to defining biome or species boundaries change? Sala et al. (2000) attempted to account for impacts on biodiversity in 2100 due to land use, atmospheric CO2, nitrogen deposition, and biological invasions in addition to climate change. They concluded that the role of each factor varies among ecosystems, with land use the most important at the global scale. In Mediterranean-climate regions they identified land use and biotic exchange (invasions) as equally important. In general, implications of interacting factors are not accounted in projections of future biome or species distributions under climate changes. For a number of reasons, we feel that for the ecosystems of California through the twenty-first century, interacting factors are likely to be the core story and not just minor complications. We envision the process of creating new distribution boundaries as functioning like a series of environmental sieves. In the past, changes in climate were slow enough that migration distances or broken mutualistic relationships were generally not problems (see mutualism). But in the future that will not be the case. As with most processes in ecology, species responses to interacting factors will likely play out with individualistic differences among species. Some species will be relatively good at Climate Change Impacts   259

Alpine/subalpine forest Evergreen conifer forest Mixed evergreen forest Mixed evergreen woodland Grassland Shrubland Desert

A

B

FIGURE 14.7 Vegetation distribution in California during the late twentieth century and projected for the future. Source: Lenihan

et al. 2003. A 1961–​1990 baseline period map in which twenty-eight California vegetation types mapped by Küchler (1975) have been

aggregated into the seven vegetation classes simulated by a dynamic vegetation model, MAPSS-CENTURY 1 (MC1). B 2070–​2 099 projection for the vegetation using MC1 with the Hadley Climate Center HADCM2 climate predictions.

capitalizing on the ecological opportunities of the Anthropocene, and others will be relatively bad. In general, we anticipate that the characteristics that best suit species for success in the rapidly changing context of California’s next century will be the characteristics that make plants successful as weeds. Facile long-distance dispersal, ability to colonize disturbed areas, lack of dependence on coevolved mutualists, and ability to spring back after extreme events will be the most essential characteristics of successful species. In addition, characteristics that make plants friendly to assisted management from humans might play a big role in future success.

Biotic Interactions Biotic interactions are among the crowning glories of evolution. Natural selection has shaped myriad actors in diverse interactions that range from competition for light and water to pollination, herbivory, predation, nitrogen fixation, pathogenesis, and many more. The defining feature of a biotic interaction is that more than one species is involved. In real ecosystems, biotic interactions link all species to some degree. Decreased growth in one plant might increase availability of light, facilitating growth in another. Establishment of a new disease in an insect or vertebrate predator could increase the abundance of herbivores, decreasing the abundance of their food plants. Increased temperature might stimulate biological 260  Biota

nitrogen fixation, facilitating the invasion of more nitrogendemanding species. These and other indirect effects might unfold over very different time periods (Van der Putten et al. 2010). Climate change and biotic interactions is an area of expanding work on specific interactions, but the search for general principles is still in its early stages. The potential importance of understanding this area is profound, but research is challenging. Integrative studies to date have focused mainly on opening doors to understanding possibilities. For example, Aslan and colleagues (2013) combined information on at-risk vertebrates and plants pollinated or dispersed by vertebrates to estimate the global threat to plant diversity from future losses of vertebrate species. Their study, while addressing important issues, did not look at herbivory or effects of altered competition from other plants. For trout in North America, future changes in distribution appear sensitive to temperature but also to competition with other trout, changes in the flow regime, and changes in food availability (Wenger et al. 2011).

Invasive Species Invasive species are major components of the flora and fauna of California (see Chapter 13, “Biological Invasions”). Characteristics of successful invaders have been difficult to define with any precision, but they are relatively simple to define

HAD PCM

Alpine/subalpine forest Evergreen conifer forest Mixed evergreen forest Mixed evergreen woodland Grassland Shrubland Desert -100

0

100

200

Change in total cover (%) FIGURE 14.8 Comparison of percentage changes in the total cover of vegetation classes from the 1961–​1990 baseline period to the 2070–​ 2099 period as projected by the MC1 vegetation model under climate scenarios by the Hadley Climate Center HADCM2 model (HAD) and the National Center for Atmospheric Research’s Parallel Climate Model (PCM). Source: Lenihan et al. 2003.

in terms of outcomes. Invaders tend to be good at getting to new locations, establishing quickly—​especially in disturbed sites—​a nd persisting in the face of environmental variation once established. Under any scenario of future climate change, California will see myriad instances of extant species stressed by novel conditions. This stress could lead to decreased vigor, decreased abundance or mortality, creating opportunities for other species. Though knowledge is far from complete, it is hard to imagine that invaders will not be common among the plants and animals most likely to capitalize on these opportunities. The juxtaposition of stress from climate change and the existence of a large pool of species good at moving to and establishing in new places seems like a recipe for creating a future “Homogenocene” (Samways 1999), with fewer and fewer species that are narrowly restricted or poor at relocating and greater abundances of generalist species that are good at moving. Many examples document invasions mediated by climate change (Walther et al. 2009). In some examples, climatemediated pressure to move blurs the distinction between a range shift and an invasion. In others, the winners are cosmopolitan invaders. For California, which is already so wellendowed with invaders, it will be surprising if climate change does not eventually yield ecosystems more weedy and more dominated by cosmopolitan invaders. The species that are already problem weeds are likely to be among the winners, but climate change could expand opportunities for other exotic species as well. In general, species rearrangements in a changing climate are likely to unfold like a massive game of musical chairs. When the music stops, it could be all about which species are best at scrambling for the remaining chairs.

Biodiversity Estimating impacts of climate change on biological diversity is one of the key research challenges of the twenty-first century. The topic is central for understanding the consequences of climate change, but it is also incredibly difficult

because it involves the interaction of many different kinds of factors beyond climate change, the potential for important roles of a wide suite of both direct and indirect effects, and a strong likelihood that processes will differ in importance on a range of time scales. Efforts to estimate the number or fraction of species at risk have taken a number of approaches (Bellard et al. 2012). Bioclimatic envelope models (BEMs), dynamic global vegetation models (DGVMs), and species area relationships (SARs) all define new ranges based on a new climate and work from the new ranges to estimate biodiversity. Approaches based on IUCN status or dose-response relationships attempt to characterize the level of local threat (Bellard et al. 2012). All these approaches provide important insights, but they fail to account for a broad range of potentially important factors, especially factors like the breaking of pollinator and disperser relationships (Aslan et al. 2013). Barnosky and colleagues (2011) discuss current and potential future extinctions in the context of the five mass extinctions in Earth’s history, times when the planet lost more than 75% of its species. While the list of known extinctions since the start of the anthropogenic era is still seemingly short, many species are already threatened. Comparing past mass extinctions with the cast of threatened species leads to the sobering conclusion that the loss of only species that are already “critically endangered” could make the Anthropocene a sixth mass extinction. Losses of species that are “endangered” or “vulnerable” could precipitate this outcome in only a few hundred years. None of the existing methods for estimating future changes in biodiversity produces a result that should appropriately be called a conclusion. Each method provides a way to think through the implications of a particular mechanism or set of interactions. But the actual trajectory of biodiversity will be determined by a much broader set of interacting mechanisms, with the potential for many unique aspects of each local setting. Vertical relief and refugia could serve as effective buffers against biodiversity loss. Habitat destruction, air and water pollution, and large numbers of invasions could accelerate losses. The net outcome for a region like California is far from clear. Still, it is difficult not to sense the risk of a major disruption arising from these multifactor interactions. As a stress multiplier, climate change has the potential to push a challenging situation into crisis. The temporal pattern of future changes in biodiversity is exceedingly difficult to predict. During a reshuffling, species diversity could move smoothly to a new steady state, but it could also overshoot, lag, or move initially in the opposite direction from the long-term pattern. In addition, the concept of anything approaching a steady-state could be irrelevant in the future, where changes in climate, land use, sea level, atmospheric composition, and invasions could continue well beyond the end of the twenty-first century. Even if these exogenous factors stop changing, internal ecosystem processes could lead to changes that continue to unfold over many centuries.

Managing California Ecosystems in the Changing Future For both the era of committed climate changes and the longer-term era of climate options, climate change is a challenge of managing risks. A range of outcomes is possible, with conClimate Change Impacts   261

sequences that depend on the outcome. Even for a particular mean temperature the distribution of actual conditions around that mean, including extreme events, is difficult to predict. Response uncertainties, interactions, and modulation by other anthropogenic factors all add to the range of possible ecosystem outcomes. Effective approaches to managing risks need to consider the full range of possible outcomes as well as the probability of each outcome. In general, risk scales approximately with the product of consequence and probability, meaning that a low-probability, high-consequence outcome can present risk that is comparable to a modest-probability, modest-consequence outcome. California has warmed substantially over the past century, a period in which rapid changes in land use and other human activities were almost certainly the dominant drivers of ecosystem change. Over the twenty-first century, ongoing effects of land use change will doubtless continue to be important. If climate change is managed such that we stabilize temperatures in the range of 2°C above preindustrial levels, then climate change will have pervasive impacts on ecosystems. But there is a realistic chance that biome boundaries, biodiversity, and ecosystem function will change gradually and incrementally, likely toward a weedier world. If greenhouse gas emissions continue to grow rapidly and warming in the twentyfirst century is on the order of 4°C, then the picture looks very different. The +4°C world is so far outside Earth’s geologically recent, normal operating parameters that the main message from the science is that all bets are off. We could get lucky, with incremental changes, but we could also see massive, discontinuous impacts. As risk managers, avoiding this +4°C world is a top priority.

Summary Climate change has had and will continue to have a wide range of direct and indirect effects on California’s diverse ecosystems. This chapter focused on the ecological effects of climate change in the state to date and in the future. California has experienced an increase in mean annual temperature since the beginning of the twentieth century of approximately 0.9°C, close to the global average. Many cases exist of well-documented responses to this past warming. These cases most often compare current species phenologies, distributions, abundances, and physiologies with historical data. Collectively they show a discernable pattern of ecological response to climate change across the state including elevational and latitudinal shifts in terrestrial and marine environments as well as changes in migration timing, species performance, and other biotic responses. Studies of the interactions of climate change and other stressors including atmospheric and land use change and biological invasions remain relatively scarce, as do studies focused on the effects of climate extremes rather than means. Future climate change impacts in the state depend on many unknowns, especially on what emissions trajectory the world follows in the next century. Even with ambitious mitigation, the world at the end of the twenty-first century will be about +2°C warmer than preindustrial levels, with widespread ecological impacts. Continuing high emissions could yield a +4°C future for California by the end of the twentyfirst century. At this level, all bets are off as to how ecological systems will respond. Rates of change are one to two orders of magnitude higher than in past periods of climate change, 262  Biota

raising the question of whether and how species can respond effectively. California’s strong topographic relief and climate gradients from coast to inland mean that species can track shifting climate spaces over relatively short distances. However, varying mobility across taxa, long lags to maturity in many plants, declining available area and soil development with increasing elevation, and local limits to upslope movement such as in the Coast Ranges could all pose barriers to effective responses by native biota. Moreover, many of the species best equipped to respond rapidly to change are California’s weedy invasive species, which could benefit from rapid climate changes at the expense of native biodiversity. Biodiversity changes in California in response to climate change are difficult to predict, but climate change will likely emerge as an increasingly strong threat multiplier for species already stressed by forces like land use change and invasions. The highest priority for managing risks to California’s ecosystems in the long term is to reduce emissions and avoid a +4°C future.

Acknowledgments Thanks to Mike Goulden, Anne Kelly, Ron Neilson, Terry Root, and Matthew Forister for checking our interpretation of their published datasets. Thanks to Leslie White and Geert Jan von Oldenburgh for help with Figure 14.2. We acknowledge support from the National Science Foundation program in Ecosystem Studies (Award # 0918617) and the NSF CAREER program via the Division of Atmospheric and Geospace Sciences (Award # 0955283). We acknowledge the World Climate Research Programme’s Working Group on Coupled Modelling, which is responsible for CMIP, and we thank the climate modeling groups for producing and  making available their model output. For CMIP the U.S. Department of Energy’s Program for Climate Model Diagnosis and Intercomparison provides coordinating support and led development of software infrastructure in partnership with the Global Organization for Earth System Science Portals.

Recommended Reading IPCC. 2014. Climate change 2014: Impacts, adaptation, and vulnerability. Contribution of Working Group II to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, UK, and New York, New York. Kadir, T., L. Mazur, C. Milanes, and K. Randles. 2013. Indicators of climate change in California. California Environmental Protection Agency, Sacramento, California. 258 pp. National Research Council. 2010. Advancing the science of climate change. National Academies Press, Washington, D.C.

Glossary Anthropocene  A proposed term for a geological epoch extending to the present from the beginning of the period of significant human impact on the world’s ecosystems. Assemblage  A group of fossil organisms found together, indicating co-occurrence during a given period of prehistory. Biomes  The world’s major ecological communities, defined by major vegetation type (such as forest, desert) and by organismal adaptations to a particular climate and environmental setting.

Edaphic endemic  An organism restricted in distribution to a particular, restricted soil type. Glacials (versus interglacials)  A geologic period of colder temperatures that alternated with warmer, or interglacial, periods during an ice age. Isotherm  A contour line on a map indicating areas with similar temperature. Krummholz  Describes a stunted form of trees at high elevations exposed to strong winds and freezing temperatures. Lapse rate  The rate at which atmospheric temperature declines with increasing altitude. Mitigation  Here, action to reduce the rate of increase of atmospheric greenhouse gas concentrations, through reduced emissions and/or increased sequestration of greenhouse gases. Mutualism  A relationship between two organisms of different species in which each benefits from the other. Palynological  Based on the study of pollen or pollen records. Pathogenesis  The biological mechanisms underlying the development of disease. Phenology  The timing of periodic life events in organisms, such as leaf drop and budburst in deciduous trees. Planktonic foraminifera (in sedimentary assemblages)  Single-celled organisms in the water column belonging to a particular class, which accumulate in bottom sediments as they rain down over time. Refugia  A location where an isolated population of a oncewidespread taxon persists.

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FIF TEEN

Introduction to Concepts of Biodiversity, Ecosystem Functioning, Ecosystem Services, and Natural Capital REBECCA CHAPLIN-KR AM ER , LISA MANDLE , ELIZ ABE TH R AUER , and SUZ ANNE L ANGR IDGE

Introduction Ecosystems, the diversity of life within them, and their basic ecological processes support and enhance human life. We explore the relationship between humans and nature from basic principles of the ecology underlying that relationship to their implications for societal decisions. We hope this chapter provides the reader with a sense of the interconnectedness between human and natural systems, and of why it is important to understand nature’s benefits to people from a variety of angles.

Biodiversity: The Variety of Life on Earth Facets of Biodiversity Biological diversity—​or biodiversity—​is the variety of life across all ecological levels, from genes to species to ecosystems. Biodiversity includes everything from the vast array of life on land and in the seas to the foods we eat and the landscapes in which we live and visit. With so many dimensions

to biodiversity (see, for example, Gaston 1996, Wilson 2010), scientists have recognized several aspects, including taxonomic, genetic, phylogenetic, and functional diversity. The spatial partitioning of diversity across heterogeneous landscapes is another important dimension of biodiversity. Species richness, a simple measure of the number of species in a given area, is one of the most common metrics for biodiversity, and part of the basis for the biodiversity hotspot concept that highlights highly altered regions with the highest numbers of unique species (Myers et al. 2000). Species richness provides one measure of taxonomic diversity, or the diversity of taxonomic units, which can also include genera, families, and orders. A variety of diversity indices have been developed to incorporate the numbers of species as well as their relative abundances within a community, in order to differentiate between communities with the same number of species but differences across dominance or evenness of the species within them. However, despite its ease of application, taxonomic diversity alone provides an incomplete picture of the multidimensional nature of the world’s biodiversity. 265

Genetic diversity, the genetic variation within and between species, is critical to a species’ ability to withstand and adapt to change and thus underpins all other forms of biodiversity. Species with low genetic diversity may have greater vulnerability to parasites and infectious diseases (O’Brien and Evermann 1988), and low adaptive potential, which means greater vulnerability to a changing environment. The California condor (Gymnogyps californianus) was reduced to fewer than twenty-five individuals left in the wild, mainly due to habitat loss and poisoning (Kiff et al. 1996). Today, nearly four hundred condors exist, with over two hundred in the wild (California Condor Recovery Program 2013). However, because they all descended from the same small group of individuals, genetic diversity is low, contributing to the continued threat of extinction for this species (Ralls and Ballou 2004). Phylogenetic diversity is a measure of the difference in evolutionary history across species. Phylogenetic diversity is often used in cases where taxonomic units are not well-defined or described, such as in the case of bacteria. However, even where taxonomic diversity is easy to measure, phylogenetic diversity can provide additional useful information. Measures of species diversity treat any two species as equally different from each other, while phylogenetic diversity provides a measure of the evolutionary distance between species. A correlation between species diversity and phylogenetic diversity has been demonstrated across California grassland plant communities, but a particular community’s rank in diversity depends on which characteristic is used (Cadotte et al. 2010). The difference between phylogenetic and genetic diversity is that phylogenetic diversity excludes shared genetic differences among species. Phylogenetic diversity therefore emphasizes clades and species that have relatively more unique diversity and underemphasizes species in clades that are species rich and radiating rapidly. These differing definitions of biodiversity have important implications for priority-setting in conservation and the designation of protected areas (Devictor et al. 2010). Functional diversity is the range of functional traits, characteristics that influence ecological functions, and processes present in a system. Functional traits of an organism can include everything from seed or egg size, to metabolic and photosynthetic rates, to trophic level (e.g., McGill et al. 2006). Functional traits are often divided into two categories: effect traits, which determine an organism’s effect on ecosystem processes such as nutrient cycling, and response traits, which govern an organism’s response to environmental conditions, such as resource availability or disturbance (Lavorel and Garnier 2002). Some traits, such as an animal’s body size or a plant’s leaf nitrogen content, may relate to both an organism’s effect on and response to its environment. Ultimately, an ecosystem’s response to environmental change will be influenced by how the composition of the ecological community responds to change and what the effect of the resulting community is on ecological processes.

Photo on previous page: Laguna Creek, California wetland in first light. Coastal marshes like this one can protect surrounding areas from flooding and erosion caused by storm surge and act as filters of agricultural and industrial run-off to the sea, securing and enhancing services like recreation and fishing provided by healthy coastal habitats. California’s size and diversity of ecosystems allow the provisioning of many different ecosystem services, including the supply and purification of water from the mountains to cities downstream, forest carbon sequestration and timber production, a wide array of agricultural systems and the services like pollination and pest control that support them, and myriad recreational opportunities. Photo by Ed Dickie. 266  Biota

Some conservation efforts have undergone a recent shift in focus from taxonomic diversity towards functional diversity, in part because functional diversity can provide a mechanistic link between biodiversity and ecological processes (Cadotte et al. 2011). There is frequently a lack of correspondence between changes in functional diversity and changes in species richness, as drivers of environmental change can affect the two different aspects of biodiversity independently (Mayfield et al. 2010). However, a continuing practical challenge when measuring functional diversity is identifying the functional traits that are truly important to ecological processes. Phylogenetic diversity is sometimes assumed to be a proxy for functional diversity and can potentially be a better predictor of ecosystem function than metrics of functional diversity, because it may capture variation in traits that has not yet been recognized or measured (Cadotte et al. 2009). On the other hand, phylogenetic distances are not always well correlated with functional differences, so reductions in phylogenetic diversity may underestimate losses in functional diversity (Prinzing et al. 2008, Fritz and Purvis 2010). Another important aspect of biodiversity is its spatial partitioning across the landscape (see, for example, Magurran 1988). The diversity found in a particular ecosystem—​which can be quantified in terms of species richness, functional diversity, or any of the other previously described measures of diversity—​is called alpha diversity. Landscapes can include a diversity of ecosystem types or habitats, and the total species diversity found across this larger region is referred to as gamma diversity. Beta diversity is the difference in composition across ecosystems within a region and provides the link between alpha diversity at the local and gamma diversity at the regional scale. Landscapes with high turnover in species composition or diversity across ecosystem types have high levels of beta diversity, while more homogeneous landscapes have low beta diversity.

Levels and Losses of Biodiversity Taxonomists have already identified nearly two million species (Chapman 2009), but we know this represents only a fraction of the total biodiversity on the planet. There are substantial biases in the kinds of organisms that have been cataloged so far; conspicuous and easily accessible organisms like mammals and plants are far better studied and understood than those that are harder to see and find, such as fungi, insects, and other invertebrates (Scheffers et al. 2012). The rate of discovery of new species remains high, at approximately 15,000 species per year (May 2011). Recent estimates of global biodiversity suggest that the planet currently supports 5 million to 10 million eukaryotic species (Mora et al. 2011), although the figure could be substantially higher. Within the United States, California leads in species richness, with more than 5,500 vascular plant, vertebrate animal, and freshwater mussel and crayfish species (Stein et al. 2000). Other taxa such as insects, ferns, or fungi have not been cataloged thoroughly enough to allow for a comparison across states. California also leads in number of species unique to this state, with 1,500 endemic species described. As of 2013, 281 plant and 157 animal species were state or federally listed as rare, threatened, or endangered in California (CDFW 2013a, 2013b). By one estimate, 78% of California’s 129 native freshwater fishes are at risk of becoming extinct, with 5% (7 species) already extinct (Moyle et al. 2011).

Why is the risk of species extinctions of such concern? The fossil record makes clear that extinction is a natural process that has been occurring since long before humans arrived on the scene. Today’s rates of extinction, however, are a hundred to a thousand times higher than background natural rates (Millennium Ecosystem Assessment 2005). At five previous times in geologic history, more than 75% of species in existence went extinct (Jablonski 1994)—​these have been dubbed the “Big Five” mass extinction events. By some estimates, current rates of extinction are at the extreme high end of the range of extinction rates in the “Big Five” (Barnosky et al. 2011). If the present pace of extinction continues, we could reach a sixth mass extinction event in the coming centuries. Even before species become globally extinct, their disappearance locally can have important consequences for ecosystem processes and the ecosystem services that are critical to our own well-being. The consequences of biodiversity loss for ecosystem functioning—​both generally and within California—​are explored next, while further discussions of biodiversity loss and its relationship to other forms of environmental change in California’s ecosystems can be found in other chapters in this volume.

Ecosystem Functions Introduction to Ecosystem Ecology Ecosystem structure is the temporal and spatial patterns in the diversity and composition of the elements (both living organisms and abiotic environment) of an ecosystem and the relationships between them (Odum 1971). This structure influences and is influenced by individual species traits such as resource use, trophic dynamics, and contribution to disturbance (Chapin et al. 1997) and can also be described by each species’ relative contribution to ecosystem function (Balvanera et al. 2005). The relationship between ecosystem structure and function is thus complex and bidirectional, with functional processes feeding back onto the structural elements of an ecosystem, but ecosystem function itself is determined by the identity, relative abundance, diversity of and relationships between the species that comprise the ecosystem. Ecosystem functions are ecological processes, such as primary production, nutrient cycling, and decomposition, that control the flows of energy, nutrients, and organic matter through an environment (Cardinale et al. 2012). While ecosystem function has also been used to refer to the subset of ecological processes that can directly or indirectly fulfill human needs (De Groot 1987, De Groot et al. 2002), we use the more encompassing definition to differentiate between ecosystem functions and ecosystem services.

Biodiversity and Ecosystem Functions Determining how biodiversity affects ecosystem functions is critical to understanding the consequences of local and global species extinctions for humanity. The relationship between biodiversity and ecosystem functions emerged as a focus of ecological inquiry in the 1990s, but it is only within the past decade that enough information has been amassed to begin to determine general relationships between biodiversity and ecosystem function (Hooper et al. 2005, Balvanera et al. 2006, Cardinale et al. 2012, Naeem et al. 2012). To date, our

understanding of the biodiversity-ecosystem function relationship comes primarily from studies in temperate grassland systems, as the relatively short lifespans of the plants and the moderate levels of diversity found at small scales make these systems particularly suitable for experimental studies. Several well-studied California ecosystems have contributed to this understanding, particularly the state’s serpentine grasslands and annual-dominated grasslands (Hooper and Vitousek 1997, Dukes 2001, Lyons and Schwartz 2001, Zavaleta and Hulvey 2004, Selmants et al. 2012). California’s kelp forests and rocky intertidal zones have also provided a model system for deepening our understanding of the role of biodiversity in the marine realm (Dayton and Tegner 1992, Sala and Knowlton 2006). In experimental studies, increasing species richness generally increases several key ecosystem functions, including biomass production, resource capture, and nutrient cycling efficiency (Cardinale et al. 2012). The mechanisms by which this occurs have been termed the “selection effect” and “complementarity effect” (Loreau and Hector 2001). The selection effect occurs when the species pool from which communities are assembled contains some species that outperform others in terms of a particular ecosystem function. When communities are assembled by randomly drawing species from a pool, the more species that are included in the community, the higher the likelihood that the community will contain and become dominated by a top performer, resulting in a positive biodiversity-ecosystem function relationship. The complementarity effect can produce the same pattern through an entirely different process. When species in a community facilitate each other or draw on different resources, this complementarity among species can lead species-rich communities to outperform species-poor communities. Both the selection effect and complementarity effect contribute to the observed increases in ecosystem functioning with increasing biodiversity. Biodiversity also promotes the stability and resilience of ecosystem functions in the face of environmental change in both terrestrial and aquatic systems (Cardinale et al. 2012, 2013). For example, in California’s Sierra Nevada, conifer forests show greater resilience to drought with increasing species richness, returning more quickly to pre-drought levels of stand productivity (DeClerck et al. 2006). Variation in the magnitude and timing of individual species’ responses to environmental change can promote stability of communityand ecosystem-level properties. Ecological theory points to a number of possible mechanisms at the root of this pattern. Statistical averaging, also called portfolio effect, results increased stability with increasing diversity when fluctuations in species abundances through time are not perfectly correlated (Doak et al. 1998). Stability can also result from compensatory dynamics, which occurs when declines in abundance of some species are coupled with increases in other functionally similar species, such as through competitive release (Tilman 1996). As communities increase in diversity, there is a greater likelihood that sets of functionally similar species will be present. Overyielding is a third mechanism that can increase stability. When complementarity among species increases a community property such as biomass, the relative amount of variation in the system decreases, leading to greater stability with higher diversity (Lehman and Tilman 2000). More research is needed to understand the relative importance of these mechanisms (Cardinale et al. 2012). The positive relationship between levels of biodiversity and Biodiversit y and Ecosystem Services   267

Ecosystem function

Increasing realism

Biodiversity FIGURE 15.1 Ecosystem function (e.g., net primary production, nutrient cycling, decomposition) typically increases with biodiversity (e.g., species richness, number of functional types) in a saturating relationship, where at a certain point additional biodiversity does not continue to increased ecosystem function. The saturation point depends on the level of realism incorporated into studies of this relationship; as observations are made over longer time scales, at larger spatial scales, and for multiple ecosystem functions, the observed relationship between biodiversity and ecosystem function will shift from a steeper, more immediately saturating curve (represented by lighter gray lines) to a more gradual increase saturating at higher levels of biodiversity (represented by darker gray lines).

an ecosystem function is frequently not linear but instead shows saturating effects (Cardinale et al. 2006). Starting from low levels of biodiversity, increasing diversity often increases a particular ecosystem function only up to a point. Beyond that, additional biodiversity may do little to increase a given function. This pattern can result from functional redundancy within a community: multiple species may contribute to a particular ecosystem function in similar ways, so the addition or loss of another functionally similar species may have little effect (Lawton and Brown 1993). Saturating effects of increasing biodiversity suggest that a low level of species losses will have only minimal consequences for an ecosystem function, if those species that are extirpated are functionally redundant (Cardinale et al. 2011). However, recent studies provide evidence that the saturating relationship between biodiversity and ecosystem function diminishes as ecosystem function is observed over longer time scales, at larger spatial scales, and when considering multiple ecosystem functions (Figure 15.1; Hector and Bagchi 2007, Zavaleta et al. 2010, Isbell et al. 2011, Reich et al. 2012, Pasari et al. 2013). Although these studies focus exclusively on temperate grassland ecosystems, some studies of seagrass systems (Gamfeldt et al. 2008) suggest that these patterns may apply more generally, both in terrestrial and marine systems. Species that appear functionally similar in their effect on a single ecosystem function and under one set of environmental conditions may not perform similarly for all functions and under all conditions. Maintaining a suite of ecological functions, especially in the face of environmental change, requires greater levels of biodiversity than would be concluded from considering only a single function over short time scales, which unfortunately has been the scope of most ecological studies investigating the biodiversity-ecosystem function relationship. High-intensity agriculture illustrates this principle well—​in selecting 268  Biota

crops that maximize a single function (e.g., biomass production of a particular crop), we have created systems that often require high inputs of fertilizers and pesticides, and are susceptible to disease, pests, and drought. The short-term, smaller-scale studies demonstrating saturating effects often assume that species losses will occur randomly. In reality, extinction is not a random process and functionally similar species are likely to disappear from a community at the same time (Zavaleta et al. 2009). Studies of the effects of land use intensification on animal diversity—​ including birds, mammals, amphibians, and arthropods—​ have found that functional diversity often declines more rapidly than species richness, indicating that land use change did not cause random extinctions but instead eliminated functionally similar species (Ernst et al. 2006, Schweiger et al. 2007, Flynn et al. 2009). In California’s native-dominated serpentine grasslands where early-season annuals were most likely to be extirpated, and in its annual-dominated grasslands where perennials and late-season annuals were most at risk, both communities became more susceptible to invasion by exotic plant species when species losses followed realistic patterns (Zavaleta and Hulvey 2004, Selmants et al. 2012). A recent synthesis of approximately two hundred studies found that the effects of species losses on productivity and decomposition are as significant as the effects of climate warming, increased CO2, nutrient pollution, and other drivers of global change (Hooper et al. 2012). Conservation of biodiversity is therefore important for maintaining the diversity of ecological processes on which human welfare depends through time and in the face of environmental change.

Ecological Interactions Ecosystem functions depend not just on the identity and diversity of species within an ecosystem but also on interactions among species. Within communities, the presence or absence of some species have large effects on community structure, composition, and community; these species have been termed strong interactors (MacArthur 1972). Keystone species and ecosystem engineers are two types of strong interactors. Keystone species are those species that have a large effect on community structure and function, compared to their biomass within the system. The keystone species concept was put forth by Paine (1969), based in part on his observations of the role of starfish (Pisaster ochraceus) in structuring the Pacific rocky intertidal community. Sea otters (Enhydra lutris) along the Pacific coast are another classic example of a keystone species. They regulate the structure of kelp forest (Macrocystis sp.) communities through a trophic cascade: by preying on sea urchins (Strongylocentrotus sp.), which in turn graze kelp, sea otters maintain the diversity and structural complexity of kelp forest communities (Estes and Palmisano 1974). When sea otters are eliminated—​as occurred in throughout much of their range in the eighteenth and nineteenth centuries due to overhunting for the fur trade—​u rchin populations increase, leading to overgrazing of algae and even shifts from kelp forest to urchin barrens. While keystone species shape ecosystem functioning through trophic cascades, ecosystem engineers exert their influence on ecosystem structure and functions through the physical formation or modification of habitat (Jones et al. 1994). In California, both the native and invasive cordgrass

(Spartina spp.) act as ecosystem engineers by trapping sediment and providing structure (Brusati and Grosholz 2006). However, native cordgrass (S. foliosa) increases benthic invertebrate densities relative to uncolonized mudflats, whereas the invasive cordgrass (a hybrid, S. alterniflora × S. foliosa) reaches such high densities aboveground and belowground that it excludes invertebrates. Strong interactors such as keystone species and ecosystem engineers were an early focus of ecological research. However, the majority of species in a given community are considered weak interactors, with their presence or absence in a community usually having small effects. This does not mean that weak interactions are without ecological importance. Weak interactions have been found to promote stability in productivity and community composition (e.g., McCann et al. 1998, Neutel et al. 2002). In addition, a species’ status as a weak interactor is not fixed—​species that have little effect on average or under certain conditions can play a large role in structuring communities under other circumstances (Berlow 1999). Finally, although weak interactors may have small effects on a per capita basis, at high densities their populationlevel effects may be substantial. Returning to California’s kelp forests, relative to sea urchins, amphipods (Ampithoe humeralis) have weak per capita interactions with kelp, in part due to their small individual size (Sala and Graham 2002). However, during El Niño events, amphipods can reach high densities and in these circumstances play a significant role in structuring kelp communities (Sala and Sugihara 2005).

Ecosystem Services Ecosystem Services and Natural Capital The Millennium Ecosystem Assessment (MA 2003, 2005) succinctly defined ecosystem services as the benefits people obtain from ecosystems, combining two prior definitions focused on including both natural and human-modified ecosystems as sources of ecosystem services (Costanza et al. 1997) and both the tangible and intangible benefits that comprise ecosystem services (Daily 1997) to highlight how ecosystems sustain and fulfill human life. The value of ecosystem services can be considered a flow of benefits from nature to people. The stock of these benefits is what we call natural capital (Tallis, Polasky et al. 2012). The difference between these two is akin to the difference become income and wealth—​t hey are often positively correlated, and while poor management of a capital asset may produce short-term gains in income, long-term degradation of that asset will ultimately erode the benefits it can provide. Natural capital, as part of a broader economic framework including four other types of capital (manufactured, human, social, and financial), has gained more traction than ecosystem services in certain applications, particularly in national accounting and inclusive wealth accounting (Tallis, Polasky et al. 2012). However, recent polling has determined that the terms “natural capital” and “ecosystem services” are unpreferred by and unfamiliar to the U.S. populace, especially in comparison to more straightforward terms like “nature’s benefits” or “nature’s value” (Fairbank et al. 2010). Regardless of the terms chosen, these stocks and flows of ecosystem assets are often related to biodiversity for many of the same reasons as for ecosystem function. Research on the relationships between biodiversity and ecosystem func-

tion and biodiversity and ecosystem services, respectively, has involved quite different methods of inquiry, with the former employing manipulative (often small-scale) experiments and mathematical theory, and the latter consisting of mostly correlative, landscape-scale studies comparing major habitat modifications (Cardinale et al. 2012). In the smaller-scale controlled studies, a mixed or negative association has been shown between biodiversity and certain ecosystem services, especially for food production and disease control, where more species richness can detrimental to the ecosystem service over the short term. Still, mutually beneficial or win-win outcomes have been documented for biodiversity and many ecosystem services over the longer-term, especially in cases where the stock of natural capital is more valuable than the consumption or extraction of the flow of ecosystem services from that stock (Reyers et al. 2012).

Distinction between Ecosystem Function and ­Ecosystem Services Ecosystem functions support ecosystem services, and in common parlance these two terms are often used interchangeably, but they are not the same thing. The MA framework includes four types of services: provisioning (such as food, fuel, and fiber), regulating (such as air and water purification, climate stabilization, and hazard mitigation), cultural (such as spiritual, aesthetic, and educational enjoyment of nature), and supporting (such as soil formation, primary production, and nutrient cycling). By these definitions ecosystem functions can be considered supporting services, the processes necessary for the production of all other services, and indeed the MA suggests that supporting services differ from the other services in that “their impact on people are either indirect or occur over a very long time.” This definition can lead to confusion or misinterpretation of the ways that we measure and communicate nature’s value to people. For example, according to this framework, pollination is defined as a regulating service and soil formation as a supporting service, though both can be thought of as contributing to crop production and therefore to human nutrition and food security. Furthermore, it is acknowledged that some processes could be classified as either a supporting or a regulating service, depending on the time scale and immediacy of their impact on people. To address this confusion, and to make a clearer distinction between supporting ecosystem functions and ecosystem services, we consider the flow of benefits from ecosystems to humans as a supply chain: the supply of ecosystem services, the ecosystem services themselves, and the value of those services to people (Tallis, Lester et al. 2012; Tallis, Mooney et al. 2012).

Supply, Service, and Value Tracing ecosystem services along the chain of supply, service, and value helps clearly delineate the ecological and social processes that intersect in human experience of and dependence on nature. The supply of an ecosystem service is the ecosystem function that supports the service, the total potentially available to society, irrespective of actual consumption of or demand for that supply. The ecosystem service is the intersection of the supply of a service with the demand for or use of that service, as determined by location or action of the beneBiodiversit y and Ecosystem Services   269

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2 3

Habitat within serviceshed

FIGURE 15.2 A serviceshed is the area that provides a specific service to a specific beneficiary. A serviceshed’s location and extent depend on the biophysical processes that determine the supply of a particular service as well as on the ability of people to access those benefits. This figure shows how servicesheds differ across three services—​water provision, crop pollination, and carbon storage—​for four beneficiaries, specifically three farms and a city.

Left: For water provisioning services, the serviceshed is the area upstream of a beneficiary’s point of water access, as the quantity and quality of water will depend on the state of upstream ecosystems. Some servicesheds overlap: the servicesheds of Farm 1 and Farm 2 fall entirely within Farm 3’s serviceshed because Farm 3 takes water from further downstream. Because the city pipes water across a distance, its water provision serviceshed is also located far from the city itself. Middle: Servicesheds for crop pollination depend on pollinators’ flight ranges. The current provision of pollination services occurs where pollinator habitat overlaps with a farm’s serviceshed, as indicated in red. In some cases a pollinator habitat is too far away from a farm for that habitat to provide crop pollination services, as is the case of the habitat located below Farm 2. Right: In the case of carbon storage, all beneficiaries share a global serviceshed. Because the Earth’s atmosphere is well mixed, everyone is affected by carbon sequestration and carbon emissions that occur anywhere else in the world.

ficiaries. The value of an ecosystem service encompasses the economic and/or social impacts of the service provided, the contribution of the service to human well-being, and can be assessed by combining the magnitude of the ecosystem service provided with social preference (including both needs and desires). Another popular method of ecosystem service assessment, benefits transfer, skips over ecosystem function to apply economic value estimates from one location to a similar site in another location. Benefits transfer treats all units of a given landscape type as identical, which makes resolving the specificity of the landscape type and deciding what constitutes “similar enough” particularly important and notably challenging (Plummer 2009). The supply chain approach, in contrast, connects ecological processes to social outcomes, without any intervening assumptions. The supply of an ecosystem service can be thought of as an ecological production function, the outputs of ecosystem processes that are relevant to supporting and sustaining human life (Tallis and Polasky 2009). In this sense, the biotic structure of ecosystems and environmental conditions interact in processes that regulate and regenerate the natural capital upon which we depend. Societal decisions about land and marine management and natural resource use affect the structure of ecosystems and therefore the functions they perform. Many efforts have modeled the structural and functional responses of ecosystems to human-induced change in 270  Biota

order to better inform decisions (reviewed by Waage et al. 2008, Nemec and Raudsepp-Hearne 2012). Moving beyond the biophysical supply of an ecosystem service, which is where many ecosystem service assessments stop, requires a consideration of a serviceshed—​the area that provides a specific service to a specific beneficiary (Tallis, Polasky et al. 2012). Servicesheds can operate over vastly different scales, depending on the nature of the service (Figure 15.2). Carbon sequestration and its contribution to climate regulation benefits all humans on the planet, which suggests its serviceshed would be global. In contrast, the serviceshed for water production or purification would in nonengineered systems be the watershed upstream from a particular city’s extraction point. Modern proclivity for pumping and delivering water from watersheds far removed from consumption points (such as is the case throughout much of California) can create distant and noncontiguous servicesheds for these services. Some services cannot be transported, and their benefits will be experienced much more locally; the servicesheds of pollination and pest control services provided by naturally occurring insects (rather than managed honeybees or biocontrol agents) will be constrained by the foraging distance of those insects from a particular farm. Selecting the appropriate scale over which to delineate these servicesheds is necessary to connect the supply of the ecosystem service to the value accruing to the people who benefit from it.

Valuation of ecosystem services can take many different forms and produce different metrics, both monetary and nonmonetary. Monetary valuation has focused on distinguishing between market and nonmarket values of different services and various methods for assessing nonmarket values (Mendelsohn and Olmstead 2009). Market-based valuation is applied for ecosystem services that are already traded—​primarily the “provisioning” services, such as crop, livestock, fish, and timber production, but also potentially “regulating” services, if a cap-and-trade program or similar regulatory framework is in effect, such as the Clean Air Act for air quality or California’s AB 32 for carbon sequestration. Nonmarket valuation includes stated preference methods, what people say they are willing to pay for an ecosystem service or willing to accept when losing access to a service, and revealed preference methods, which use proxies of related consumer behavior, such as the cost to travel to a place where ecosystem services are consumed or a comparison of property values that differ only in their access to a particular ecosystem service (Freeman 2003). Avoidance or replacement costs can also fall under this category of revealed preference; the water-filtering capacity of a wetland, for instance, may be valued as the cost of building and maintaining a water treatment plant that would be necessary if the wetland were not providing that service. Approaches for valuing marginal changes in ecosystem services are important in the consideration of trade-offs between different decisions, and the value of such marginal changes can be dramatically different depending upon where they fall on the ecosystem service supply curve (Fisher et al. 2008). Marginal values of services can increase exponentially as they grow scarcer, especially those services responsible for critical life-support functions and without an adequate or sustainable technological substitute (Heal 2000). These vastly different approaches to monetary valuation can produce substantially different estimates of value, and their appropriateness should be considered carefully for the services and decision in question. There are many aspects of human well-being that can be captured through monetary valuation, but it is not always desirable to express ecosystem service values in such currencies to inform decisions. For example, human health risk metrics already are widely adopted and accepted by public and private decision-making bodies. Thus the contribution of water quality to human health can be valued as the reduced risk of mortality or morbidity (Keeler et al. 2012), or can also be valued in terms of the percentage of population with access to clean drinking water and/or resulting reduction in disease burden that may be set through national or international goals. Likewise, crop production, fisheries landings, or nontimber forest products could be assessed at their market value, or by their contribution to alleviating malnutrition (Tallis, Mooney et al. 2012). Cultural services span a particularly wide range of values, from market-based values of visitation rates for recreation, to religious values for sacred sites or plants, to nonuse values of inspiration and wonder; new approaches to integrating such disparate values are necessary to include them in environmental assessment (Satz et al. 2013). A new branch of cultural services is emerging in the realm of psychological ecosystem services, the benefits of human experience of nature to cognitive function, including concentration, impulse inhibition, short-term and working memory, and mood (Bratman et al. 2012). Expanding beyond our traditional notions of “value” will help more

thoroughly capture the full benefits of ecosystem services to people.

Examples of Ecosystem Functions and Services in Californian Ecosystems Agricultural Production California supplies a diverse mix of agricultural products, producing $43 billion in revenues in 2011 across more than eighty crop and animal products (USDA National Agricultural Statistics Service 2012) (see also Chapter 38, “Agriculture”). Crop production is an ecosystem service in and of itself and is also supported by pollination and pest control services provided by beneficial insects that inhabit the farm and surrounding areas. Livestock that graze on California pastures, grasslands, and open woodlands represent another important ecosystem service in the agricultural system. As provisioning services with a direct market value, service valuation is fairly well advanced, even for the partial contributions to the production value of supporting services like pollination and pest control. The ecosystem function or biophysical supply (crop or grassland productivity, delivery of pollen to crops, or suppression of pests) is well-documented in the literature, the service to people (available food for livelihood generation and consumption) is readily understood, and its value (revenues from crop or livestock production and the contribution of pollination and pest control to generating those revenues) is easily accounted for in economic analysis. Less well explored is the full value of agricultural production and supporting services to human nutritional health and overall food security, which do not always align with economic returns from farming. While growing cash crops provides an economic resource base from which to buy other goods, including a more diverse selection of food, this can leave growers vulnerable to environmental and economic shocks and ultimately less food secure. California’s high diversity in agricultural production allows for an unusual degree of localized food and nutritional security in the modern globalized food economy.

Supporting Agricultural Services: Pollination and Pest Control Pollination by wild bee populations enhances production of many high-value crops in California, including watermelon (Kremen et al. 2004), almond (Brittain et al. 2013), sunflowers (Greenleaf and Kremen 2006b), and tomatoes (Greenleaf and Kremen 2006a). Seminal work in watermelons has illustrated a general trend found in California and throughout the world: crop pollination services provided by native bee communities strongly depends on the proportion of noncrop or unmanaged habitat around the farm site. Kremen and colleagues (2004) found that native bee populations could fully support watermelon yields (without managed honeybees) in farms situated in areas containing more than 40% of natural habitat within a 2.4 kilometer radius. This supporting service is becoming increasingly important as colony collapse disorder and other threats posed to honeybee colonies compromise their availability for crop pollination on many farms. Furthermore, wild bees can enhance production in certain crops that honeybees cannot. Tomatoes are self-pollinating Biodiversit y and Ecosystem Services   271

to a certain degree, but bumblebees in California have been shown to pollinate with a particular vibration frequency that releases more pollen than would otherwise be achieved, increasing yields accordingly (Greenleaf and Kremen 2006a). Wild bees can also enhance the efficiency or productivity of honeybees through their direct and indirect interactions with this managed species. Wild bees have been shown to increase bee movement between male and female flowers by disrupting honeybee visits to sunflowers in the Central Valley of California, where 90% of U.S. hybrid sunflower seed production occurs (Greenleaf and Kremen 2006b). This enhanced pollination is worth an estimated $10.4 million annually, which nearly doubles the value for direct pollination by honeybees. Such bee movement between male and female flowers is critical in all crops with separate male and female flowers, such as melons, pumpkins, and kiwis, or with self-incompatibility, such as apples, sweet cherries, and almonds. Wild bees have similarly been shown to enhance honeybee efficiency in California almonds—​a crop so high in value that even the modest gain in efficiency of a 5 percent yield increase would augment economic returns to the farmer by $591 per hectare (Brittain et al. 2013), or $182 million across 307,560 hectares of almond in California (USDA National Agricultural Statistics Service 2012). Pest control services can also be enhanced by natural or unmanaged perennial habitat surrounding the farm, which concentrates or sustains populations of predators and parasitic insects, natural enemies of agricultural pests. Diversity and abundance of natural enemies has been found to increase with the proportion of noncrop habitat in the landscape (Chaplin-Kramer, O’Rourke et al. 2011), and this trend has been linked to pest suppression in many systems in California. Insect populations in tomatoes in the Central Valley responded more strongly to landscape features than to onfarm management (conventional or organic), with higher natural enemy populations and lower pest populations on farms located in more natural landscapes (Letourneau and Goldstein 2001). Lepidopteran pests in the Central Coast region of California were better contained by parasitism in landscapes with less annual cropland (Letourneau et al. 2012), and aphid pests in the same system showed suppressed population growth by natural enemies in landscapes with more natural habitat (Chaplin-Kramer and Kremen 2012). Vertebrate predators can play important roles in maintaining pest control on farms as well; conservation practices that benefited the western bluebird in vineyards in the Russian River Valley doubled to tripled the removal rate of important grape pests (Jedlicka et al. 2011). It is important to recognize that natural or unmanaged habitat in agricultural landscapes could provide ecosystem disservices as well, by providing resources for pests and pathogens in addition to natural enemies. For example, the non-native brassicas that dot many Californian roadsides and fallow fields or pastures provide an enemy-free refuge for aphids that can then spill over onto vegetable crops (ChaplinKramer, Kliebenstein et al. 2011). Non-native predator species can also outcompete more specialized native predators, reducing effective pest control, as has been shown for spiders in Napa Valley vineyards (Hogg and Daane 2011). Recently, concern over food safety emerging from a 2006 disease outbreak of Escherichia coli in California spinach has added to the perception that unmanaged habitat adjacent to farms harbors harmful and undesirable species. New protocols put in place in the Salinas Valley in response to this outbreak have 272  Biota

resulted in the loss of 13.3% of remaining riparian habitat in the region, and if these practices were implemented statewide, it is estimated that up to 40% of riparian habitat and 45% of wetlands in some counties would be impacted (Gennet et al. 2013). This “sterilization” of farm landscapes to reduce the risk of microbial food contamination events should be weighed against the value of possible pollination, pest control, and other ecosystem services that would be lost in the process. In many cases, the types or qualities of habitat that contribute either ecosystem services or disservices are very different, and it may be possible to manage to promote the services and constrain the disservices (Chaplin-Kramer, Kliebenstein et al. 2011). Valuation of pollination and pest control could take several forms. Pollination services have been valued in terms of the replacement costs (of renting honeybees or hand-pollinating) to perform the same service provided by wild pollinators (De Groot et al. 2002). The replacement cost for pest control services could be considered the amount spent on pesticides, although it is somewhat controversial whether the two methods provide equivalent levels of control (Pimentel et al. 1992). A more commonly used approach for valuing pollination services estimates the value of crop production attributable to pollination. Crop yield reduction in the absence of pollinators has been approximated for all major crop types (Klein et al. 2007), and this yield reduction is then multiplied by the market value of production (Morse and Calderone 2000). The analogous approach in pest control is to quantify the amount that yield is reduced by pest damage in the absence of natural enemies, and value the avoided damage (Losey and Vaughan 2006). Recently a new approach to valuing these supporting services in agriculture has been developed, called the attributable net income method, which subtracts the cost of inputs to crop production from the value of the ecosystem service and does not attribute value to the service in excess of the plants’ requirements (Winfree et al. 2011).

Forage Production The rangelands that account for about half the land area (approximately 24 million hectares) of California (Brown et al. 2004) produce grass and hay for livestock forage but also provide many other ecosystem services, including pollination ­(Chaplin-Kramer, Tuxen-Bettman et al. 2011) and carbon sequestration (Silver et al. 2010; see below; see also Chapter 37, “Range Ecosystems”). Unlike many industrialized cropland systems, rangelands can provide habitat for wildlife and support an aesthetic and a quality of life cherished by many Californians. Over half of oak woodlands landowners in California are estimated to produce livestock, though this figure has been declining steadily since 1985 as productive land is converted to vacation homes (Huntsinger et al. 2010). Most valuation in these systems, however, focuses on the forage production supporting a livestock industry that constitutes a major component of California’s economy. By rough estimates, California produces 13 Teragrams (1 million Megagrams or metric tons) of forage annually, which could support an average of 1.5 million cows without any supplemental feed (according to simple approximations assuming 4,300 kilograms of dry matter per year per cow with 50% consumption to account for trampling; Shaw et al. 2011), which is approximately a third of all livestock produced in California (USDA National Agricultural Statistics Service 2012). This 13 Tg of forage production can be valued as $165

million in profits from livestock or $555 million in cost of replacement for hay (Shaw et al. 2011). More local-scale assessments suggest high spatial heterogeneity in the supply and value of this service, with much higher capacity for production in the wetter northern than the drier southern portions of the San Francisco Bay Area (Chaplin-Kramer and George 2013) and much higher values for production in Kern County than in neighboring San Luis Obispo County due to proximity to feedlots, slaughterhouses, and transportation routes (Chan et al. 2006).

Timber and Nontimber Forest Products California’s 16.1 million hectares of forest covers almost 40% of the state, 6.9 million hectares of which are classified as timber lands, supporting a large timber industry (Shih 1998, Laaksonen-Craig et al. 2003; see also Chapter 36, “Forestry”). Although timber does not play as large a role in the state’s economy as in other west coast states, 3.04 million cubic meters (1.29 billion board feet) of timber was harvested in California in 2011, worth $272 million (California State Board of Equalization 2012). Including what the Forest Service calls “multiplier effects,” the additional economic activities from processing the timber into final products, yields a much higher estimate—​over $12 billion for the 5.0 million cubic meters (2.1 billion board feet) produced in 1999 ­( Laaksonen-Craig et al. 2003). This forest products industry and related economic sectors support more than 220,000 jobs in California, which is only 1% of the state’s total employment, but comprises a vital livelihood for many people living in rural areas (Laaksonen-Craig et al. 2003). Managed timber forests also provide a complex web of ecosystem services that contribute to human well-being, along with the state’s 9.3 million hectares of nonproduction forest. California forests provide nontimber forest products such as mushrooms, floral greens, a variety of edible and medicinal plant species, and wildlife for hunting (Krieger 2001). As reviewed in other sections, forests also sequester carbon, provide recreational opportunities, support pest control and pollination services to nearby farmland, and regulate water supplies to people.

Fisheries and Aquaculture The California commercial fishing industry landed over 181 million kilograms of fish and shellfish in 2011, worth over $200 million of landed value (CDFG 2012a) (see also Chapter 35, “Marine Fisheries”). Like farming and ranching, fishing provides not only revenues but a livelihood and a way of life; the seafood industry employs over 122,000 people in California, more than in any other state (National Marine Fisheries Service 2013). Meanwhile, aquaculture is growing rapidly in California and around the world. It is the fastest growing animal food production sector, and within a few years is expected to produce half of the fish we eat (Klinger and Naylor 2012). The income gained from commercial fishing and marine aquaculture operations reached $4.3 billion in 2011, adding a total of $7.2 billion to California’s economy in direct and related revenues (National Marine Fisheries Service 2013). The supply of California’s seafood is supported by healthy marine ecosystems and the water quality of streams feeding into the ocean. Many commercially important species rely on offshore habitat, such as rockfish on rocky reefs, and tuna

in pelagic ecosystems, but some of the highest value species, such as farmed oysters and salmon, are particularly sensitive to nutrient and sediment loadings, freshwater inputs, and other services provided by upland and coastal ecosystems. Low streamflows contributed to the rapid decline of California’s salmon populations, and the complete closure of California’s commercial salmon fishery in 2008 and then again in 2009, resulting in the loss of 1,823 jobs and $118.4 million in income (Macfarlane et al. 2008, Business Forecasting Center 2010). In addition to supporting commercial fishing and aquaculture, healthy coastal and marine ecosystems also support a recreational fishing industry in California, providing recreational opportunities to over 1.6 million anglers, supporting over 7,700 full- and part-time jobs in the state, and generating $1 billion of value added to the local economy in 2011 (National Marine Fisheries Service 2013).

Water Production and Regulation The relationship between water production and vegetation is complex, with dynamic processes operating across multiple spatial and temporal scales. Supply of this service, the total wateryielding capacity of a watershed, can be more difficult to measure and understand than the service provided to people, in the form of water available for human uses such as irrigation, drinking water consumption, and hydropower. Monetary valuation of these uses is based on market values of streamflow, which range from $0.21 per 1,000 cubic meters ($0.26 per acre foot) for electricity generation to as much as $40 per 1,000 cubic meters ($50 per acre foot) for irrigation and municipal uses (Krieger 2001). Nonmarket values of water include supplying natural systems for biodiversity, aesthetic value, and recreation. Southern California residents were willing to pay $115 per household to raise the water level in Mono Lake (Brown 1992). However, the pathways by which water is captured and delivered to people and ecosystems are often uncertain, especially the interplay between surface water and groundwater resources, making it difficult to assign value to particular ecosystems providing the service. In general, greater evapotranspiration rates from increased vegetative growth can reduce overall water supply downstream, although the biophysical mediation of hydrologic processes results in more regular and reliable flows of water (as opposed to long dry periods punctuated by flash flooding during rain events) (Brauman et al. 2007). Some habitats create their own microclimates to capture more water from the air than would otherwise be possible. The coastal redwood forest of California provides more intercepting surfaces to capture moisture in its tall vegetation, storing twice as much water from fog as treeless sites (Brauman et al. 2007).

Water Purification Water filtration or purification through the trapping of sediments and nutrients by vegetation is a regulating service that is the basis for many payments for ecosystem services (PES) programs (Long Tom Watershed Council 2008, GoldmanBenner et al. 2012), but very often the assessment of this ecosystem service stops at the ecosystem function or supply of the service (biophysical nutrient retention, or reduction in nutrient loadings in streams), without consideration of service to people (water quality at intake points for drinking water) or the value of that service (avoided cost of building Biodiversit y and Ecosystem Services   273

or maintaining a treatment plant). One of the most famous examples of a valuation of water quality services is New York City’s $2 billion investment in restoration of the Catskills to avoid spending $8 billion to build a water treatment plants plus $300 million annually in operating costs (Turner and Daily 2007). In California there has not yet been the same level of economic analysis, but the biophysical supply of the service is well documented throughout the state. In the San Diego region restoration of the 8 hectare Famosa Slough reduced nitrogen loading and resulting algal blooms in Mission Bay, with co-benefits including enhanced mudflat habitat to support better wildlife viewing opportunities for residents (Zedler and Kercher 2005). Such coastal wetlands have been shown to be effective nutrient traps across California, including Elkhorn Slough in the Monterey Bay region, where this function is not saturated despite high background nitrogen inputs from intensive upstream agriculture in the Salinas Valley (Nelson and Zavaleta 2012). Quantitative analysis of area requirements to perform this ecosystem service indicates less than 3 percent of the land in agricultural watersheds in the San Joaquin Valley would need to be in wetlands in order to reduce nitrate concentrations below a required threshold of 0.5 mg L –1 (Karpuzcu and Stringfellow 2012). Similar analyses have shown that the current wetland area enrolled in the USDA Wetlands Reserve Program in the Central Valley could denitrify nitrate loads to the delta in as little as eighteen days (Duffy and Kahara 2011). Riparian buffers can provide similar services to wetlands over much smaller areas, and valuation of the incremental impact of changes in water quality is simpler over this finer scale; a net benefit of vegetated buffer strips to farmers has been demonstrated in California, with improvements in water quality and reductions in soil loss outweighing the opportunity cost of land taken out of production (Brauman et al. 2007).

Carbon Sequestration and Storage A key component of climate regulation is reducing the greenhouse gases responsible for global warming through carbon sequestration. In this case, the supply is the same as the service, in that all people on the planet benefit from lower concentrations of carbon dioxide in a globally mixed atmosphere. The value may be different for different populations, depending on their vulnerability to climate change, with low-lying coastal areas more sensitive to sea level rise and certain areas at higher risk of losing agricultural, forest, or water production than others (Mimura et al. 2007). California’s varied ecosystems store carbon to different degrees, and the total live aboveground carbon stored in California is estimated at 1025 Tg, which, with a social cost of carbon between $23 and $185 per metric ton, could be valued in the range of $23.5 billion and $190 billion (Shaw et al. 2011). The abundant biomass of forests make them especially important carbon sinks, and California has some of the most carbon-rich forests on the planet; the living aboveground biomass in the California coastal redwoods alone is thought to contain five times more carbon than any other forest (Sillett et al. 2010). Freshwater and tidal wetlands are also critically important in terms of carbon storage in California (Trulio et al. 2007). However, there is a difference between the stock of stored carbon, for which value should be considered from the perspective of avoided loss, and the flow of carbon sequestration, for which value can be thought of as the contribution to reduc274  Biota

ing total global greenhouse gas emissions (Tallis, Polasky et al. 2012). California’s working landscapes are in constant flux of growth and harvest, and thus there is important potential for greater carbon sequestration through improved management of these lands (Ackerly et al. 2012). Managed forests are considered to be at only 50% of their carbon carrying capacity and could considerably increase sequestration through selective logging that increases the stocking rate of large (>110 cm diameter at breast height) trees (Roxburgh et al. 2006). Similarly, rangeland and cropping systems could sequester more carbon through grazing management, pasture improvement, no-till agriculture, and conversion to perennial crops. Soil carbon pools in California rangelands are already higher than in the rest of the U.S., likely due to the prevalence of woody plants, which must be balanced against forage production needs for livestock (Silver et al. 2010). Improved grazing practices such as moderate stocking rates can significantly increase rates of soil carbon sequestration (Conant et al. 2001), and a high degree of variability in soil carbon pools in California’s rangelands across similar soil types and climate suggests there is a considerable role for management to increase soil carbon (Silver et al. 2010). Meanwhile, California’s agriculture sequesters an average of 19 grams carbon per meter per year, which is equivalent to 0.7% of the state’s fossil fuel emissions, and this amount could be doubled if conservation tillage and improved orchard pruning and waste management practices were implemented (Kroodsma and Field 2006).

Hazard Prevention: Flood Mitigation and Coastal Protection Many California ecosystems have the capacity to moderate extreme events such as floods and coastal storms. Maintaining wetland or riparian habitat that is part of a natural flood regime can reduce the risk of catastrophic flooding by retaining in soils and vegetation high volumes of water that would run off of impervious surfaces in a watershed. Similarly, maintaining benthic habitat such as mangrove or eelgrass along shorelines can provide coastal protection from erosion and flooding by attenuating storm surge. The water retention or wave attenuation resulting in reduced flooding or coastal erosion is the biophysical supply of this regulating service; intersecting this ecosystem function with the location and activities of people yields the service itself, in the form of reduced damage to property. The value of this service is therefore typically measured by avoided damage costs, although consideration of the distributional impacts of this value is particularly important in this case, as poorer populations with low property values may be disproportionately situated in high-risk areas (Arkema et al. 2013). Restoration of 150 hectares on floodplain in Napa Valley has provided flood mitigation along a 10 kilometer stretch of the Napa River, as well as a host of co-benefits such as recreation and tourism activities (Turner and Daily 2007). Flood mitigation services are often correlated with recreation, whether in floodplain or narrower riparian habitats, and has also been shown to overlap with pollination in the Salinas Valley and forage production in the San Jose area (Chan et al. 2006) and nutrient retention in the Central Valley (Duffy and Kahara 2011). Coastal habitats providing protection from erosion and coastal flooding are important as well; in fact, a nationwide assessment of coastal hazards found California’s

coastal ecosystems to provide protection for the greatest number of total people, socially vulnerable populations, and properties in the country (Arkema et al. 2013).

Recreation California’s open space and coastal areas provide significant economic and social benefits. The supply of this service is determined by the existence and quality of habitat in which recreation activities could occur, while the service itself depends on access to the recreation area, such that aesthetically pleasing areas nearer to urban centers will have higher service values for recreation. Recreation can be valued a variety of ways: in market terms, through visitation fees and other revenue generated by tourism; nonmarket methods include stated willingness to pay or revealed preference methods such as travel cost to recreation areas or property values near these areas. Outdoor recreation in California generates revenues of approximately $46 billion annually and supports 408,000 jobs (Outdoor Industry Foundation 2006). Beaches are a particularly important recreational resource in California, both to residents and to tourists. Over 15 million people a year use California’s beaches, making over 150 million trips a year (Pendleton and Kildow 2006). California residents spend over $4 billion a year on beach recreation (Pendleton et al. 2011). On top of this, California’s beaches generate $14 billion a year in related revenues from beachgoers spending money on amenities such as food, parking, and beach-related activities, making California’s state beaches a multibillion-dollar natural resource (Pendleton et al. 2011).

Biodiversity, Ecosystem Functioning, and Ecosystem Services in Decision Making Parts of the conservation field have broadened in scope in recent years to include goals to maintain and enhance human well-being in addition to preserving biodiversity (Kareiva and Marvier 2012). This has created some tensions within the conservation community, amid concerns from many conservation scientists and advocates that ecosystem services and “valuing nature” will distract from and dilute efforts to conserve biodiversity (Soule 2013, Doak et al. 2014). It has been shown that integrating ecosystem services into decisions can expand the potential partners and funders for conservation rather than detracting from funding available for biodiversity-focused conservation (Goldman et al. 2009, Reyers et al. 2012). However, it is true and worth considering that winwins between biodiversity and human well-being are not always possible; in some cases, nature provides disservices or comes into direct conflict with people (McCauley 2006). Integrating ecosystem service information into decisions can expand the potential partners that support environmental protection and restoration beyond those within the traditional conservation community (Reyers et al. 2012) and can help make explicit the conflicts and synergies among stakeholders with different goals (Guerry et al. 2012). Recognizing the value of multiple ecosystem services for human wellbeing is a first step in connecting ecosystem service value to land and ocean management decisions, but it is also necessary to have a clear understanding of the needs and expectations of different institutions involved in decision making (Daily et al. 2009). Creating this type of institutional change in natu-

ral resource decisions requires a clear accounting of the condition of the ecosystem (supply metrics), the amount of natural resources used by people (service metrics), and the people’s preference for that level of service (value metrics) (Tallis, Lester et al. 2012; Tallis, Mooney et al. 2012). Information about biodiversity, ecosystem function, and ecosystem services will guide decisions in different ways, and if considered as separate approaches, they will often result in different visions of what our world should look like. However, there is often room for compromise and finding complementarities among biodiversity and ecosystem services in these approaches. Three specific decision contexts are reviewed here (conservation planning, climate adaptation, and impacts assessment), and biodiversity- and ecosystem services–​based approaches are compared for each.

Conservation Planning Maintaining biodiversity, ecosystem function, and ecosystem services in the face of humanity’s increasing dominance of Earth’s ecosystems requires landscape-level decisions about where to manage for production of food, materials, and other benefits to people, and where to limit human activities. Conservation planning involves identifying the location and configuration of areas that, if properly managed, can promote the persistence of biodiversity and other natural values (Pressey et al. 2007). An unavoidable reality is that human activities very often threaten biodiversity. While wildlife-friendly forms of management can lessen the severity of such threats, successful conservation requires an explicit recognition of their existence, and guiding activities that pose a threat away from the most sensitive or most diverse or unique places will generally benefit biodiversity. Conservation planning is thus critical to effectively balancing human needs with the preservation of nature, but when framed in terms of biodiversity only, conservation decisions are viewed as inherently in conflict with development. From this perspective, setting aside areas for protection of biodiversity requires forgoing uses that might otherwise benefit people. This can cause misunderstanding and resentment, and the perception that conservation prioritizes endangered species over livelihoods, other species over humans. Ecosystem-service approaches to conservation planning can reduce this apparent conflict, by considering a wider range of nature’s values to people and better capturing both trade-offs and synergies that emerge from conservation. An ecosystem services–​based framework can illuminate the ways in which conservation in specific places might purify water, enhance fish stocks, provide recreational opportunities, mitigate hazards like flooding, or buffer against extreme heat events in cities. For example, marine protected areas designed to reduce degradation of coastal ecosystems also provided an estimated $2.5 million per kilometer of coastline in recreational benefits to beach visitors in southern California (Hall et al. 2002). Furthermore, ecosystem services approaches can create opportunities for conservation that would not exist with a biodiversity-only approach. Napa County restored 250 hectares of floodplain that would never have been undertaken for preserving biodiversity alone; this massive conservation investment was viable due to the perceived win-win outcomes for flood mitigation as well as enhancement of fish and wildlife populations, scenic beauty, and the accompanying recreation and tourism benefits (Turner and Daily 2007). Residents were willing to pay an additional $50 million to Biodiversit y and Ecosystem Services   275

implement this ecosystem services approach to flood control over the traditional infrastructure approach because of the anticipated value of those additional benefits. While biodiversity- and ecosystem services–​ b ased approaches to conservation planning can often be employed together, the two approaches are not necessarily substitutable, especially with services that are negatively correlated with biodiversity. This is often the case with provisioning services that are maximized by more intensive management of the land. In the Central Coast region of California, reserve design that maximized forage production and pollination services would only protect 44% of the biodiversity targets achieved by a reserve system based solely on biodiversity goals; however, more strategic targeting of biodiversity and only the services with which it was positively correlated (water production, carbon storage, flood control, and recreation) achieved nearly the same biodiversity outcomes as when targeting biodiversity alone, with many additional benefits (Chan et al. 2006). In practice, the complementarity of ecosystem services–​ based and biodiversity-based approaches will depend upon how well the scales over which the services are produced match the habitat requirements of the species of interest. Flows of some services (e.g., climate regulation, waterrelated services) can originate at great distances from the ultimate beneficiaries, but many services are provided at more local scales (e.g., agriculture-related services, coastal protection), requiring conservation of smaller areas close to people and embedded within human-dominated landscapes. Such configuration may be difficult to reconcile with a biodiversity-based approach that prioritizes large contiguous areas removed from anthropogenic impacts. Merging the two approaches will be most successful when connectivity between protected areas can also establish connectivity between ecosystems and people and the flows of benefits. For example, large coastal no-take marine protected areas can severely reduce the ability of small, nearshore fleets to access fish stocks, thereby reducing the provision of fishing services (Roberts et al. 2005). In some cases, smaller but more numerous no-take areas could better provide benefits to the nearshore fleets by maintaining access to fishing grounds, while still protecting fish stocks and associated biodiversity. In contrast to marine systems, grassland ecosystems in California are largely privately owned and managed as rangelands for livestock production, making large protected areas unfeasible. However, integrated management of California rangelands can provide many benefits both for species and people, including livestock production, recreation opportunities, carbon sequestration, and improved water supply (Kroeger et al. 2010). Protecting public land targeted to maximize conservation of species while promoting incentives for best management practices on private rangelands in these regions can connect biodiversity- and ecosystem services–​based approaches (Community Foundation Sonoma County 2010). Despite some incompatibilities between certain services and at certain scales, biodiversity and ecosystem services approaches to conservation have much in common (Reyers et al. 2012). Many argue that integrating the two objectives in landscape-scale conservation planning decisions can increase the range of people in support of conservation while achieving better outcomes for both nature and society than if either objective is considered alone (Goldman and Tallis 2009; Ruckelshaus, McKenzie et al. 2013). However, integrating biodiversity and ecosystem service approaches will not eliminate 276  Biota

trade-offs, and ultimately conservation planners will need to decide what to prioritize

Climate Adaptation Climate change impacts individual species, the ecosystems they comprise, the processes or functions of those ecosystems, and ultimately the ecosystem services provided to people. The effects of climate change on species and ecosystems include range shifts, asynchronies in phenology, altered food web dynamics and community interactions like pollination, and disruption of key functions such as seasonal water flows upon which other ecosystems rely (Moritz et al. 2008, Mawdsley et al. 2009, Hoegh-Guldberg and Bruno 2010, Walther 2010, Null and Viers 2012, National Climate Assessment 2013). Climate adaptation planning addresses these impacts, though the approaches taken and results of these efforts differ depending on whether biodiversity or ecosystem services are the focus. With a biodiversity-based approach, the focal unit is species or ecosystems, and adaptive measures may include increasing the number or area of reserves, improving management within reserves, and maintaining and enhancing connectivity (Heller and Zavaleta 2009, Mawdsley et al. 2009). Species-specific adaptation strategies tend to focus on managing and restoring reserves and corridors specifically for the focal species’ needs or reintroducing or relocating species at risk (Rahel et al. 2008, Grewell et al. 2013). However, such species-specific climate adaptation does not address resilience of the system in the face of novel species and interactions that may occur with climate change (Seastedt et al. 2008, West et al. 2009, Bernhardt and Leslie 2013). Uncertainty associated with projections of climate change and species’ range shifts could also lead to adaptation efforts that do not provide an appropriate range of conditions for species’ persistence. Climate adaptation planning can also focus on maintaining biodiversity at the ecosystem level—​for example, prioritizing conservation and restoration of marshes for a diversity of habitat types (Stralberg et al. 2011) or managing river flows to support aquatic systems (Palmer et al. 2008). An ecosystem services approach to climate adaptation will consider changes in the provision of services rather than tracking movement of species or even habitats across landscapes. A serviceshed depends upon the location and activity of beneficiaries of the service as well as the scale at which they can benefit from it, and its delineation will change primarily in response to demographic changes. Thus, while the magnitude of service provided will certainly change with the structure of the ecosystem responding to climate change, servicesheds will tend to be more stationary than species and habitats in the face of climate change. It is therefore important in an ecosystem services approach to adaptation to anticipate changes in the structure of ecosystems within the serviceshed that will lead to changes in function that will impact people. Impacted services may be species-specific, such as pollination and pest control of crops (Luedeling et al. 2011, Vanbergen and Initiative 2013), and fisheries (Brander 2010; Doney et al. 2012; Ruckelshaus, Doney et al. 2013), or at the whole habitat level, as with coastal protection from storms and sea level rise (Hayhoe et al. 2004, Heberger et al. 2009, Doney et al. 2012) or recreation on beaches and in the mountains (Hayhoe et al. 2004, King et al. 2010). Understanding and adapting to climate impacts on species-specific services will require much of the same research

as biodiversity-based adaptation, particularly with regards to shifts in resource availability and the timing of community interactions (Brander 2010). However, a higher degree of substitutability exists when conserving species-specific services than when conserving specific species; pollinators or commercially important fish that respond more positively to climate change could potentially compensate for the loss of function of species harmed by climate change. Habitat-specific services may be less substitutable in many cases, such as sea level rise causing inundation of marsh or beach habitat when there is no space for the habitat to migrate upshore (Heberger et al. 2009; Hanak and Moreno 2011; Committee on Sea Level Rise in California, Oregon, and Washington et al. 2012). In these cases, the service of coastal protection or recreation will be lost, with no option for replacing it. Other services, such as carbon sequestration and water provision, are produced by many different habitats, though to different degrees. Woody encroachment throughout much of California’s grasslands and the retreat of many of the state’s oak woodlands expected in response to climate change (Cornwell et al. 2012) will alter the patterns of carbon and water cycling, resulting in changes to climate-regulating carbon sequestration and to water supply in certain regions. An ecosystem services approach to climate adaptation prioritizes the best locations for protection and restoration of habitats within the same region or serviceshed, to produce optimal outcomes for the people who remain there, in contrast to the biodiversity-based approach of planning reserve networks to follow species or habitats. These two approaches will often not be complementary and in some cases may conflict with each other. For example, in the San Francisco Bay two endangered estuarine plants are less successful at upslope colonization when they are closer to anthropogenic activity, while coastal protection services provided by estuaries are of greatest benefit in areas of high population density, where they will protect the most people (Grewell et al. 2013). It is therefore important for climate adaptation projects to have very clear and specific goals, whether conserving biodiversity or ecosystem services or both, and identify where and how to meet each goal, separately, if needed.

Environmental Impact Assessment and Mitigation Development can create new jobs, increase food production, improve transportation and communication, and provide cheaper energy. However, these benefits need to be weighed against the environmental costs, which often include water and air pollution, loss of recreation opportunities, and threats to species and habitats. Development in California is governed by a suite of federal and state legislation (e.g., the California Environmental Quality Act, the California Endangered Species Act, the National Environmental Policy Act, the Clean Water Act, and the Endangered Species Act) that requires developers to identify significant environmental impacts of proposed activities and determine how to avoid, minimize, and mitigate these impacts, where feasible. Mitigation offsets in the United States primarily require restoration of habitat or ecosystem functions, though sometimes protection that prevents future losses can also count as mitigation. Mitigation decisions—​how much mitigation is needed and where—​ depends to a large degree on whether environmental impact is viewed through the lens of biodiversity, ecosystem function, or ecosystem services.

In California, mitigation requirements have been established for the functions (water storage and transport; nutrient retention and cycling) and habitats (maintenance of native vegetation and wildlife) provided by wetlands as well as for endangered species. This mitigation for wetlands was traditionally undertaken on a project-by-project basis, with each developer held responsible for restoring or creating a replacement wetland to compensate for unavoidable impacts. However, this approach proved to be inadequate—​by some estimates, one-third of the required wetlands were never built, and those that were successfully created or restored tended to be small and fragmented (Rolband et al. 2000). Wetland mitigation banks—​consolidated areas of restored, enhanced, constructed, or preserved wetlands—​emerged in the 1980s and 1990s as a response to this problem (CDFG 2012b). Banks sell credits based on the area and quality of wetland the bank possesses to those in need of mitigation within the bank’s “service areas” (often watersheds) (Ruhl and Salzman 2006). In California as of 2012, thirty-three banks across more than 1090 hectares sell or have sold wetland mitigation credits (CDFG 2012b). In addition to wetland mitigation banking, California pioneered conservation banking, which, instead of providing credits for wetlands, provides credits for particular (often endangered) species or habitats. To date, California has approved credits for thirty-four species of animals and plants as well as twenty-seven habitat types (CDFG 2012b). Thirtytwo conservation banks have been approved by the California Department of Fish and Game, covering over 11,300 hectares of habitat (CDFG 2012b). While mitigation banking represents an improvement on the previous, piecemeal approach, there are still limitations to mitigation banking’s effectiveness at biodiversity conservation. Mitigation credits are calculated using simple metrics of habitat or ecosystem quality. Given the many facets of biodiversity and ways to measure it, as discussed earlier in this chapter, these simple approaches can obscure important aspects of biodiversity, such as genetic diversity or connectivity between populations, making it difficult to establish whether a mitigation credit adequately offsets the biodiversity lost with development (Burgin 2008, Walker et al. 2009). Finding robust yet practical ways to assess equivalence between damage to biodiversity from development and benefits to biodiversity from mitigation remains a challenge. However, even if mitigation successfully maintains certain aspects of biodiversity or ecosystem function, it does not necessarily restore ecosystem services at the same time. When mitigation redistributes biodiversity and ecosystems/ecosystem function across the landscape, ecosystem services can be transferred across servicesheds, altering who receives the benefits from biodiversity and ecosystem functions. In other words, mitigation that does not explicitly account for the link between the landscape and people can reduce ecosystem service benefits even when the total supply of the service stays the same (Tallis and Polasky 2011, Mandle et al. in prep). This redistribution of ecosystem services can create inequities. Mitigation banks, though they are likely to be more effective at achieving mitigation targets than the conventional, project-by-project, piecemeal approach, may also lead to the concentration of ecosystem services in areas near mitigation banks. Wetland mitigation banking is leading to the redistribution of ecosystems services, taking services such as pollution control, flood mitigation, and recreational opportunities away from urban populations and concentrating them in rural areas (Ruhl and Salzman 2006, Biodiversit y and Ecosystem Services   277

BenDor et al. 2008). Furthermore, biodiversity and some ecosystem services may be restored more slowly than habitat, and such time lags may lead to temporary inequity that needs to be addressed (Bullock et al. 2011). Explicitly incorporating ecosystem services into mitigation requirements would help provide a more complete picture of the environmental impacts of development and ensure that the benefits and costs of development are distributed equitably within society. Although there is growing interest from governments, business, and multilateral organizations in including ecosystem services in impact assessments and mitigation requirements (Council on Environmental Quality 2009, Landsberg et al. 2011, IFC 2012), the development of standardized approaches to determining suitable offsets for ecosystem services lags behind those that have been developed for wetlands and for particular species. This is an issue for institutions like the California Coastal Commission, which is mandated by the Coastal Act to protect biodiversity, such as environmentally sensitive habitats and species, as well as to ensure the continued provision of several ecosystem services including recreational opportunities, fisheries, aesthetics, and ocean access. When coastal development is permitted, the California Coastal Commission may levy mitigation fees to compensate for environmental losses, but there is currently not a comprehensive, systematic approach for determining mitigation requirements or fees. For example, the mitigation fee for development of a seawall in Monterey County only took into account the loss of recreational services, not other services such as coastal protection provided by the beach (Caldwell and Segall 2007). A holistic impact assessment and mitigation framework that simultaneously assesses impacts on biodiversity, ecosystem function, and ecosystem services would better guide development and mitigation decisions in the diverse decision contexts facing California.

Conclusion Ecosystem services can illuminate the many different benefits that functioning ecosystems can provide, rather than just benefits for a single species or single function, and relate those benefits to the people receiving them. Biodiversity has intrinsic value, and many conservation efforts do not require a broader approach to consider ecosystem function or ecosystem services to successfully preserve biodiversity. However, ecosystem functions are both affected by and can affect biodiversity, and therefore the functions underpinning ecosystem services can be correlated with biodiversity (Cardinale et al. 2012). This is not always the case, especially for sensitive species or habitats that are threatened by proximity to human impact and services that must be produced and received locally. Making such trade-offs explicit is better than not understanding or appreciating the consequences of natural resource decisions for multiple objectives. Where synergies do exist, considering biodiversity and ecosystem services together can create a stronger approach to conserving all aspects of nature that humans care about. An integrated framework for biodiversity and ecosystem services allows for scientific measurement of nature’s diverse benefits to assist policy makers in defining goals and provides managers with tools to track progress toward these goals, with improved outcomes for humans and the ecosystems they rely upon. 278  Biota

Summary This chapter introduces the concepts of biodiversity, ecosystem functions, ecosystem services, and their relationships to human well-being. The many facets of biodiversity are examined, along with current levels and losses of biodiversity in California in particular, and the consequences of those losses to ecosystem functioning. Ecosystem functioning is first presented within the framework of ecosystem ecology and ecological interactions and then extended as a foundation for the concepts of ecosystem services and natural capital. The socioecological systems thinking that defines ecosystem services and natural capital is described to help differentiate among ecosystem functioning, ecosystem services, natural capital, and the value of these numerous benefits of nature to people. Examples from Californian ecosystems illustrate a diverse set of ecosystem services provided by natural and managed systems. These include agricultural production, pest control, pollination, forage production for livestock, timber and nontimber forest products, fisheries and aquaculture, water production and regulation, water purification, carbon sequestration, hazard mitigation (from coastal and inland flooding), and recreation. The chapter closes with an examination of how information about biodiversity, ecosystem functioning, and ecosystem services can help inform decisions in different ways and for different policy contexts, such as conservation planning, climate adaptation, and permitting and mitigation. These three contexts give a brief but diverse view of the many ways to value nature, and of how using science to reveal these values can translate into actionable policy.

Acknowledgments The authors wish to thank the Gordon and Betty Moore Foundation, which supported their work during the researching and writing of this chapter, and Heather Tallis and Mary Ruckelshaus for valuable insight to inform its discussion.

Recommended Reading Daily, G., and K. Ellison. 2002. The new economy of nature: The quest to make conservation profitable. Island Press, Washington, D.C. Gaston, K., and J. Spicer. 2004. Biodiversity: An introduction. Second edition. Blackwell Publishing, Oxford, UK. Kareiva, P., H. Tallis, T. H. Ricketts, G. C. Daily, and S. Polasky. 2011. Natural capital: Theory and practice of mapping ecosystem services. Oxford University Press, Oxford, UK. Naeem, S., D. E. Bunker, A. Hector, M. Loreau, and C. Perrings. 2009. Biodiversity, ecosystem functioning, and human wellbeing. An ecological and economic perspective. Oxford University Press, Oxford, UK. National Research Council. 2004. Valuing ecosystem services: Toward better environmental decision-making. National Academies Press, Washington, D.C. Solan, M., R. J. Aspden, and D. M. Paterson. 2012. Marine biodiversity and ecosystem functioning: Frameworks, methodologies, and integration. Oxford University Press, Oxford, UK.

Glossary Ecosystem engineers  Species that exert their influence on ecosystem structure and functions through the physical formation or modification of habitat.

Ecosystem services  The benefits people receive from ecosystems. The value of ecosystem services can be considered a flow of benefits from nature to people. The four main types of ecosystem services are provisioning services (such as food, fuel, and fiber), regulating services (such as air and water purification, climate stabilization, and hazard mitigation), cultural services (such as spiritual, aesthetic and educational enjoyment of nature), and supporting services (such as soil formation, primary production, and nutrient cycling). Functional redundancy  When multiple species contribute to a particular ecosystem function in similar ways, so the addition or loss of another functionally similar species may have little effect. Gamma diversity  The total species diversity found across a landscape. Keystone species  Species that have a very large effect on community structure and function, compared to their biomass within the system. Mitigation offsets  The restoration of habitat or ecosystem functions to compensate for habitat or ecosystem services lost due to development. Natural capital  The stock of ecosystem services. Overyielding  When complementarity among species increases a community property such as biomass, the relative amount of variation in the system decreases, leading to greater stability with higher diversity. Payments for ecosystem services (PES)  A financial arrangement where beneficiaries of an ecosystem service pay landowners to manage their land in a way that will provide them with an ecological service. Portfolio effect  When ecosystem stability increases with diversity because fluctuations in species abundances through time are not perfectly correlated. Psychological ecosystem services  The benefits of human experience of nature to cognitive function, including concentration, impulse inhibition, short-term and working memory, and mood. Resilience  The ability of an ecosystem to bounce back from a stressor or disturbance. Serviceshed  The area that provides a specific ecosystem service to specific beneficiaries. Species richness  The number of species in a given area. Strong interactors  A species whose presence or absence has large effects on community structure, composition and community. Trophic cascade  When predators in a food web greatly reduce the abundance of their prey, minimizing predation on their prey’s prey.

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ECOSYSTEMS

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SIX TEEN

The Offshore Ecosystem STE VEN J. BOGR AD, ELLIOT T L . HA ZEN, SAR A M . MA X WELL , ANDRE W W. LEISING , HELEN BAI LE Y, and R ICHARD D. BRODEUR

Introduction The California Current System (CCS) is one of the world’s four eastern boundary upwelling systems, which are among the most productive ecosystems in the ocean. Prevailing northwesterly winds at the California coast drive a southward-flowing current that moves cold water to lower latitudes. These winds also drive surface water offshore, causing upwelling of colder, nutrient-rich subsurface water to replace it. This seasonal abundance of nutrients results in phytoplankton blooms at the base of a food web that supports an abundance of animal species from zooplankton to small fish (e.g., anchovy and sardines) up to large predators such as tuna, sharks, and whales. This rich marine ecosystem sustains California’s many important fisheries and contributes to its value for ecotourism. The highly dynamic nature of the CCS makes it difficult to strictly define its boundaries. In general, the CCS extends from the Transition Zone near 50°N, separating the Alaskan Gyre and North Pacific Subtropical Gyre, south to the subtropical waters off Baja California, Mexico (20°N–​25°N). Distinct biogeographical domains within the CCS have been proposed, with divisions at Cape Mendocino (40.4°N) and

Point Conception (34.4°N) (Parrish et al. 1981, Allen et al. 2006, Longhurst 1998). In the cross-shore direction, the CCS extends from the intertidal and coastal zones (see Chapters 17, “Shallow Rocky Reefs and Kelp Forests,” and 18, “Intertidal”) to hundreds of kilometers offshore into the North Pacific Subtropical Gyre. For this chapter we define the offshore region of the CCS as the area from 4.8 kilometers (3 miles) offshore, the outer boundary of state jurisdiction, to 200 kilometers offshore, which represents the western boundary of the federal Exclusive Economic Zone (Figure 16.1). We further focus on the waters offshore of the state of California. The CCS is one of the most thoroughly monitored and studied oceanic regions in the world (Bograd et al. 2003, Peña and Bograd 2007, Checkley and Barth 2009). One key research program in this area is the California Cooperative Oceanic Fisheries Investigations (CalCOFI), which has provided consistent monitoring of the physics, chemistry, and lower-trophic biology of the CCS (primarily the southern component) since 1949 (McClatchie 2014). Numerous other monitoring programs and process studies have been conducted by academic, federal, and state government scientists over the past 287

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FIGURE 16.1 Map of California Current Ecosystem (CCE) showing dominant currents and oceanographic features.

several decades (see summary in Checkley and Barth 2009). We begin with a review of the key taxa and species that compose the offshore component of the CCS ecosystem, from primary producers to top predators. We also describe some of the dominant processes of trophic interactions and ecosystem functioning within the offshore CCS. We then move to the human dimensions of the CCS, focusing on ecosystem services, conservation issues, and management structures. We conclude with a brief discussion of the future of the CCS offshore ecosystem and highlight potential climate-driven changes that could occur.

Offshore Ecosystems and Food Webs of the California Current System The oceanography of the CCS is characterized by strong physical forcing across multiple time scales and drives high biological productivity and a complex ecosystem structure (see Chapter 6, “Oceanography”). Its seasonal pulses of productivity lead to a high biomass of pelagic species that in turn support a diversity of top predators and economically important marine fisheries (see Chapter 35, “Marine Fisheries”). Many mobile species migrate seasonally over the latitudinal extent of the CCS (Horne and Smith 1997, Agostini et al. 2008, Checkley and Barth 2009, Block et al. 2011), while others migrate long distances across the Pacific to the CCS to exploit seasonally recurrent resources (Block et al. 2011). Geographic and bathymetric features (e.g., capes, islands, and submarine Photo on previous page: Offshore hotspot of shearwaters, common dolphins, and minke whales at sunset, southern California. Photo: Elliott L. Hazen. 288  Ecosystems

canyons) and oceanographic features (e.g., eddies and fronts) affect the distribution of many species (Checkley and Barth 2009). Vertically, pelagic habitat is defined as below the surface and above the bottom, but more specifically as the epipelagic (0–​2 00 meters, euphotic), mesopelagic (200–​1,000 meters), and bathypelagic (>1,000 meters bottom depth). In this section we describe the diversity of species occurring in the CCS from primary producers to top predators, their trophic interactions, and hotspots, focusing on the epipelagic and mesopelagic.

The base of the offshore food web is primary production by unicellular algae—​the phytoplankton (Kudela et al. 2008; Figure 16.2). The predominant phytoplankton groups within the California Current include three classes:

. Diatoms—​eukaryotic cells with hard silica–​based . .

shells, dominant in strong coastal upwelling areas, occasionally forming a harmful algal bloom. Dinoflagellates—​eukaryotic cells, many slightly motile, often dominant in stratified regions and more commonly forming harmful algal blooms than diatoms. Cyanobacteria—​prokaryotic cells, predominant in further offshore regions but still abundant in nearshore regions (about 20% of phytoplankton productivity).

Diatoms, which range in size from a few microns to chains of cells several hundred microns long, are probably the most critical phytoplankton contributors to overall productivity and as a food resource for higher trophic levels. Diatoms grow rapidly in nearshore regions where upwelling provides cool, nutrient-rich water. Diatoms thus typically dominate the phytoplankton biomass during strong upwelling. In turn, diatoms are grazed by secondary producers (i.e., the microzooplankton and mesozooplankton, described below, and certain small and larval fish). Occasionally, certain species of diatoms may form harmful algal blooms. Specifically, the diatom Pseudo-nitzschia multiseries produces the powerful neurotoxin domoic acid, which can bioaccumulate in the tissues of fish and potentially harm humans, marine mammals, and possibly seabirds (Kudela et al. 2005). Although diatoms are important prey for copepods, their protective silica casings (known as frustules) and larger size provide some protection from smaller microzooplankton, another factor allowing them to form large blooms. Dinoflagellates can outcompete diatoms under certain conditions when silica is limiting, because dinoflagellates do not require silica for growth. Many dinoflagellates are also somewhat motile and can swim to deeper waters at night to obtain nutrients when the water column is stratified and surface nutrient levels are low. Dinoflagellates, with their relatively enriched nutrient content and lack of hard silica encasements, are typically preferred over diatoms as a food source by other microzooplankton and small crustacean zooplankton (Kleppel 1993, Leising et al. 2005). Because of this, when dinoflagellates predominate, a longer chain of organisms often exists between phytoplankton and higher predators, reducing total energy transfer to higher trophic levels (only about 30% to 35% of energy is transferred upwards from each trophic level) (Paffenhofer 1976, Fenchel 1987). In

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FIGURE 16.2 Food web diagram for the California Current Ecosystem. Red represents the benthic ecosystem while blue represents the pelagic. The estimated trophic level is on the y-axis and the size of each box represents its relative biomass in the food web. Width of lines represents biomass flux of prey to predator. Source: Field et al. 2006.

contrast, in diatom-dominated systems the diatoms may be directly consumed by krill, copepods, and other zooplankton, which in turn are eaten by small fish and some fish larvae. Cyanobacteria are more important in offshore regions where, although they do not have a high biomass, they can have high growth rates, causing rapid nutrient turnover (Sherr et al. 2005). Cyanobacteria are consumed primarily by unicellular microzooplankton that may be prey for other microzooplankton. Thus far offshore food webs dominated by cyanobacteria, which already have low total productivity due to lower biomass compared to the nearshore, tend to have an even greater reduction of biomass at higher trophic levels due to the relatively large number of trophic links (see Figure 16.2).

include the following groups ordered approximately from smallest to largest by individual body size:

. Microzooplankton—​unicellular zooplankton that feed . . .

Primary Consumers Primary consumers are species that feed either primarily or partially on the primary producers (phytoplankton). They

.

at high rates on phytoplankton, other microzooplankton, and bacteria. Copepods—​smaller crustacean zooplankton, often the numerically dominant multicellular organism in many areas of the CCS that feed on phytoplankton, other zooplankton, and microzooplankton. Other crustacean zooplankton—​this group includes shrimps, mysids, and other less numerically dominant but important organisms that consume phytoplankton, other zooplankton, and microzooplankton. Gelatinous zooplankton—​soft-bodied zooplankton, such as jellyfish, ctenophores, pelagic gastropods (primarily pteropods), salps, doliolids, and appendicularians; chaetognaths can be important in some areas. Euphausiids—​also known as krill, relatively large, often swarm- or school-forming crustacean zooThe Offshore Ecosystem   289

.

.

plankton that feed on both phytoplankton and zooplankton. Ichthyoplankton—​small larval stages of fish that feed on both phytoplankton and zooplankton, including the larvae of the small pelagic fish, plus the larval stages of large pelagic fish and groundfish, such as Pacific hake, jack mackerel, and rockfish. Small pelagic fish—​includes baitfish and other forage fish, such as sardine, anchovy, and smelts, which are relatively small as adults and feed on phytoplankton and/or zooplankton.

Unicellular microzooplankton include a diverse array of organisms, such as heterotrophic dinoflagellates, ciliates, and choanoflagellates. These organisms primarily eat other microzooplankton, phytoplankton, cyanobacteria, and bacteria. The CCE biomass of unicellular microzooplankton is usually low; however, their grazing rates are comparable to the growth rates of phytoplankton (Li et al. 2011). Thus, contrary to common belief, it is these unicellular microzooplankton, not crustaceans or fish, that consume the majority of phytoplankton standing stock and production within many areas of the CCE (Calbet and Landry 2004). Particularly in the regions offshore of the main winddriven upwelling front, a large portion of the energy that flows into microzooplankton does not reach higher trophic levels but is returned to detrital pools or recycled within the microzooplankton trophic level (see Figure 16.2). This retention of energy within the unicellular microzooplankton trophic level is known as the “microbial loop” and, when prevalent, decreases the overall productivity of higher trophic levels (see review in Fenchel 2008). Unicellular microzooplankton are a key prey source for copepods, gelatinous zooplankton, and other small crustacean zooplankton due to their enriched nitrogen content compared to similarly sized phytoplankton. Copepods and other small crustacean zooplankton have similar roles to krill within the CCE (Figure 16.3). Unlike krill, copepods and small crustacean zooplankton do not tend to form large, dense schools, although for brief periods (a few hours to days) they can occur at locally higher densities when they aggregate near physical (e.g., horizontally along physical fronts or vertically near the main thermocline) or biological discontinuities (e.g., phytoplankton “thin layers”). Copepods, which often dominate the zooplankton numerically, eat phytoplankton, microzooplankton, and other smaller crustacean zooplankton, and in turn are food for krill, fish larvae, and small pelagic fish. Many of the larger, crustacean zooplankton undergo daily vertical migrations from as deep as several hundred meters during the day to near the surface at night, mainly to avoid visual predators such as fish (Enright and Hamner 1967, Hays 2003). This vertical migration could contribute significantly to carbon export from surface layers to depth, as copepods feed near the surface and then can die, be eaten, or produce fecal pellets while at depth (Stukel et al. 2013). Unlike many other zooplankton, several dominant species of copepods, those of the genera Calanus and Neocalanus in particular, undergo a wintertime dormant period wherein they descend to great depths (around 400 to 1,000 meters) for four to eight months of the year (Dahms 1995). They then emerge in the spring to reproduce. Copepods thus have marked seasonal variation in their availability to higher trophic levels, often leading to timing mismatch problems critical to the overall phenology of the CCE and potentially sensitive to climate change. Other 290  Ecosystems

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FIGURE 16.3 Photographs of dominant zooplankton species in the California Current. Photos: (A) NOAA Fisheries; (B) Moira Galbraith, Fisheries and Oceans Canada; (C) Russ Hopcroft, University of Alaska, Fairbanks; (D) Steve Haddock, Monterey Bay Aquarium Research Institute. A Calanus pacificus B Euphausia pacifica C Metridia pacifica D Thysanoessa spinifera

small crustaceans, such as shrimps and mysids, tend to be less abundant but can be important in some areas. Mysids often form swarms in shallow nearshore waters and can be an important food source for outmigrating smolts (Brodeur 1990). Crab larvae can also be seasonally important during their larval planktonic phase. When prevalent, gelatinous zooplankton provide an alternate pathway for energy flow that might or might not lead to production in higher trophic levels (Brodeur et al. 2011). Gelatinous zooplankton include a variety of forms, from freefloating jellyfish that passively ambush zooplankton and small larval fish prey, to appendicularians that build large gelatinous “houses” used to filter large quantities of the smallest phytoplankton from the water column. While gelatinous zooplankton grow and feed at high rates, their bodies are composed mostly of water. As a result, they are not typically a good food source for larger organisms, with the exception of certain turtles (e.g., leatherbacks, Dermochelys coriacea) and fishes (e.g., ocean sunfish, Mola mola) that specialize in gelatinous prey. Systems dominated by gelatinous zooplankton as the primary predators of phytoplankton tend to have limited fish production and are generally considered “dead-end.” An exception are pteropods—​pelagic gastropods that form large, gelatinous nets much larger than their body size, used to capture falling detritus in the water column. Unlike the other taxa in this group, pteropods are known to be an important food source for some salmon (Brodeur 1990) and possibly other fish species. Gelatinous zooplankton blooms (especially salps, a type of planktonic tunicate) can be found offshore in oligotrophic regions, while medusae and ctenophore blooms can be important nearshore during warmer periods. Euphausiids, primarily the species Euphausia pacifica and Thysanoessa spinifera, are another key link in the trophic web of the CCE (Brinton and Townsend 2003; see Figure 16.2, Figure 16.3). These species eat primarily phytoplankton (diatoms) and small zooplankton, and in turn provide food for many fishes, birds, and marine mammals. Euphausiids can

form large, conspicuous schools and swarms that attract larger predators, including baleen whales. Due to their high feeding and growth rates and key prey status for many species, euphausiids play a critical role in the overall flow of energy through the CCE. Ichthyoplankton, the larvae of larger fish and small pelagics, are a key resource for larger fish and other marine predators. Although only within the water column for a short period of time, from weeks to a few months, ichthyoplankton abundance may at times dominate the total abundance of secondary consumers, making them important grazers on phytoplankton and smaller zooplankton classes. Finally, small pelagic fish, such as sardine and anchovy, comprise an integral part of the CCE, feeding nearly exclusively on phytoplankton (typically diatoms), small pelagic crustaceans, and copepods (Emmett et al. 2005; Figure 16.4). This group, often termed the “forage” fish (discussed in more detail elsewhere in this chapter), functions as the main pathway of energy flow in the CCE from phytoplankton to larger fish and to the young life stages of larger predators (Crawford 1987, Cury et al. 2000; see Figure 16.2).

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FIGURE 16.4 Key forage and mid-trophic fish species in the California Current. Photos: (A) NOAA Oceanic and Atmospheric Research/ National Undersea Research Program; (B, C) NOAA Southwest Fisheries Science Center; (D) California Academy of Sciences. A Northern anchovy (Engraulis mordax) B Chilipepper rockfish (Sebastes goodei) C Pacific sardine (Sardinops sagax) D Coho salmon (Oncorhynchus kisutch)

The Small Pelagics Small pelagic fishes such as northern anchovy (Engraulis mordax), Pacific sardine (Sardinops sagax), Pacific mackerel (Scomber japonicus), and jack mackerel (Trachurus symmetricus) dominate the fish fauna of the epipelagic California Current. These schooling planktivores supported major fisheries in the past and in some cases still do (Allen and Cross 2006; see Chapter 35, “Marine Fisheries”). Their abundances vary dramatically over time, often peaking asynchronously with each other. These species spawn primarily in the Southern California Bight but are distributed along the entire California shelf region as juveniles and adults (Zwolinski et al. 2012). As in other upwelling ecosystems, they consume primarily smaller planktonic prey including phytoplankton, copepods, pteropods, decapod larvae, and small larval and juvenile euphausiids (van der Lingen et al. 2009), although larger mackerel can eat adult euphausiids and small fishes and squids. The ability of sardines and anchovies to filter feed places them at a relatively low trophic level and allows them to pass production on to higher trophic levels relatively efficiently. In addition to these dominant forage species, other regionally important fishes include Pacific pompano (Peprilus simillimus), Pacific bonito (Sarda chiliensis), and yellowtail (Seriola lalandi) in southern California, Pacific saury (Cololabis saira) in offshore waters, and Pacific herring (Clupea pallasi) and smelts (family Osmeridae) off central and northern California (Brodeur et al. 2003, Allen and Cross 2006, Harding et al. 2011). Another dominant species in terms of biomass and fishery harvest levels in the California Current is the Pacific hake (Merluccius productus), which is distributed widely along the coast of California and undergoes seasonal migrations from spawning grounds off southern California to feeding grounds off the Pacific Northwest and Canada (Ressler et al. 2007). This species inhabits midwater regions along the shelf and is important as prey (mainly in the larval and juvenile stage) and predator throughout the year (Livingston and Bailey 1985). A number of small squid species occur in the CCE, but off central and southern California the most important species by far is the market squid (Doryteuthis [formerly Loligo] opalescens). Although this species is relatively short-lived (lifes-

pan about one year), it is the mainstay of a pelagic fishery for several California ports (Zeidberg et al. 2006). Market squid feed mainly on copepods early in life, switching to euphausiids, small fish, and other squid as adults (Karpov and Cailliet 1978). They in turn become an important component in the diets of many fishes, seabirds, and marine mammals in coastal waters (Morejohn et al. 1978).

The Mesopelagics Mesopelagic species occur primarily at 200–​1,000 meter

depth and are present in all of the world’s oceans, but the difficulty of sampling them has led to uncertainty about their abundances and ecological roles (Brodeur and Yamamura 2005). Mesopelagic fish in particular are an important component of the deep scattering layer, and their sheer biomass make them important components of the open ocean food web globally. They inhabit depths below the photic zone (200–​1,000 meters) by day but exhibit diel vertical migration to the surface to feed at night, presumably to minimize predation risk while maximizing foraging capabilities (Robison 2004). Mesopelagic fish in the CCS include over twenty genera (Ahlstrom 1969), making them an extremely diverse component of the CCE. From the limited diet studies available, they are consumers primarily of zooplankton including mainly copepods and euphausiids but also some species of amphipods, ostracods, molluscs, and larvaceans (Mauchline and Gordon 1991, Suntsov and Brodeur 2008). They serve as important prey species for many large fishes, sharks, seabirds, Humboldt squid (Dosidicus gigas), and marine mammals (Pearcy et al. 1988; Fiedler, Barlow et al. 1998; Arizmendi-Rodriguez et al. 2006; Barlow et al. 2008; Field et al. 2012; Preti et al. 2012). Over the past fifty years, mesopelagic fish abundance in the CCS has been in decline, with a hypothesis that links it to the shoaling (movement to shallowed depths) of the oxygen minimum zone (OMZ) over the same time period (Bograd et al. 2008, Koslow et al. 2011). Given the sheer abundance of these organisms in the CCS, perhaps an order of magnitude The Offshore Ecosystem   291

greater than currently estimated (Kaartvedt et al. 2012), and their repeated presence in predator stomach contents, they are likely an integral but poorly understood part of the CCS food web and nutrient cycle (Davison et al. 2013).

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Top Predators The California Current is a hotspot for a high diversity and abundance of top predators because of the abundant prey supported by its seasonal upwelling and nutrient-rich waters (Figure 16.5). A number of top predators reside in the CCS year round, while even more travel vast distances to forage seasonally in the CCS (Block et al. 2011; Figure 16.6). Large pelagic fishes are abundant and support a number of fisheries, including salmon, rockfish, billfish, sharks, and a few species of tuna (Field et al. 2010, Block et al. 2011, Glaser 2011, Preti et al. 2012, Wells et al. 2012). Seabird species include local breeders and oceanic migrants, both of which rely on the CCS as their foraging grounds (Shaffer et al. 2006, Yen et al. 2006, Mills et al. 2007, Kappes et al. 2010). Six pinniped species breed on the coast of California, with many of these animals foraging in offshore waters (Antonelis and Fiscus 1980). A high diversity of cetacean species also occur in the region (Barlow and Forney 2007). These predators have all evolved strategies to benefit from the seasonal productivity of the CCS while minimizing interspecific competition.

F E

FIGURE 16.5 Key top predators in the California Current. Photos: (A, E, F) Elliott Hazen, NOAA; (B) Mark Conlin, NOAA Southwest Fisheries Science Center; (C) Dan Costa, UC Santa Cruz; (D) NOAA. A Black-footed albatross (Phoebastria nigripes) B Blue shark (Prionace glauca) C Northern elephant seal with electronic tag (Mirounga angustirostris) D Humpback whale (Megaptera novaeangliae)

FISHE S

E Short-beaked common dolphin (Delphinus delphis) F Transient killer whales (Orcinus orca)

Salmon species in California include coho (Oncorhynchus kisutch) and Chinook salmon (Oncorhynchus tshawytscha), both populations federally endangered (see Chapter 35, “Marine Fisheries”). Adult salmon serve as important predators in the pelagic ecosystem (Thayer et al. 2014) and support economically valuable fisheries. Juvenile salmon are important predators of euphausiids and small fish. The largest Chinook runs are in the Central Valley (Sacramento–​San Joaquin River system) and Klamath River. Chinook salmon inhabit a relatively narrow temperature range in the CCS. This may underlie observed connections between ocean conditions and salmon survival (Hinke et al. 2005). The first year at sea for Chinook salmon is thought to be the most sensitive, with surviving individuals returning to spawn as adults four to seven years after hatching (MacFarlane 2010, Wells et al. 2012). Because of their ties to ocean conditions, both salmon species are often used as indicators of ecosystem status, with spawning returns (escapement) providing a time series of salmon abundance in the CCS back to 1970 (Lindley et al. 2009, Levin and Schwing 2011). A 2009 collapse in Central Valley Chinook was attributed to a combination of a historical reduction in the population’s diversity of run timings and life history strategies caused by water competition with human uses, and poor ocean conditions for the yearling stage in 2004 and 2005 (Lindley et al. 2009). This event illustrated how salmon are intimately tied to both land and sea conditions and vulnerable to changes in both. Adult rockfishes are a complex of over seventy species, many of which are late maturing and extremely long-lived (Love et al. 2002, Mills et al. 2007, Field et al. 2010). They support important fisheries, both recreational and commercial (see Chapter 35, “Marine Fisheries”). The deep-water rockfishes found in offshore environments are particularly sensi292  Ecosystems

tive because they are the longest-lived and suffer the greatest proportional mortality when brought to the surface (Love et al. 2008). Long-term fishing pressure appears to have resulted in changes in community composition (Love et al. 2008) and decreases in fish size (Mason 1998). Similar to salmon survival, rockfish growth has been correlated with winter upwelling, indicating that some rockfish species are highly dependent on the early physical processes in the CCS (Black et al. 2011). Their role as predators given fishery depletion is less clearly quantified but could be important, as they feed primarily on pelagic forage fish. Both juvenile salmon and rockfish serve as important forage fishes in the CCS and are found in the diets of many larger fish, seabirds, marine mammals, and even Humboldt squid (Mills et al. 2007, Field et al. 2010, Wells et al. 2012). A number of species of tuna, billfish, and sharks use the CCS as a seasonal foraging ground, relying on the upwelling dynamics to provide an abundance of food. Albacore (Thunnus alalunga), yellowfin (T. albacares), skipjack (Euthynnus pelamis), and Pacific bluefin tuna (T. orientalis) are all found off the coast of California. Albacore and skipjack use the entire U.S. West Coast, while bluefin and yellowfin occur primarily in the central and southern CCS, respectively (Block and Stevens 2001). Tuna specialize on forage fishes such as anchovy, Pacific sardine, Pacific saury, and squids as adults (Pinkas 1971, Bernard et al. 1985, Glaser 2011). North Pacific albacore could provide significant top-down pressure on anchovy populations previously believed to be largely regulated by the physical forcing and upwelling dynamics (Glaser 2011). Swordfish (Xiphias gladius) and striped marlin (Kajikia audax) are the two main billfish species in the CCS and are ocean migrants (Bed-

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Longitude FIGURE 16.6 Fidelity and attraction to the California Current Ecosystem (CCE). Source: Block et al. 2011. A Examples of pelagic predators released and electronically tracked in the CCE that show fidelity to deployment locations and the CCE.

We show the release locations (square), pop-up satellite end-point locations (triangle), and daily mean positions (circles) of the following species: yellowfin tuna (yellow), bluefin tuna (white), white shark (red), elephant seal (blue), and salmon shark (orange). B Individual tracks of pelagic animals released 2,000 kilometers away from the CCE indicate cross-basin or ecosystem attraction to, and

temporary residency within, the eastern North Pacific. Symbols as in (A); for leatherback sea turtles (green), sooty shearwaters (pink), fur seals (pale yellow), black-footed albatrosses (black), and salmon sharks (orange).

ford and Hagerman 1983). Swordfish and striped marlin are both known to target forage fishes and cephalopods, based on diet studies in the tropical Pacific (Markaida and Hochberg 2005). The high energetic demand of these pelagic predators has made them susceptible to the OMZ in the tropical Atlantic (Stramma et al. 2011), and it is plausible that the shoaling OMZ in the CCS may also compress their pelagic habitat (Bograd et al. 2008, Koslow et al. 2011). Tagging data from bluefin, albacore, and yellowfin tuna show high site fidelity to the CCS with seasonal migrations from the southern CCS northward as temperatures increase (Block et al. 2011). Shark species occupy a range of foraging niches in the CCS, from planktivore to scavenger to top predator. Much research on sharks has stemmed from fisheries bycatch or directed tagging studies. Basking sharks (Cetorhinus maximus) that feed on zooplankton aggregations were once abundant in the CCS, but fisheries harvest and directed mortality has reduced them to a remnant population with little remaining ecological impact (Ebert 2003). Blue (Prionace glauca), salmon (Lamna ditropis), shortfin mako (Isurus oxyrinchus), and white sharks (Carcharodon carcharias) show significant niche partitioning in the CCS. Blue, mako, and white sharks primarily forage in the CCS, while salmon sharks likely use it mostly as a pupping ground (Block et al. 2011). Salmon sharks migrate throughout the eastern North Pacific, seasonally taking advantage of salmon, pollock, and herring in the northern Gulf of Alaska in the summer through the winter and giving birth in the spring off the California coast (Weng et al. 2008, Carlisle et al. 2011). Thresher (Alopias vulpinus), mako, and blue sharks are commonly caught by fisheries, providing insights into their distribution and diet. Thresher sharks feed primarily on pelagic forage fish including anchovy, sardine, and hake, with a smaller contribution from squid species. Mako and blue sharks feed on Humboldt squid (Dosidicus gigas), particularly when they

are abundant in the CCS (Vetter et al. 2008, Camhi et al. 2008, Preti et al. 2012). Blue sharks consume both pelagic fish and cephalopods. White sharks use the CCS most heavily during the pinniped pupping season, then move offshore to an area nicknamed the “White Shark Café,” where they exhibit very different diving behaviors than in the CCS and may be foraging and/or mating (Camhi et al. 2008, Jorgensen et al. 2010, Jorgensen et al. 2012). When farthest offshore, many of the shark species likely supplement their diet with deep scattering layer organisms, such as mesopelagic fishes and squid (Jorgensen et al. 2012).

SE ABIRDS

Nearly 150 seabird species occur in the CCS. These include locally breeding species, such as auklets and murres, and long distance migrants, such as albatrosses and shearwaters. These species all use the CCS as a foraging ground. A number of local species have become key indicators of climate variability because their breeding success is closely tied to CCS upwelling regimes (Black et al. 2011). A late start to the upwelling season in 2005 led to unprecedented and complete breeding failure in Cassin’s auklet (Ptychoramphus aleuticus) that was also apparent but less extreme in other top predator taxa (Sydeman, Bradley et al. 2006). Further research has confirmed that upwelling is critical for these seabirds’ reproductive success and, more specifically, that winter conditions and early season upwelling are most important (Schroeder et al. 2009, Black et al. 2011). Weak upwelling combined with less stratified waters than usual can provide nutrients to jump start the food web at critical times in seabird life histories (Schroeder et al. 2009). Summer upwelling was closely correlated with Cassin’s auklet fledgling success, while common murre (Uria aalge) fledgThe Offshore Ecosystem   293

ing success and both species’ egg laying dates were associated with the anomalous winter upwelling mode (Black et al. 2011). Prey availability near seabird colonies is the proximate cause of local seabird breeding success, with changes in key prey species such as juvenile rockfish and krill directly influencing breeding success in a number of bird species (Ainley et al. 1995, Sydeman et al. 2001). The proportions of juvenile rockfish found in the diets of three species—​common murre, pigeon guillemot (Cepphus columba), and rhinoceros auklet (Cerorhinca monocerata)—​have been directly related to their reproductive success, highlighting their importance as forage fish in the system (Ainley et al. 1995, Mills et al. 2007). In more recent studies, however, prey switching and conservative life history strategies (laying at most a single egg per clutch) seem to have buffered these three species from the effects of juvenile rockfish declines of 50% to 75% (Field et al. 2010). Cassin’s auklet breeds later, with lower fledging success, in warmer years such as during El Niño events, indicating that long-term climate change could result in later breeding and reduced recruitment for a number of these indicator species (Sydeman et al. 2009). Many highly migratory bird species, including sooty shearwaters (Puffinus griseus), blackfooted albatross (Phoebastria nigripes), and Laysan albatross (Phoebastria immutabilis), depend upon the CCS as primary foraging grounds. These migrants depend highly on wind patterns to reach their foraging grounds but can spend over half of the year in the CCS. Sooty shearwaters are one of the world’s greatest migrants, traveling from New Zealand breeding grounds to the CCS foraging grounds in April through October (Shaffer et al. 2006) where they feed on an abundance of fish, squid, and krill. They can also migrate to other foraging grounds as productivity wanes in the CCS or depart early when the CCS is affected by anomalous conditions such as delayed upwelling or El Niño events. The two albatross species breed in the northwest Hawaiian Islands, with an additional Laysan albatross colony on Guadalupe Island off Baja California, Mexico (Pitman et al. 2004, Kappes et al. 2010). During the incubation period, both species migrate to the CCS, although Laysan albatross use the western Pacific more heavily and are less common in the CCS (Kappes et al. 2010). Both species are largely diurnal scavengers on fish and squid, and their migration patterns are less constrained during the incubation period than during chick-rearing. This allows them to better adapt to both interannual and decadal variability (Pitman et al. 2004, Kappes et al. 2010). Several seabird species, including the blackfooted albatross, red pharalope (Phalaropus fulicaria), and Leach’s storm petrel (Oceanodroma leucorhoa), are associated with mesoscale features in the CCS, such as topographic or upwelling fronts, that could be a response to enhanced local productivity (Yen et al. 2006).

SE A TURTLE S

No sea turtles nest in California, although leatherback turtles (Dermochelys coriacea), a small population of green turtles (Chelonia mydas), and occasionally loggerhead turtles (Caretta caretta) forage in the CCS. The green turtles occur only in San Diego Bay and breed in Mexico (MacDonald et al. 2012). Loggerheads are common off the Baja California peninsula and present in low numbers off California during El Niño years (Carretta et al. 2004, Wingfield et al. 2011). Leatherback turtles migrate to the CCS foraging grounds from their breeding 294  Ecosystems

grounds in the tropical western Pacific (Benson et al. 2011, Block et al. 2011, Bailey et al. 2012). They will also return in multiple years to these CCS foraging grounds for the summer and autumn, then migrate to low-latitude, eastern tropical Pacific wintering areas without returning to their nesting beaches. Three main high-use areas in the CCS have been identified for leatherbacks (Benson et al. 2011). The first is off central California, where relatively cool water over the coastal shelf is characterized by high levels of chlorophyll-a concentration, indicative of high productivity. The second is in offshore waters off central and northern California, where sea surface temperature fronts occur in early summer. Finally, the third is on the continental shelf and slope off Oregon and Washington, particularly in the area surrounding the Columbia River Plume, which supports seasonally high abundances of the turtles’ gelatinous zooplankton prey (Benson et al. 2011). Leatherback turtles are critically endangered and have recently been named California’s marine reptile due to their dependence on CCS resources. Beach development at nesting grounds outside the CCS and bycatch in fisheries could both be hindering its recovery (Carretta et al. 2004, Benson et al. 2011, Tapilatu et al. 2013).

CETACE ANS

Twenty-one species of cetaceans, including odontocetes and baleen whales, have been sighted off the California coast during marine mammal surveys since 1991 (Barlow and Forney 2007). The development of species-habitat models from these survey sightings and environmental data have revealed that the distribution of many cetacean species can be explained by both temporally dynamic variables, such as sea surface temperature and its variance, and more static variables, such as water depth and seabed slope (Becker et al. 2010, Forney et al. 2012). Odontocete species (toothed whales) include warm water, cool water, and cosmopolitan species, and their sighting rates vary based on the oceanographic regime, such that during El Niño conditions warm water species are much more common in the CCS (Barlow and Forney 2007). Some of the rarest and deepest diving species, beaked whales, have declined in abundance in the CCS since 1991, which could reflect increased anthropogenic use of the offshore environment or broaderscale ecosystem changes (Moore and Barlow 2013). Beaked whales are known to forage on mesopelagic fish and squid (Moore and Barlow 2013). Another deep diver that also forages on squid, the sperm whale (Physeter macrocephalus), did not show an abundance decline over the 1991–​2 005 survey period (Barlow and Forney 2007). Baleen whales, characterized by baleen plates for filtering their food, feed on pelagic forage fish and krill (Croll et al. 2005, Barlow et al. 2008, Burrows et al. 2012). Although baleen whales have a 2.5-fold greater biomass than odontocetes, their primary production requirement in the CCS is only 13% of that required by odontocetes because they feed at a lower trophic level (Barlow et al. 2008). Occurrences of baleen whales in the CCS are affected by the distribution and abundance of dense prey aggregations that they target (Croll et al. 1998; Fiedler, Reilly et al. 1998). Sightings data have revealed important patterns in interannual variability, such as decreased presence of blue whales (Balaenoptera musculus) in the CCS following the anomalous upwelling year of

2005, presumably due to poor krill recruitment (Barlow and Forney 2007). Integrative studies in central California have shown that increased krill density lags seasonal upwelling and increased productivity by three to four months and that blue whales target the densest patches of krill in the California vicinity before moving to other foraging hotspots in the northeast Pacific (Croll et al. 2005, Bailey et al. 2009, Calambokidis et al. 2009). Marine mammals in Monterey Bay generally moved nearshore in El Niño conditions to more productive waters, while anomalous upwelling conditions led to migration to other, less-affected foraging areas (Burrows et al. 2012). Many large whale populations are still recovering from harvest, suggesting they might have played a greater role in the ecosystem before the historical whaling era. Recent research suggests that Blue whales (Balaenoptera musculus) have recovered to their carrying capacity since the cessation of whaling (Monnahan et al. 2015), although human impacts including noise and ship strikes may still affect the Eastern Pacific population (Goldbogen et al. 2013, Irvine et al. 2014).

PINNIPEDS

Six species of pinnipeds breed on the California coast, many of which use the CCS as their foraging grounds (Antonelis and Fiscus 1980). Two of the most abundant offshore visitors are the California sea lion (Zalophus californianus) and the northern elephant seal (Mirounga angustirostris). California sea lions are extremely abundant in the CCS, and typically forage opportunistically on fish, decapods, and cephalopods on the continental shelf (Weise et al. 2006). Their plasticity in foraging strategy is apparent in anomalous years. For example, they foraged up to 600 kilometers offshore in 2005 during a delayed upwelling event compared to 100- kilometer migrations in normal years. Sea lion pup mortality has been directly linked to productivity, with higher mortality during El Niño regimes (Sydeman and Allen 1999). Northern elephant seals have one of the richest tagging histories of any top predator and are central place foragers returning to California beaches to breed and molt (Robinson et al. 2012). Most female northern elephant seals move far offshore to the transition zone to feed, but a significant portion exhibit a distinct and more nearshore strategy (Robinson et al. 2012). Males tend to migrate northward, foraging along the continental margin from Oregon to the western Aleutian Islands (Le Boeuf et al. 2000).

Trophic Interactions and Ecosystem Functioning The CCS food web is driven by variability in upwellingdriven, bottom-up processes (Checkley and Barth 2009). Once upwelling has kick-started primary production, nutrients drive a pelagic food web highly dependent on the phytoplankton base, with copepods, euphausiids, forage fish, and hake serving as key pathways of biomass to higher trophic levels in the northern California Current (Checkley and Barth 2009; see Figure 16.2). The importance of forage fish in the pelagic food web suggests potential wasp-waist dynamics, in which the food web is strongly controlled both up and down by these midtrophic organisms. Similar food web structure occurs in other eastern Pacific boundary currents (Cury et al. 2000). However, recent studies have challenged the waspwaist hypothesis. For example, recent food web models suggest that alternative trophic pathways are available in the CCS when a key forage species is depleted (Fréon et al. 2009). Stable

isotopes have also shown that many top predators are more plastic than once believed and can switch prey types and even trophic levels based on what is available (Madigan et al. 2012). The diversity in midtrophic species in the CCS may also buffer the ecosystem from wasp-waist dynamics compared to systems where a single species of forage fish dominates. While fishing impacts have had an effect on forage fish in the CCS (Essington et al. 2015), the variability in seasonal upwelling is equally, if not more, important in regulating forage fish and the ultimate dynamics of the system (Fréon et al. 2005). The fact that juveniles of a high trophic level species can serve as an important forage fish could also provide alternate food web pathways when other forage fish are less prevalent (Field et al. 2010). As in other eastern Pacific boundary currents, the large biomass of lower trophic level species supports a high diversity of large predators, including fishes, seabirds, and marine mammals (Checkley and Barth 2009). The strong influence of bottom-up forcing through climate variability in the CCS (Ware and Thomson 2005) is highlighted by the range expansion of Humboldt squid (Field et al. 2007). Following the 1997–​1998 El Niño, Humboldt squid were observed in unprecedented numbers from central California up to British Columbia, Canada (Zeidberg and Robison 2007, Field et al. 2007, Field et al. 2012). Initial hypotheses included a combination of environmental changes, including shoaling low oxygen and warming temperatures, and population release following overfishing of sharks and tuna that could have reduced predation pressure on this species (Zeidberg and Robison 2007). However, the timing mismatch between fisheries pressure and the expansion indicate that environmental mechanisms, combined with a fast migration by the squid into the CCS, was the more likely driver (Watters et al. 2008, Stewart et al. 2012). Humboldt squid are important predators of mesopelagics and forage fish as well as a large portion of the diet of many top predators in the CCS (Field et al. 2012, Preti et al. 2012). A concurrent decline of Pacific hake during Humboldt squid expansion could indicate competition with or predation of this species (Zeidberg and Robison 2007). The shoaling OMZ, which has compressed the habitats of numerous top predators with high energetic demand, could also provide a new ecological niche for expansion of the hypoxia-tolerant Humboldt squid (Stramma et al. 2010, Hoving et al. 2013). The dynamics of these diverse trophic groups underscore how both the timing and strength of upwelling can have large effects on the CCS (Bograd et al. 2009). The drastic effects of the spring transition’s timing on CCE food webs and production are highlighted by the events of the 2005 upwelling season. In that year, upwelling was delayed by up to several months, producing warmer waters, lower nutrients, fewer lipid-rich copepods, and failed recruitment of fish and seabird species (Brodeur et al. 2006; Mackas et al. 2006; Sydeman, Bradley et al. 2006). Anomalous wintertime upwelling can lead to greater recruitment of seabird species and rockfish growth and higher salmon ocean survival, while strong summer events are important for other seabird species and for salmon returns (Wells et al. 2008, Black et al. 2011, Schroeder et al. 2013). These differential responses to upwelling highlight the complexity of marine species interactions and dynamics and offer insight into potential long-term changes as the North Pacific warms. Broad-scale climate variability including ENSO (El Niño) events, the NPGO (North Pacific Gyre Oscillation), and the PDO (Pacific Decadal Oscillation) also drive major responses in the CCE, arguably including some irreversible regime shifts (Ohman et al. 2013). During El Niño events, warmer surface The Offshore Ecosystem   295

waters and deeper thermoclines result in lower CCS productivity. Two key forage fish (anchovy and sardine) exhibit strong, decadal-scale variation in ecosystem dominance, total abundance, and recruitment (Barange et al. 2009, Checkley and Barth 2009). In warm years, sardine dominate the CCS in both surveys and top predator diet, while anchovy are more dominant in cooler years (Barange et al. 2009). Debate persists about the relative roles of climate and fishing effects on suppressing forage fish populations, but fished species exhibit greater sensitivity to climatic variability than their unfished counterparts (Hsieh et al. 2008, Barange et al. 2009, Essington et al. 2015). Highly migratory predators can often shift their behavior and distribution during El Niño years in response to lower trophic level processes, but extreme events can result in poor recruitment or juvenile survival among these migratory predators (Sydeman and Allen 1999, Benson et al. 2002, Weise et al. 2006). Since the early 1970s, variance in broadscale indices like the PDO and ENSO has remained constant, but the NPGO has shown increasing variance since 1985 (Sydeman, Santora et al. 2013). This increased variance may propagate through the CCE, particularly for climate-sensitive indicator species (Sydeman, Santora et al. 2013). Some extreme events, such as the 1998 El Niño, may have led to a shift in climate forcing from a warm, low productivity regime to a cool, highly productive regime (Peterson and Schwing 2003). Concurrent changes occurred in species communities and perhaps even ecosystem functioning, but it is difficult to determine mechanistically whether these constitute a cycle or a regime shift (Peterson and Schwing 2003, Overland et al. 2008). The differential responses of marine species to CCS processes highlight the need for multiple indicators of ecosystem state, as well as composite indicators that combine physical forcing and species, to ensure effective management of the CCE that incorporates both anthropogenic use and climate variability and change (Levin and Schwing 2011).

Spatial Distributions Spatial features and temporal processes can generate marine hotspots—​predictable and persistent areas of productivity or aggregation of lower trophic level organisms with greatly increased trophic flow and ecosystem importance (Sydeman, Brodeur et al. 2006; Hazen et al. 2013). Here we discuss processes that can create marine hotspots and identify recurrent hotspots in the CCS. Bathymetric features such as seamounts, shelf breaks, or islands can generate increased upwelling of nutrients and retention of forage species (Reese and Brodeur 2006). A high diversity of top predators relies on the seasonal variability and productivity of the CCS. These top predators aggregate based on specific oceanographic and bathymetric features to forage (Block et al. 2011). Based on a decade of tracking data, many top predator species use the entire CCS as a regional hotspot, with use patterns seasonally shifting northward with increasing productivity as temperature rises in the Southern California Bight (Block et al. 2011). Mesoscale features such as eddies and fronts can result in marine hotspots when increased productivity is entrained in an eddy or the mixing of two water masses increases productivity and aggregation (Logerwell and Smith 2001, Palacios et al. 2006, Yen et al. 2006). For example, mesoscale eddies created from meanders of the California Current form important hotspots for top predators including blackfooted alba296  Ecosystems

tross, red phalaropes, Leach’s storm petrel, and elephant seals (Yen et al. 2006, Robinson et al. 2012). Upwelling can directly produce hotspots, particularly when particular locations have persistent and/or stronger upwelling than surrounding areas (Palacios et al. 2006 and references within). In the CCS, Palacios et al. (2006) identified three coastal hotspots driven by upwelling: Cape Mendocino to Point Arena, Bodega Head to Point Sur, and Cape San Martin to Point Arguello. Chlorophyll-a persistence indices corroborated these locations as productive for a large portion of the year (Suryan et al. 2012). Additional work in the upwelling region off Point Conception in southern California showed that the high chlorophyll-a concentrations in turn supported krill (Santora et al. 2011) and that the same areas are important for foraging by seabirds (Yen et al. 2006) and blue whales (Bailey et al. 2009). Krill hotspots generally are also located near known upwelling centers, but krill avoid the regions of strongest Ekman transport where they would be advected offshore and instead are more strongly associated with areas of retention (Santora et al. 2011). Two critical forage species, the euphausiids (krill) E. pacifica and T. spinifera, are distributed throughout California waters but form patchy aggregations where they serve as important food resources for top predators (Croll et al. 2005, Santora et al. 2011). Blue whale foraging hotspots closely overlap with krill hotspots, particularly in these regions (Croll et al. 1998; Fiedler, Reilly et al. 1998; Bailey et al. 2009; Santora et al. 2011). One of the best studies decomposing a seasonal hotspot examined upwelling, phytoplankton blooms, krill aggregations, and blue whale foraging in Monterey Bay (Croll et al. 2005). After the start of the upwelling season, when wind-driven upwelling provides nutrients to the photic zone, phytoplankton blooms occurred six to ten days after upwelling events. Densities of both T. spinifera and E. pacifica adults were greatest in late summer a few months after the peak upwelling. Blue whales tended to exploit the lateseason patches of krill found along the edge of the canyon in Monterey Bay, taking advantage of the increased productivity from submarine upwelling but also using the shelf break as a buffer from currents to minimize energetic costs (Croll et al. 2005). These hotspots recur seasonally, with some interannual variability (Croll et al. 2005, Bailey et al. 2009). The Gulf of the Farallones and the waters surrounding the Channel Islands are two important foraging areas for marine predators, particularly seabirds and pinnipeds. Both areas provide terrestrial haul-out and nesting habitats near a shelf break and corresponding area of increased productivity (Ainley and Lewis 1974, Sydeman et al. 2001, Hyrenbach and Veit 2003, Carretta et al. 2009). Finally, areas of curl-driven upwelling, higher salinity, and sea surface temperature values between 12°C and 16°C were important in defining regional hotspots for sardine spawning aggregations, while finer-scale hotspots important for foraging predators are less well understood (Weber and McClatchie 2010, Zwolinski et al. 2011). Finally, spawning anchovy regional hotspots were best predicted using depth of the chlorophyll maximum and geostrophic flow (Weber and McClatchie 2010). Both of these species exhibit seasonal migrations along the coast with patchy aggregations, making their distributions potential mobile hotspots for foraging predators. These concentrations of high trophic transfer indicate important hotspots and potentially conservation areas, yet they can move or wane with seasonal, interannual, and even long-term changes in ecosystem dynamics, making adaptive and dynamic approaches critical for long-term management of these hotspots in the CCS

(­Palacios et al. 2006, Žydelis et al. 2011, Hazen et al. 2013, Scales et al. 2014, Maxwell et al. 2015).

Ecosystem Services, Threats, and Management Structure in the Offshore California Current In the CCE, ecosystem services include provisioning services, such as food and water; regulating services, including regulation of climate, wastes, and water quality; and additional services such as transportation via shipping and oil and gas extraction. Habitats that provide nursery areas and foraging grounds support commercial and recreational fisheries. Finally, the CCE provides important cultural services, such as ecotourism. Although the CCE provides many ecosystem services, human activities pose threats to these marine resources through stresses that include pollution and habitat degradation. We focus here on three major services in the offshore California Current: carbon sequestration, shipping, and oil and gas production. We also discuss threats from the shipping and oil and gas industries and their potential impact to the offshore ecosystem.

Services Provided by Offshore Ecosystems CARBON SEQUE STR ATION

The offshore ecosystem plays a large part in modulating a number of atmospheric processes. Key among these is regulating the amount of carbon dioxide in the atmosphere. Through the “biological organic carbon pump,” photosynthetic organisms in the upper layers of the ocean intake carbon. As they die, this carbon is sequestered in the deep ocean as the photosynthetic organisms sink toward the ocean floor. The ocean is responsible for removing approximately 48% of human carbon emissions (Sabine et al. 2004) and is the second largest sink for anthropogenic carbon after the atmosphere itself (Riebesell et al. 2007). Given increasing anthropogenic production of carbon dioxide and resulting climate change, this ecosystem service is of increasing importance. Oceanic carbon sequestration is not, however, without consequences (see the “Ocean Acidification” section later in this chapter).

SHIPPING AND CRUISE LINERS

California waters are a major economic driver of California and the entire U.S. Shipping and cruise industries provide millions of U.S. jobs for millions of people, billions of dollars in tax revenue, and billions of dollars to the U.S. economy. Transportation via container ships is one of the largest services provided by the CCE, facilitating massive domestic and, particularly, international trade. California is home to some of the largest ports in the U.S., including the ports of Long Beach, Los Angeles, and Richmond—​a ll in the top fifty U.S. ports by total trade volume for 2011 (AAPA 2012). Shipping is considered a more environmentally sound means of transporting goods than road, train, and air (Butt 2007). The cruise industry in the U.S. is another major economic driver and employed over 350,000 people in the U.S. in 2011, including many at the ports of San Francisco, Los Angeles, Long Beach, and San Diego.

OIL AND GAS PRODUCTION

Oil and gas production has a long cultural history in terrestrial systems in California, but 16% of the state’s extraction occurs offshore of California. California was the location of the world’s first offshore oil development, with the first well placed offshore of Santa Barbara in 1896. Production quickly climbed, with over 180 wells placed offshore of Santa Barbara County by 1902, and with quick expansion south to Orange, Los Angeles, and Ventura Counties. Regulation of the industry followed in 1921 (McCrary et al. 2003). Despite its long history in California, offshore oil and gas activity in the state contribute a small proportion of the country’s production. The Bureau of Ocean Energy Management (BOEM) reports that in the Pacific region, which includes the offshore regions of Washington, Oregon, and California, 43 of the 49 active leases were actively producing an average of 61,113 barrels of oil and 113 million cubic feet of gas per day in 2009 (BOEM 2013). This amounts to 1.24 billion barrels of gas and 1.67 trillion cubic feet of gas per year, which is less than 1% of the total daily production of oil and gas in the U.S. (U.S. Energy Information Administration 2013, BOEM 2013). Six companies operate the twentythree oil and gas platforms along the U.S. West Coast, nine of which are active in California, in the southern California region between Santa Barbara and San Pedro Bay/Long Beach Harbor (BOEM 2013). By comparison, the Gulf of Mexico has over four thousand active offshore platforms (McCrary et al. 2003).

Ecosystem Threats The scale and continued rise of the shipping and cruising industries along the U.S. West Coast, particularly in California, pose a number of threats to the larger ecosystem. Key among those affecting the offshore ecosystem are noise emitted by the oil and gas industry and military, discharge of pollutants (such as oil), risk of collisions, transport of invasive species from distant regions, and pollution from trash and sewage discards.

NOISE

Understanding of noise impacts on marine organisms remains poor, and noise threats to cetaceans has been of particular concern (Nowacek et al. 2007). Sound is the primary sense used by cetaceans for a number of life functions, and many cetaceans in the CCE use low-frequency sounds to communicate (Nowacek et al. 2007, Weilgart 2007, Clark et al. 2009). The baleen whales such as blue, fin, humpback, and gray whales, communicate over long distances of approximately 100 kilometers (Payne and Webb 1971, Payne and McVay 1971). Although much about the significance of these sounds is unclear, we know that some of the sounds, also known as “songs,” play critical roles in socialization and mating (Darling and Berube 2001, Simon et al. 2010). While mysticete whales do not echolocate like dolphins, they may still use these low-frequency sounds for navigation, as evidenced by the wide berth they take around ice floes even when the floes are out of visual range (George et al. 1989). Ship traffic in California has increased exponentially over the past sixty years (McDonald et al. 2006). CommerThe Offshore Ecosystem   297

cial vessels emit loud (180–​195 dB re 1μPa), low-frequency (10–​500 Hz) underwater sound. This noise can impair cetaceans’ ability to detect relevant noises (“masking”), affecting the hearing threshold or behavior of the animals (Nowacek et al. 2007). This in turn can impair their ability to navigate, detect prey, and communicate with conspecifics (Clark et al. 2009). Some noise impacts are acute, with behavioral or physiological impacts. For example, an animal’s hearing might be damaged, or individuals might leave an area that may be important for feeding or breeding (Weilgart 2007). Chronic impacts, often in the form of masking, can cause stress and lead to effects on individual or reproductive fitness (Weilgart 2007, Clark et al. 2009). Animals might attempt to change vocalization patterns in response to masking by changing vocalization frequencies, volume, and timing (Weilgart 2007, Parks et al. 2011, Castellote et al. 2012, Rolland et al. 2012). The effects of chronic noise impacts are particularly hard to detect at a population level, although reduction in noise levels has been shown to significantly decrease stress levels in North Atlantic right whales (Rolland et al. 2012). Similar noise impacts are likely to occur to other marine taxa such as diving birds, fish, reptiles, and invertebrates that use sound, either passively or directly, to communicate, detect prey, or navigate (Wysocki and Ladich 2005, Codarin et al. 2009, Popper and Hastings 2009, Simpson et al. 2010). For example, some fishes respond to increased noise by increasing vocalization rates, with unknown consequences on fitness (Picciulin et al. 2012). Brown shrimp (Crangon crangon) exposed to sound levels 30 dB above ambient noise had decreased growth and reproductive rates (Lagardere 1982). Very few studies have investigated the impacts of chronic noise on species beyond cetaceans. Sounds produced by the oil and gas industry, such as during geological and geophysical surveys used to identify new potential hydrocarbon resources, tend to be loud (>200 dB re 1μPa) and to include mid- to high frequencies (>1 kHz) and acute noises. This is in contrast to commercial vessels, which produce low-frequency chronic noise (Nowacek et al. 2007). In particular, seismic surveys for oil and gas fields often produce loud (220–​255 dB re 1μPa) sounds that expand across a range of frequencies (>300 Hz to over 2 kHz). Baleen whales are considered the most sensitive to seismic surveys. Effects from these acute noises fall into behavioral, acoustic, and physiological categories (Nowacek et al. 2007). Mitigation measures, such as soft starts to operations that gradually increase the sound, should help to reduce the likelihood of physiological impacts such as hearing damage and threshold shifts. Most observed responses have been behavioral, with animals displaced from previous habitat more than 30 kilometers away and sometimes rapidly changing course to avoid survey areas (Richardson et al. 1995, Nowacek et al. 2007). Odontocetes have also been observed to change vocalization patterns in response to air guns (Goold and Fish 1998). Noise impacts from military sonar can be similar to those from seismic surveys, with displacement of animals in the range of sonar activities. Active sonar produces loud (>210 dB re 1μPa) sounds across a range of frequencies, including those over 10 kHz. Changes in vocalization may also result. For example, humpback whales lengthened or delayed songs during playbacks of sonar (Miller et al. 2000, Fristrup et al. 2003). Military sonar has also been suggested as a cause of marine mammal strandings (Cox et al. 2006). While the link between sonar and strandings is not confirmed, a probable hypotheses is that decompression-like symptoms (e.g., “the 298  Ecosystems

bends”) can occur when whales surface quickly in response to the loud sounds (Jepson et al. 2003). Additional research into noise effects, particularly to understand stress thresholds, behavioral responses, and population impacts, are warranted (Southall et al. 2012).

SHIP STRIKE S

The shipping industry also affects offshore ecosystems through collisions with large marine species. Most studies on ship strikes have focused on cetaceans because they are struck frequently relative to their small numbers (Clapham et al. 1999, Kraus et al. 2005). Ship strikes have been of concern for cetacean populations only since the early 1950s, when ship engines gained the power to travel at speeds sufficient to injure whales (Laist et al. 2001). Ships of any size can strike whales; however, those over 80 meters and traveling faster than 14 knots are most likely to have lethal effects (Laist et al. 2001). The commercial shipping industry is the primary concern due to vessel size as well as the volume of vessel traffic in California waters. Along the California coast, the large whales most susceptible to ship strikes include blue, humpback, fin, and gray whales (Redfern et al. 2013, McKenna et al. 2015). Between 1988 and 2007, twenty-one blue whale strandings attributed to ship strikes were recorded (Berman-Kowalewski et al. 2010). Recent data indicate that two humpback whale, four gray whale, one sperm whale, and four blue whale deaths were caused by ship strikes between 2004 and 2008, with additional historical records of a strike on a minke whale (Carretta et al. 2010). These numbers undoubtedly underestimate mortality caused by ship strikes, as many animals do not strand on land following injury or death, and many strikes are not reported or even detected by large container ships (Kraus et al. 2005). Other offshore marine species undoubtedly are victims of ship strikes, but these species have been less studied or strikes are less easily detected. Boat strikes to leatherback sea turtles are known to occur but are not considered a major source of mortality (Turtle Expert Working Group 2007). Seabirds are attracted to the lights of ships at night and may become disoriented, striking hard surfaces on the boats and suffering injury or mortality; this has not be studied in the CCS (Black 2005).

POLLUTION

Assessments of ship discharge effects have focused on oil and have been studied primarily for individual taxa, such as seabirds and marine mammals, rather than for the entire offshore ecosystem. Seabirds exposed to even small amounts of oil can lose flight capability and experience wasting fat and muscle tissues and abnormal conditions in the lungs, kidneys, and other organs (Clark 1984, Henkel et al. 2012). Strandings of seabirds occur regularly in areas of heavy vessel traffic because of exposure to oils associated with many commercial vessels. The number of stranded birds is likely a gross underestimate, as approximately 95% of bird carcasses sink within several days (Wiese and Robertson 2004). In the Monterey Bay region, seabird strandings are regularly attributed to oiling in the absence of large-scale spills and include offshore species such as sooty shearwaters (Newton et al. 2009). Chronic exposure to oil might have larger impacts on seabird populations than occasional, large oil spills often

considered more catastrophic (Wiese and Robertson 2004). Overall impacts of chronic, low-level oil exposure on marine mammals and sea turtles are less than on seabirds. Moderate exposure to oil can affect these organisms’ skin, blood, digestive systems and salt glands, behavior, and thermoregulatory abilities. However, acute impacts from low-level exposure are considered minimal, and recovery occurs within several weeks (Engelhardt 1983, Lutcavage et al. 1995, Milton et al. 2003). Oil spills can affect many offshore ecosystem components, such as deep-water corals, seabirds, sea turtles, pinnipeds, and cetaceans. Oil spill impacts on fish stocks extend from ecological to economic as people reduce fish consumption out of fear of contamination (Engelhardt 1983, Clark 1984, McCrea-Strub et al. 2011, Henkel et al. 2012, White et al. 2012). Our knowledge of oil pollution effects on offshore marine ecosystems is increasing rapidly following the Deepwater Horizon spill in the Gulf of Mexico, from an oil well approximately 80 kilometers from shore in waters 1,500 meters deep (Lubchenco et al. 2012). Large oil spills have occurred in the CCS, including the famous 1969 Santa Barbara oil spill that resulted from oil seeping from a well in the Santa Barbara channel. Approximately 13,000 to 16,000 m3 of crude oil spilled into the channel and surrounding area, causing large-scale environmental damage. The Santa Barbara oil spill, despite its devastation, impelled a number of marine protective measures including the creation of the National Marine Sanctuaries.

OCE AN ACIDIFICATION

While carbon uptake helps to regulate Earth’s climate, increased oceanic carbon uptake due to increases in fossil fuel emissions are causing changes in ocean chemistry. Ocean acidification—​a decrease in the pH of the ocean—​influences the ability of animals with carbonate skeletons to uptake carbonate for skeletal structures. It could pose greater risks in coastal upwelling areas such as the CCE because upwelling brings more acidic waters to the surface waters (Harris et al. 2013, Hauri et al. 2012, 2013). Ocean acidification has affected a number of species common in the CCE including deep-sea corals, krill, fish, mollusks, and gelatinous zooplankton (Attrill et al. 2007, Fabry et al. 2008, Guinotte and Fabry 2008). Changes in ocean chemistry may also change the way that sound travels, particularly by decreasing sound absorption (Hester et al. 2008, Sehgal et al. 2010). This could affect species such as whales that use long-distance calls to communicate (McDonald et al. 2009). Changes to the CCE resulting from ocean acidification will be compounded by other climate change impacts (e.g., ocean warming, ocean deoxygenation) and are expected to include range shifts in many CCE species, reduced fishery landings, and potentially reduced prey availability for apex marine predators (Ainsworth et al. 2011).

OTHER ISSUE S AND RISKS

Other risks associated with commercial shipping include transportation of invasive species and pollution resulting from direct trash and sewage discards. Through ballast water, marine species ranging from crabs to fish to microorganisms may be transported across entire ocean basins. Measures

are in place to reduce invasions. For example, the International Maritime Organization recommends ballast water be exchanged at-sea (known as “reballasting”) to reduce potential invasions of species from ports of origin; however, the impacts on offshore ecosystems are unknown (Tsolaki and Diamadopoulos 2010). The shipping industry, particularly cruise ships, also generates large amounts of waste. It is estimated that while cruise ships account for only a small fraction of the commercial vessels, they produce over a quarter of all merchant vessel waste (Butt 2007). International laws allow the discharge of raw, untreated sewage on the high seas beyond 12 miles from land; however, the U.S. Clean Cruise Ships Act of 2005 prohibits the discharge of gray water, sewage, or bilge waters within U.S. waters. Under the International Convention for the Prevention of Pollution from Ships (MARPOL 73/78), there is a complete ban on dumping of plastics at sea; these must be incinerated or disposed of at ports. Despite these restrictions, a large amount of waste is disposed of at-sea, particularly offshore, and while little is known of its environmental impact at the ecosystem level, the Pacific Garbage Patch illustrates the long-term effects of regular dumping at sea (Howell et al. 2012). This highlights the importance of reducing waste and increasing the ability to better handle and treat waste onboard vessels (Johnson 2002, Butt 2007).

Conservation Issues and Management Structure Management of the offshore ecosystem is complex because it is highly dynamic and, in many regards, a poorly understood system. In addition, there are multiple impacts on this system, occurring simultaneously and resulting in cumulative impacts with complex interactions. Effectively managing the offshore ecosystem requires being able to effectively characterize the threats, impacts, and management needs of the system, as well as developing appropriate management tools. Here we describe some of the potential cumulative impacts on the offshore ecosystem, current management tools, and an example of where science and management have worked together in an effort to improve protection of whales in the offshore ecosystem.

CUMUL ATIVE IMPACTS ON THE OFFSHORE ECOSYSTEM

Despite our still limited knowledge, the CCS is one of the best-studied ecosystems in the world, and a number of studies have characterized cumulative impacts on this system. Building on a global assessment of human impacts on marine ecosystems (Halpern et al. 2008), Halpern and colleagues (Halpern et al. 2009) conducted an assessment of human impacts on the CCS. Looking across twenty-five human activities, they ranked the threats and their impacts on nineteen different marine ecosystems within the CCS including surface (>30 meters) and deep (>200 meters) pelagic ecosystems, and the hard and soft substrates found in deep waters including canyons and seamounts. Climate change impacts, particularly ocean acidification, and activities that affect the benthic surfaces, such as oil rigs and destructive demersal fishing practices, were identified as being the greatest threat to offshore ecosystems. The spatial pattern of impact showed an increase in cumulative impacts moving further north, largely due to the affect of climate change. The Offshore Ecosystem   299

12

Cumulative utilization and impact

Another assessment (Maxwell et al. 2013) considered the impacts specifically on pelagic predators. Using similar threats and weighting methods as Halpern et al. (2009), they found that the continental shelf was the region most impacted by human activities across the eight pelagic predator species studied. The impact of climate change stressors were again rated highest, along with various forms of pollution (Figure 16.7).

10 45˚ N

8 6

MANAGEMENT OF THE CALIFORNIA OFFSHORE ECOSYSTEM Marine protected areas (MPAs) are management tools that

have been utilized extensively in California waters, particularly nearshore via the California Marine Life Protection Act (see Chapter 37, “Range Ecosystems”) and through the U.S. National Marine Sanctuaries. Five National Marine Sanctuaries have been designated along the U.S. West Coast. These five sanctuaries are the Olympic Coast, Cordell Bank, Gulf of the Farallones, Monterey Bay, and Channel Islands National Marine Sanctuaries, totally over 31,263 km2 in the CCS (Table 16.1, see Figure 16.1). The newest addition to the sanctuaries was an extension of the Monterey Bay Sanctuary to include Davidson Seamount in 2009. Davidson Seamount is 129 kilometers southwest of Monterey, California, 42 kilometers long, 13 kilometers wide, and 2,280 meters tall but still 1,250 meters from the water’s surface. It hosts a dense and diverse array of deep-water corals and sponges, particularly along the flanks of the seamount (Clague et al. 2010). Due to its depth, the benthic ecosystem is still pristine, and restrictions on bottom fishing have been put in place to maintain the benthic habitat. Expansion of the Cordell Bank and Gulf of Farallones Sanctuaries north to Point Arena is also currently under consideration (NOAA 2013). The Sanctuary Program was created primarily in response to concerns over pollution through dumping of waste and oil spills in coastal regions. The sanctuaries began with grand intentions, and some of its early backers originally envisioned the sanctuaries to be akin to the National Parks System on land, with protections at the level of those afforded under the Wilderness Act (Owen 2003, Chandler and Gillehan 2004). Over time, however, Congress has defined the primary purpose of National Marine Sanctuaries as multiple-use regions, often with little restrictions on uses within their boundaries, and the act as a whole is considered to be weakly structured with little ability to afford significant protection to marine ecosystems (Chandler and Gillehan 2004). While many activities, such as fishing, are allowed throughout many of the sanctuaries, dumping of waste is prohibited in all sanctuaries, with some minor exceptions, and oil and gas development is prohibited in all of the sanctuaries in the CCS (Owen 2003). Furthermore, the sanctuaries have played an important role in scientific research and education in the regions where they have been designated. The highly dynamic and integrated system of the CCS has resulted in calls for management that is as dynamic as the marine system (Dunn et al. 2011, Hazen et al. 2013, Maxwell et al. 2013). Dynamic management can be put into place through large-scale protection for species or ecosystems (e.g., mobile marine protected areas [Game et al. 2009]), or through managing multiple or individual sectors using management measures that move and shift with the changing ecosystem (e.g., TurtleWatch, a tool that highlights probable bycatch regions 30 0  Ecosystems

4 Cape Mendocino

40˚ N

2

Point Arena

National Marine Sanctuaries

35˚ N

Point Conception

128˚ W

126˚ W

124˚ W

122˚ W

120˚W

118˚W

116˚W

FIGURE 16.7 Cumulative utilization and impacts (CUI) that combines tracking data and impact data for eight marine predators in the U.S. exclusive economic zone. Outer solid line represents the U.S. exclusive economic zone; solid inner lines represent the U.S. National Marine Sanctuaries; hashed lines represent the 200 meter depth contour. Source: Maxwell et al. 2013.

of endangered loggerhead sea turtles within pelagic longline fishing grounds off Hawaii [Howell et al. 2008]). While there is still a need to develop the legal and ecological frameworks to ensure feasibility of dynamic ocean management, incorporating the dynamic nature of marine ecosystems into management strategies will undoubtedly be the future of marine management (Lewison et al. 2015, Maxwell et al. 2015).

Integrating Science and Management CASE STUDY: MODIF YING SHIPPING L ANE S TO AVOID WHALE STRIKE S

The region around San Francisco Bay is known as a hotspot for marine mammals, particularly whales (Keiper et al. 2005). San Francisco Bay is also home to a number of the largest ports on the U.S. West Coast, with approximately twenty tankers, container ships, or barges entering and exiting the port everyday via the predesignated shipping lanes. As a result of this traffic, from 1988 to 2011 there have been thirty documented whale strikes within the Gulf of the Farallones and Cordell Bank National Marine Sanctuaries, with the real number likely much higher (Joint Working Group on Vessel Strikes and Acoustic Impacts 2012). The species affected by ship strikes include blue, humpback, fin, and gray whales, and in some cases, the mortality caused by ship strikes greatly

TA B L E 16 .1 U.S. West Coast National Marine Sanctuaries (NMS)

Sanctuary name (designation year) Channel Islands NMS (1980)

Area

Area fully protected (no-take) designated as part of the National Marine Sanctuaries

4,294 km2

2,344 km2

Key activities restricted or prohibited Alteration or construction on the seafloor Oil and gas exploration or production Removal/damage of cultural or historical resources Depositing or discharging material

Cordell Bank NMS (1989)

1,362 km

2

None

Oil and gas exploration or production Exploration or production of other minerals Depositing or discharging material Depositing or discharging material outside the NMS that may injure NMS resources

Gulf of the Farallones NMS (1981)

3,250 km2

None

Alteration or construction on the seafloor Oil and gas exploration or production Removal/damage of cultural or historical resources Depositing or discharging material Use of motorized watercraft

Monterey Bay NMS (1992)

13,784 km

2

None

Alteration or construction on the seafloor Oil and gas exploration or production Removal of natural resources Removal/damage of cultural or historical resources Depositing or discharging material Depositing or discharging material outside the NMS that may injure NMS resources

Olympic Coast NMS (1994)

8,573 km2

None

Alteration or construction on the seafloor Oil and gas exploration or production Exploration or production of other minerals Removal of natural resources Removal/damage of cultural or historical resources Depositing or discharging material Depositing or discharging material outside the NMS that may injure NMS resources

Source: Adapted from Marine Conservation Institute National Marine Sanctuaries Fact Sheets, 2006, http://www.marine-conservation.org/what-we -do/program-areas/mpas/national-marine-sanctuaries.

exceeds the sustainable human-caused mortality (also known as “potential biological removal,” or PBR) populations can withstand without negatively impacting populations. For example, the PBR for blue whales in the eastern North Pacific is 3.1 animals; documented ship strikes can be as high as 4 annually, and this number is likely an underestimate (Joint Working Group on Vessel Strikes and Acoustic Impacts 2012; Redfern et al. 2013). In an effort to reduce ship strike mortality in the San Francisco Bay Area, a joint working group (JWG) was convened by the Gulf of the Farallones and Cordell Bank National Marine

Sanctuaries. This JWG included representatives from NOAA, the agency responsible for managing and protecting whales under the Marine Mammal Protection Act and Endangered Species Act, and the U.S. Coast Guard, which is responsible for implementing the Ports and Waterways Safety Act, which promotes navigation, safety, and protection of the marine environment. The fine-scale patterns of whale habitat use in the region were identified across seasons, and this was compared with shipping vessel traffic patterns to identify the areas of greatest risk to whales (Keiper et al. 2012). Using this scientific assessment, and combining it with The Offshore Ecosystem   301

available management measures, the JWG came up with three key recommendations to reduce whale strikes (Joint Working Group on Vessel Strikes and Acoustic Impacts 2012). Carbon emissions and economic impacts on the shipping industry were considered at all stages, with the goal of maximizing whale protection while minimizing economic and emission impacts. First, the JWG recommended modifying the shipping lanes, extending them to three nautical miles beyond the continental shelf edge to minimize shipping activity within the most sensitive areas. The shelf-break region is a key area for whales because of the high productivity associated with upwelling along the shelf edge. Second, the JWG recommended the implementation of dynamic management areas (DMAs). Vessels would have to choose alternate shipping lanes or reduce speeds when concentrations of whales are present, with DMAs designated using data from real-time sightings of whales. DMAs have been applied to reduce ship strikes on the U.S. East Coast with variable success (Silber et al. 2012) and are considered to be a better option economically for the shipping industry than semipermanent restrictions that would go into effect seasonally. Finally, the JWG recommended a real-time whale sighting/monitoring network with participation from the shipping industry to support the DMAs. The Joint Working Group recommendations have been approved by both the National Marine Sanctuaries and the International Maritime Organization (IMO), and extension of the shipping lanes was implemented in June 2013. This demonstrates how science and management can interact to inform how anthropogenic impacts on marine ecosystems can be reduced while maintaining sustainable human uses. A similar process is under way to reduce ship strikes in the Santa Barbara Channel shipping lanes, which are the gateway to the ports of Los Angeles and Long Beach (Abramsom et al. 2009, DeAngelis et al. 2010, Betz et al. 2011). The inbound lane was shifted northward in June 2013 to move traffic away from known whale concentrations.

Summary The California Current System is a dynamic eastern boundary current system, driven by seasonal coastal upwelling that transports nutrients into the photic zone and supports a diverse and productive ecosystem. As a result, the offshore ecosystem is driven by phytoplankton blooms that support a diverse food web from zooplankton and forage fish to top predators. Copepods, euphausiids, and forage fish (e.g., sardine, anchovy) serve as critical prey resources for a suite of predators from seabirds to large whales. Intermittent gelatinous zooplankton blooms can alter the energy flow through the food web, although the forcing and trophic impact of these blooms are not well understood. Many mobile species migrate seasonally within the CCS following patterns of productivity, while the predictable prey resources bring other species from across the Pacific. Mesoscale features such as eddies, fronts, and upwelling shadows can be considered marine hotspots in the CCS, which are associated with increased trophic exchange (predation hotspots) and are often of increased ecosystem importance (biodiversity hotspots). Historically, forage fish provided livelihood for coastal communities throughout the region, although collapses have resulted in increased economic importance from bottom fish, salmon, highly migratory fish, and squid fisheries, to name a 302  Ecosystems

few. Population centers along the California coast use the offshore ecosystem for fishing, oil exploration, military activities, ecotourism, and shipping, although increased use also results in increased risk to resident species, including pollution, oil spills, ship strikes, gear entanglements, bycatch, and ocean noise. Given the strong bottom-up forcing of the CCS, the impacts of climate-driven changes in the system could be profound. These impacts include changes in the timing and strength of upwelling, which may lead to mismatches between prey availability and predator distribution, and increased hypoxia and ocean acidification, which may reduce viable habitat for many species and alter community structure. Understanding the interplay between natural climate variability and ecosystem services is critical for effective management of the CCS into the future.

Acknowledgments The authors thank the external reviewers for useful comments on earlier drafts of the chapter. We thank John Field for permission to use Figure 16.2. The authors also thank Hal Mooney and Erika Zavaleta for providing editorial support and leadership in developing this book.

Recommended Reading Allen, L. G., and M. H. Horn, editors. 2006. The ecology of marine fishes: California and adjacent waters. University of California Press, Berkeley, California. Block, B. A., I. D. Jonsen, S. J. Jorgensen, A. J. Winship, S. A. Shaffer, S. J. Bograd, E. L. Hazen, D. G. Foley, G. A. Breed, A. L. Harrison, J. E. Ganong, A. Swithenbank, M. Castleton, H. Dewar, B. R. Mate, G. L. Shillinger, K. M. Schaefer, S. R. Benson, M. J. Weise, R. W. Henry, and D. P. Costa. 2011. Tracking apex marine predator movements in a dynamic ocean. Nature 475:86–​90. Bograd, S. J., W. J. Sydeman et al. 2010. The California Current, 2003–​0 8. Pages 106–​141 in S. M. McKinnell and M. Dagg, editors. Marine Ecosystems of the North Pacific Ocean, 2003–​2 008. PICES Special Publication 4. North Pacific Marine Science Organization, Sidney, BC, Canada. Checkley, D. M., Jr., and J. A. Barth. 2009. Patterns and processes in the California Current System. Progress in Oceanography 83:49–​6 4. Mann, K. H., and J.R.N. Lazier. 2006. Dynamics of marine ecosystems: Biological-physical interactions in the oceans. Third edition. Blackwell Publishing, Oxford, UK. McClatchie, S. 2014. Regional fisheries oceanography of the California Current System: The CalCOFI program. Springer Press, Dordrecht, Netherlands. Norse, E., and L. B. Crowder. 2005. Marine conservation biology. Island Press, Washington, D.C.

Glossary Deep scattering layer  A horizontal zone of organisms at mid-depths that tend to rise toward the surface at dusk and descend again at dawn. Dynamic ocean management  Measures that manage ocean resources by taking into account the dynamic nature of marine organisms and processes. Euphotic  At depths exposed to sufficient sunlight for photosynthesis to occur. Marine hotspots  In marine systems, these are defined by (1) important life history areas for a particular species, (2) areas of high biodiversity or abundance of individuals, and (3) areas

of important productivity, trophic transfer, and biophysical coupling. Marine protected areas (MPAs)  Special areas established and managed for conserving marine resources, often allowing specific recreational and commercial uses. Mesopelagics  Species that primarily occur in the mesopelagic zone, which extends from 200 to 1,000 meter depth. Oligotrophic  Describes an environment with very low nutrient availability. Pelagic  Refers to the open ocean rather than nearshore waters, and vertically above seafloor habitats that are considered benthic. Plankton  Organisms that live in the water column and are incapable of swimming against a current. These include phytoplankton (plants) and zooplankton (animals). Trophic interactions  Interactions between the producers and consumers in an ecosystem (i.e., a food web).

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and 2010, estimated from acoustic-trawl surveys. Fishery Bulletin 110:110–​122. Zwolinski, J. P., R. L. Emmett, and D. A. Demer. 2011. Predicting habitat to optimize sampling of Pacific sardine (Sardinops sagax). ICES Journal of Marine Science 68:867–​879. Žydelis, R., R. L. Lewison, S. A. Shaffer, J. E. Moore, A. M. Boustany, J. J. Roberts, M. Sims, D. C. Dunn, B. D. Best, Y. Tremblay, M. A. Kappes, P. N. Halpin, D. P. Costa, and L. B. Crowder. 2011. Dynamic habitat models: Using telemetry data to project fisheries bycatch. Proceedings of the Royal Society B 278: 3191–​3200.

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SE VENTEEN

Shallow Rocky Reefs and Kelp Forests MAR K H. CARR and DAN IEL C. REED

Introduction Kelp forests are among the iconic ecosystems of California

despite most people having experienced them only remotely. From shore, viewers are captivated by the forest canopy formed at the ocean surface and the associated marine birds and mammals. Some observers have experienced kelp forests in public aquaria, but most are familiar with these ecosystems only through photographs and video. Nonetheless, these glimpses impress one with the extraordinary three-dimensional structure and biodiversity characteristic of kelp forest ecosystems. These impressions are captured in Charles Darwin’s exuberant description of the forests of giant kelp (Macrocystis pyrifera) around the coast of Tierra del Fuego, even though his observations were largely limited to the surface and samples brought to him aboard the RV Beagle: The number of living creatures of all Orders, whose existence intimately depends on the kelp, is wonderful. A great volume might be written, describing the inhabitants of one of these beds of sea-weed. . . . Innumerable crustacea frequent every part of the plant. On shaking the great entangled roots, a pile of small fish, shells, cuttle-fish, crabs of all orders, sea-eggs, star-fish, beautiful Holuthuriae, Planariae, and crawling nereidous animals of a multitude of

forms, all fall out together. Often as I recurred to a branch of the kelp, I never failed to discover animals of new and curious structures. . . . I can only compare these great aquatic forests of the southern hemisphere with the terrestrial ones in the intertropical regions. Yet if in any country a forest was destroyed, I do not believe nearly so many species of animals would perish as would here, from the destruction of the kelp. Amidst the leaves of this plant numerous species of fish live, which nowhere else could find food or shelter; with their destruction the many cormorants and other fishing birds, the otters, seals, and porpoises, would soon perish also. (C. Darwin, 1839)

Globally, a variety of species of kelp establish forests along margins of continents and islands in the temperate oceans of both the southern and northern hemispheres. Kelp forests also develop in some subtropical areas that experience considerable coastal upwelling. The distribution of these kelp forests is generally limited to areas where rocky reefs occur at shallow depths (generally 50% dur-

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ing El Niño storms (Revell and Griggs 2006, 2007). The formation of erosion hotspots during El Niño years appears related to a pattern of beach rotations resulting from a southwesterly shift in wave direction (Revell and Griggs 2006, 2007). High wave energy associated with El Niño events erodes shorelines and can leave beaches sand-starved and narrow for more than a year (Barnard et al. 2011, 2012). These events also affect the distribution of beach zones, nearshore kelp beds and wrack inputs, the survival of intertidal invertebrates, and habitat and prey resources for birds requiring recovery periods stretching from months to years (Revell et al. 2011) (see Figure 20.3). The ecological consequences of storm-related erosion on sandy coastal habitats operate on different time scales according to proximity to sand supply, persistence of beach rotations and hotspots, metapopulation and community dynamics, and plant and animal life histories (Arntz et al. 1987, Revell et al. 2011). Paradoxically, large, episodic storm events can also have positive effects on the sediment budget of the shoreline (Kniskern et al. 2011). Intense precipitation and runoff events can transport massive amounts of sediment to the coast, as exemplified in January 2005 when the Santa Clara River (Ventura County) delivered 5,000,000 cubic meters of sediment to the coast (ten times the annual average) and built the shoreline more than 170 meters seaward at the river delta (Barnard and Warrick 2010) (see Figure 20.1). The immediate effects of erosion are obvious and receive a great deal of media attention, while new sediment supplied to the coast is not generally widely noticed. In fact, the sand delivered from large storms may be deposited primarily in the nearshore zone and appreciated mostly by surfers. However, beaches are often observed to be wider following major storm, rainfall, and runoff years (Griggs et al. 2005). Finally, sea level rise is likely to exacerbate coastal erosion by raising mean water levels and consequently increasing exposure of backshores, including the toes of dunes and bluffs to wave attack. If sea levels rise high enough dunes may be breached, allowing seawater to flood behind the dune. Dune breaching can reduce the extent of coastal strand and dune habitat, increase salinity levels, and increase exposure to inundation.

Ecosystem Attributes and Food Webs To a casual visitor, the wave-swept sands of California’s beaches might appear relatively empty of life. No plants can take hold on the shifting sand of open beaches, and most characteristic animals are highly mobile and nocturnal, burrowing deeply into the sand during the day. However, sandy beaches in California support a unique and rich animal diversity that includes resident animals; fish and invertebrates that depend on beaches for a key part of their life cycle; and birds and pinnipeds that winter, forage, breed, or nest on beaches and dunes. California’s sandy beaches support some of the most diverse intertidal invertebrate communities ever reported for beach ecosystems (McLachlan 1994, McLachlan and Dorvlo 2005), with more than 45 species found in single surveys on a variety of beaches and more than 105 species recorded in southern and central regions (Straughan 1983; Dugan, Hubbard, Engle et al. 2000; Dugan et al. 2003; Schooler, Dugan and Hubbard 2014; Schooler et al. in prep.). Crustaceans, polychaete worms, and mollusks are major intertidal invertebrate

FIGURE 20.4 Connectivity between beaches and kelp forests is an important driver of beach ecosystems in California. Wrack, in the form of beach-cast macroalgae and seagrass, exported from kelp forests and reefs is a prominent feature and key ecological resource on California beaches. Imported wrack supports a major component of the beach food web and a high diversity of endemic intertidal animals. Beach-cast giant kelp (Macrocystis pyrifera), feather boa kelp (Egregia menziesii), and surfgrass (Phyllospadix scouleri) on a beach near Isla Vista, California. Photo: David Hubbard.

groups on California beaches and elsewhere. Endemic insects, including a number of flightless beetles, form an important element of the diversity of California’s beaches. Many additional species are likely present on California beaches, but identification of several important taxa including infaunal polychaete worms and wrack-associated insects is limited by current taxonomic knowledge.

Zonation The distinctive mobility of beach intertidal animals and of the sand itself limit the applicability of many classic tenets of intertidal zonation useful for exposed rocky shore biota (Peterson 1991). On rocky intertidal shores, many characteristic plants and animals survive wave and tidal action by strongly resisting movement (see Chapter 18, “Intertidal”). On sheltered muddy shores, many animals build and inhabit relatively permanent burrows in consolidated sediments or attach to plants (see Chapter 19, “Estuaries: Life on the Edge”). On sandy beaches, intertidal animals must move; although they can occupy burrows for minutes to a few days, they do not inhabit permanent burrows or locations on the beach profile. Most regularly move up and down the beach profile in response to tide height and phase (e.g., Dugan et al. 2013). Over the course of larger seasonal or event-driven erosion and accretion, these animals also move much greater distances across the shore. The remarkably high mobility of beach animals thus underpins many of the major differences in intertidal ecology observed among sandy beaches and other more stable shore types. Three relatively distinct zones of intertidal beach animals can be often be identified for a given day or tide condition (McLachlan and Jaramillo 1995). These zones generally correspond to (1) the relatively dry sand around and above the high tide strand line, (2) the damp sand of the middle intertidal, and (3) the saturated sand of the lower and swash intertidal zone (see Figure 20.2). The locations of these zones and of their characteristic inhabitants move up and down (as well as along) the beach in response to tides and water motion, shifting dramatically in just hours. As the tide floods after a low tide, ani-

mals burrowed in the swash and low zone of a beach, such as sand crabs, emerge from the sand to migrate up the shore. On the ebb (receding) tide, they move down the shore again and reburrow. The positions of many beach animals on the profile also respond distinctly to the lunar tide cycle, burrowing higher on the beach during spring tides and occupying lower strata during neap tides (Dugan et al. 2013). Although overall abundance remains the same, these semilunar movements create changes in the density of animals burrowed in a particular zone. These shifts in density can affect the intensity of biotic interactions and must be accounted for in the design of quantitative surveys. The positions of these mobile burrowing animals also react strongly to wave energy, beach erosion, and accretion. Annual shifts in position defined as the ecological envelopes of these animals generally extend across greater than 60% of the overall beach width (Dugan et al. 2013).

Beach Food Webs Because beaches lack attached plants, they have very low in situ primary production. Primary production is limited mostly to diatoms in the surf and lower intertidal zones. Thus beach food webs depend mainly on imported organic matter from other marine ecosystems. In California the main sources are phytoplankton from nearshore pelagic ecosystems and macrophytes (algae and seagrass) from nearby kelp forests, rocky shores, and estuaries (Figure 20.4). These two major marine inputs support distinct components (or subwebs) of the intertidal invertebrate community. Drift algae and seagrasses stranded on beaches as macrophyte wrack (see Figure 20.3) represent an important link between reef and kelp forest ecosystems and beaches in many regions, with estimated annual inputs exceeding 500 kg m-1 (ZoBell 1971, Dugan et al. 2011). Intertidal consumers consist of (1) suspension-feeding clams, hippid crabs, mysids, and amphipods that filter plankton from the wave wash, (2) wrack-feeding amphipods, isopods, and insects that feed on drift macrophytes deposited by waves on the beach, and (3) burrowing deposit feeders that feed on both wrack particles and phytoplankton pumped into the sand by waves and tides. The dependence of beach food

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FIGURE 20.5 On the lower beach and in the swash zone many of the invertebrates are suspension feeders that sieve or filter plankton from the surf and swash water. Photos: (A–​C, F) David Hubbard; (D) Dan Ayres, Washington Department of Fish and Wildlife; (E) Shane Anderson. A Sand crabs (Emerita analoga) on surface including females carrying

clutches of orange eggs. B Backwash of a wave showing texture of sand crab feeding

aggregation. C Bean clams (Donax gouldii) exposed at low tide at La Jolla with

inset of an individual clam. D Pacific razor clams (Siliqua patula), found only in the northern

part of the state. E A Pismo clam (Tivela stultorum) underwater with siphons extended

(the large frilly intake siphon helps to filter sandy water, the smaller open siphon is for outflow). F Legally harvestable Pismo clams from an intertidal beach.

webs on allochthonous marine resources results in strong bottom-up effects that propagate up to avian and other predators (Dugan et al. 2003).

Suspension Feeders Sand crabs, clams, and other suspension feeders make up most of the biomass (80%–​98%) of invertebrate communities on California beaches (Figure 20.5). They can be extremely abundant in the swash zone, exceeding 100,000 individuals m-1 of shoreline and providing important prey for shorebirds, seabirds, fish, pinnipeds, and even sea otters. The growth and population biology of these animals are closely coupled to ocean processes such as upwelling that stimulates growth of nearshore phytoplankton and currents that carry their pelagic larvae away from and back to beaches (see Chapter 16, “The Offshore Ecosystem”). The most widespread and often the most abundant sus-

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pension-feeding beach invertebrate is the sand crab (Emerita analoga), which can occur on almost every type and length of exposed sandy beach in California (Efford 1970, 1976). This hippid crab is a rapidly burrowing sediment generalist (sensu Alexander et al. 1993) with excellent orientation and swimming abilities that can colonize the full spectrum of exposed sandy beaches, from fully reflective to dissipative (Dugan, Hubbard, and Lastra 2000). These highly mobile crabs generally aggregate in the active swash zone and follow this zone up and down the beach with the tides (see Figure 20.5). The sand crab uses its plumose (feathery) second antennae to sieve fine particles, primarily phytoplankton, from the turbulent moving water in the swash zone (Efford 1966). Its growth is correlated with surf zone chlorophyll- a concentrations (an indicator of phytoplankton abundance) along the California coast (Dugan et al. 1994) indicating coupling of its population biology with upwelling and productivity gradients. Sand crab larvae spend three to four months as zoea in the plankton before settling on beaches as megalopa/postlarvae (Efford 1970, 1976). Populations in the northeastern Pacific appear well mixed with high gene flow (Dawson et al. 2011). Sand crabs have sensitive and plastic life history responses to environmental variation (Fusaro 1978, Dugan et al. 1991, Dugan et al. 1994, Dugan and Hubbard 1996). Size and age at maturity, growth rate, maximum male and female crab size, and survival all increase significantly from south to north across its geographic range in California (Dugan et al. 1991, Dugan et al. 1994, Wenner et al. 1993). This strong geographic pattern appears to be related to the gradient of increasingly productive yet colder waters from southern to northern California (Dugan 1990, Dugan et al. 1994). Sand crab populations have been successfully used as bioindicators (Siegel and Wenner 1984, Wenner 1988) as they readily bioaccumulate metals, hydrocarbons, pesticides (e.g., Burnett 1971, Rossi et al. 1978, Wenner 1988) and harmful algal toxins such as Paralytic Shellfish Poisoning and domoic acid (Bretz et al. 2002, Ferdin et al. 2002). Tissue loadings of DDT in sand crabs were a key factor used to describe the coastal distribution of this now-banned pesticide associated with the White’s Point Outfall on Palos Verdes in the Southern California Bight (Burnett 1971). This major Los Angeles sewage outfall discharged tons of DDT from the nation’s largest manufacturer of this pesticide, Montrose Chemical Corporation, into the ocean on the Palos Verdes shelf from the 1950s through 1971. Intertidal clams of California beaches have more limited distributions than sand crabs, preferring beaches with flatter slopes and finer sand. Species include the bean clam (Donax gouldii), the Pismo clam (Tivela stultorum), and the razor clam (Siliqua patula, as well as S. lucida) (see Figure 20.5). Colorful bean clams are a southern species most abundant and most commonly encountered on intermediate beaches south of Point Conception, where they can cover the beach like gravel at times (Coe 1953, 1955). Thick-shelled, slow-burrowing Pismo clams inhabit intermediate to dissipative beach types from Half Moon Bay south; their distribution can extend from the intertidal into subtidal sands (Fitch 1950, McLachlan et al. 1995, 1996). Thin-shelled, rapidly burrowing razor clams are restricted to dissipative or nearly dissipative beaches in the northernmost part of the state (McLachlan et al. 1995, 1996). Both razor clams and Pismo clams are fished recreationally in California. Fishing of all types has impacted beach clams around the world, but long-lived, large intertidal species appear to be particularly vulnerable (McLachlan et al.

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FIGURE 20.6 Higher on the beach, much of the food web is fueled by beach cast wrack, particularly kelps. The major consumers of wrack resources are beach hoppers (Megalorchestia spp.) that burrow in damp sand during the day and emerge at night to locate and feed on freshly deposited kelp. Photos: David Hubbard. A Raylike stripes on the surface mark burrows of Megalorchestia

californiana (inset shows a mature male perched on a stranded pneumatocyst of giant kelp). B Burrows of other species are marked by irregular mounds, the

inset shows two male Megalorchestia corniculata competing for a burrow that contains a female.

1996). The Pismo clam, which can live for more than fifty years and reach shell lengths of greater than six inches (Fitch 1950), is a classic example of this problem. This large, slowmoving broadcast spawning clam might be considered the “abalone of the beach,” although its decline preceded that of the abalone fishery by decades. In the face of declining landings, the commercial fishery for Pismo clams was closed in the state in 1948. Despite rolling closures, transplants, and changes in regulations on size and bag limits (Fitch 1950), populations of this highly desirable clam have never recovered to commercially harvestable levels. Suspension-feeding mysids (Archeomysis, Acanthomysis spp.) can be extremely abundant in the swash zone, particularly in the northern region of the state (Nielsen et al. 2013). These swimming and burrowing crustaceans provide a major prey resource for fishes. In the midbeach zone, polychaetes in the family Spionidae that capture suspended particles with their long palps can be very abundant (Dauer 1985).

FIGURE 20.7 Kelp wrack inputs are rapidly processed by a wide variety of species found only on beaches. Photos: (A-C) David Hubbard, (D-E) Jenifer Dugan.

These include two species of oniscid isopods, (A) Alloniscus perconvexus and (B) Tylos punctatus, that burrow in damp sand above the twenty-fourhour high-tide line during daylight, emerging at night to feed on kelp and other organic material; (C) herbivorous tenebrionid beetles (Phaleria rotundata); (D) pupal cases of kelp flies (Fucellia) that feed on and develop in piles of kelp wrack with the timing of emergence related to spring lunar tides; and (E) only stipes (stems) of giant kelp (Macrocystis pyrifera), surrounded by the burrows of beach hoppers remain on the beach after all the kelp blades (leaves) on this frond were consumed overnight.

Wrack Feeders In California more than 40% of the intertidal invertebrate species on a beach are directly associated with wrack (Dugan et al. 2003). Intertidal invertebrate species richness is strongly correlated with wrack abundance on beaches that are relatively unmanipulated (Dugan et al. 2003). Strong spatial and temporal variation in wrack inputs (see Figure 20.3) affect intertidal community and food web structure and dynamics (e.g., Revell et al. 2011). Wrack-dependent invertebrates are typically most abundant and diverse where marine macroalgae, particularly kelps and kelp forests, grow on nearshore reefs (see Chapter 17, “Shallow Rocky Reefs and Kelp Forests”). Although kelps are greatly preferred as food sources for intertidal consumers, macrophyte wrack of all types (brown, red and green macroalgae, surfgrass, and eelgrass) functions as habitat for many of these invertebrates (Figures 20.6, 20.7).

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FIGURE 20.8 Detritus that is pumped into the sand by waves and tides supports a number of burrowing deposit-feeders. The red polychaete (Thoracophelia mucronata), seen here excavated in a shovelful of sand at Scripps Beach, California, from the mid-beach at low tide, feeds like an earthworm, ingesting sand as it burrows through it and digesting organic material from on and between the sand grains. Photo: Nicholas Schooler.

The damp sand near the high tide strand line is preferred by many wrack consumers, such as beach hoppers, isopods, and some beetles that build temporary burrows in this zone during the day and emerge at night to feed on freshly deposited kelp wrack. Amphipods or beach hoppers (Talitridae Megalorchestia spp.) are dominant wrack consumers, greatly preferring kelps (e.g., Macrocystis pyrifera, Egregia menziesii, Nereocystis luetkeana) over other macrophytes (Lastra et al. 2008). These widespread, nocturnal amphipods can reach abundances of greater than 90,000 individuals m-1 on kelpstrewn beaches (Lastra et al. 2008). Megalorchestia is represented by six species in California, and individual beaches can support up to four species at a time (Schooler et al. in prep). Isopods (e.g., Alloniscus perconvexus and Tylos punctatus) build distinctive burrows in the vicinity of the high wrack line and can reach high abundance on some beaches (Hubbard et al. 2014). Other important wrack consumers include a great variety of intertidal insects, such as kelp and seaweed flies (Fucellia and Coelopa spp.), whose larvae feed on and develop in moist, aging piles of kelp wrack. Herbivorous beetles (Phalaria rotundata, Epantius obscurus, Tenebrionidae) and larvae of the intertidal weevil (Emphyastes fucicola) also feed on kelp wrack, while adult weevils are considered saprophagous. As stranded wrack ages on the beach, it is colonized and fed on by a successional sequence of invertebrates starting with amphipods (Talitridae), followed by flies and specialized beetles (Yaninek 1980). Piles of wrack provide both food and essential microhabitat for beach invertebrates from the first night stranded on the beach until they disappear. Beaches with high wrack inputs support dense populations of these invertebrate consumers that provide prey to a high diversity and abundance of wintering and migratory shorebirds even during high tides (Dugan et al. 2003, Hubbard and Dugan 2003). Wrack-associated invertebrates and insects appear to be particularly important for short-billed shorebirds that search for prey visually, such as black-bellied plovers (Pluvialis squatarola) and snowy plovers (Charadrius nivosus) (Dugan et al. 2003). Snowy plovers are federally listed as a threatened 398  Ecosystems

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FIGURE 20.9 Invertebrate predators and scavengers feed on the abundant lower-beach suspension feeders and the wrack-dependent upper-beach species as well as carrion delivered by waves. Photos: (A–E, G–H) David Hubbard, (F) Jenifer Dugan.

(A) Spiny sand crabs (Blepharipoda occidentalis) on the surface and (B) in typical burrowed position with only eyes and antennae above sand, (C) porcelain sand crabs (Lepidopa californica), (D) the hermit crab (Isocheles pilosus), (E) pictured rove beetles (Thinopinus pictus), (F) black rove beetle (Hadrotes crassus), (G) the purple olive snail (Callianax biplicata), and (H) a predatory polychaete worm (Nephtys californicus).

species and are one of the few shorebirds that nest and rear chicks on sandy beaches in California. Wrack consumers lack planktonic larvae and are directdeveloping species that depend on the reproduction of resident populations. In addition, many important wrack consumers are flightless and do not swim. The disappearance of populations of two species of upper intertidal isopods (Tylos punctatus and Alloniscus perconvexus) along much of the coast of southern California during the past century illustrates the vulnerability of these types of animals to habitat alteration, loss and fragmentation from coastal development, and anthropogenic disturbance (Hubbard et al. 2014). The remaining populations are largely restricted to bluff-backed beaches where vehicle access is limited (Hubbard et al. 2014).

Deposit Feeders The polychaete worm (Thoracophelia, formerly known as Euzonus) feeds on detrital particles flushed into the sand by waves

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FIGURE 20.10 Birds feed on beach invertebrates from all levels of the

food web across the entire intertidal zone. Shorebirds are particularly abundant on California beaches between fall migration (September) and spring migration (April), but the summer is the time when a rare shorebird, the threatened western snowy plover, nests on California beaches. Photos: (A–​B) David Hubbard, (C) Callie Bowdish. On a typical winter morning, (A) a southern California beach is covered by feeding birds including sanderlings (Calidris alba), marbled godwits (Limosa fedoa), willets (Tringa semipalmata), black-bellied plovers (Pluvialis squatarola), snowy egrets (Egretta thula), and ring-billed gulls (Larus delawarensis) (Isla Vista, California); (B) long-billed dowitchers (Limnodromus scolopaceus) feed on beach hoppers burrowed near kelp wrack (Isla Vista, California); and (C) a precocial snowy plover chick (Charadrius nivosus) feeds on a beach hopper (Sands Beach, California).

and tides. It can be extremely abundant in the midbeach zone (up to 40,000 individuals m-2, McConnaughey and Fox 1949) (Figure 20.8). It is found in distinct beds marked by small, irregular holes on the surface of damp, cohesive sand (see Figure 20.7). Thoracophelia is represented by three congeners in California (T. mucronata, T. dillonensis, and T. williamsi) that can be found living together on some beaches (Law et al. 2013).

Carnivores A diversity of predatory invertebrates along with predatory fish and birds occupy the upper trophic levels of beach food webs. Commonly encountered predatory invertebrates on beaches include crustaceans such as swimming crabs (Portunus xantusii xantusii), graceful Cancer crabs (Cancer gracilis), hermit crabs (Isocheles pilosus), and shrimp (Lissocrangon stylirostris, Crangon nigricauda) that prey on mysids in the surf zones; and scavenging crabs, such as the spiny sand crab (Blepharipoda occidentalis) and the porcelain sand crab (Lepidopa californica) in the surf and swash zones (Figure 20.9). Active, carnivorous polychaete worms, such as Nephtys spp. and the strong-jawed Glycera and Hemipodia spp., are found in the mid- and low intertidal zones. Fast-swimming scavenging isopods (Excirolana spp.) can be very abundant in the

FIGURE 20.11 Beach-nesting fishes. Photos: (A) Karen Martin, (B) Doug Martin, (C) California Fish Bulletin.

In southern and central California, (A) California grunion (Leuresthes tenuis) surf in to shore to spawn near the high-tide mark during spring tides. (B) Close view of a male grunion approaching a female grunion burrowed in wet sand. In northern California, (C) a Native Californian using a dip net to catch surf smelt (Hypomesus pretiosus) in the 1930s at a beach in Humboldt County, California.

mid-intertidal, often preying on sand crab eggs as well as carrion and nipping the occasional human ankle. Around the wrack and high tide line and above, a variety of predatory beetles prowl the beach including tiger beetles (Cicindela spp., Carabidae), predaceous ground beetles (Dyschirius marinus), and many species of rove beetles (Staphylinidae, e.g., Cafius spp., Aleochara spp.) that prey on or parasitize kelp flies. The endemic intertidal rove beetles of California beaches include two large, flightless species (Thinopinus pictus, Hadrotes crassus) that prey exclusively on talitrid amphipods. Shorebirds are top predators that feed on all the types of intertidal invertebrates living on beaches and respond strongly to prey availability (Figure 20.10). Species richness and abundance of shorebirds are positively correlated with the availability of wrack and the diversity, biomass, and abundance of invertebrate prey, as well as tide, beach type, and width (Connors et al. 1981, Dugan et al. 2003, Dugan et al. 2008, Dugan and Hubbard in prep., Neumann et al. 2008, Nielsen et al. 2013). Beaches in developed regions, such as southern California, can provide birds with needed prey resources that are no longer available in coastal wetlands (Hubbard and Dugan 2003). Some seabirds dive in the surf zone or swash to catch beach invertebrates. Sand crabs and spiny sand crabs are important prey for surf scoters. Gulls regularly feed on sand crabs, either by catching them on their own or stealing them from foraging shorebirds. Gulls also feed on Pismo clams, dropping them from great heights to smash their thick shells. Nearshore fishes such as barred surfperch, redtail surfperch, yellowfin and spotfin croaker, and corbina feed on swash zone invertebrates including sand crabs and mysids (Love 1991) (Figure 20.11). These fish support important recSandy Beaches   399

reational and artisanal surf fisheries in the state (Fritzsche and Collier 2001, Valle and Oliphant 2001a and 2001b, O’Brien and Oliphant 2001), though population information for these fish is limited (see Chapter 35, “Marine Fisheries”). Sea otters and sea lions regularly prey on spiny and common sand crabs. Young sea otters may feed largely on spiny and common sand crabs as they learn to catch their own food. Pismo clams are also a favored prey of sea otters. A number of terrestrial vertebrates, including birds, mammals, and reptiles, feed on intertidal invertebrates in both the wrack line and the swash zone. Terrestrial birds that regularly feed on beaches include passerines such as barn swallows (Hirundo rustica) and cliff swallows (Petrochelidon spp.), black and Say’s phoebes (Sayornis spp.), American pipits (Anthus rubescens), kingbirds (Tyrannus spp.), savannah sparrows (Passerculus spp., including Belding’s savannah sparrow, P. rostratus/sandwichensis beldingi, an endangered songbird), and Brewer’s blackbirds (Euphagus cyanocephalus). Raptors including peregrine falcons (Falco peregrinus) hunt shorebirds on beaches, while scavengers such as turkey vultures (Cathartes aura), crows and ravens (Corvus spp.), as well as a variety of gulls and now a few reintroduced California condors (Gymnogyps californianus) regularly feed on the carcasses of marine mammals washed up on beaches. Mallard ducks (Anas platyrhynchos) have recently been observed dabbling in the swash zone for sand crabs in central and southern California (Lafferty et al. 2013). Mammals such as California ground squirrels (Otospermophilus beecheyi), raccoons (Procyon lotor), feral pigs (Sus scrofa), gray foxes (Urycyon cinereoargenteus), and the endangered dwarf island fox (Urocyon littoralis) prey on a variety of the intertidal invertebrates inhabiting California beaches.

Ecosystem Functions Along with their unique biodiversity and productive food webs, beaches provide ecological functions and services not supplied other open coast ecosystems (Schlacher et al. 2007). These include filtering large volumes of seawater, accumulating and storing sand, wave dissipation and buffering, processing of organic matter, recycling of imported nutrients, support for coastal fisheries, and pupping, nesting, and foraging habitat for endangered wildlife species. Beaches can filter large volumes of seawater (~10 m3 m-1d-1 for intermediate type beaches) pumped into the porous beach by the action of waves and tides trapping particulates in the sand (McLachlan 1989, McLachlan and Turner 1994, McLachlan and Brown 2006). Particles trapped in the sand matrix include detritus, phytoplankton, wrack particles, microorganisms, pollutants, and other particular organic matter. These organic materials become available for biogeochemical processing and transformation in the beach sand. Beaches are energy sinks that function as buffers between waves and the coastline. Stored sand on beaches, in sand bars, and in dunes plays an important role in absorbing and dissipating wave energy, even during extreme events (e.g., storms, tsunamis). Wide beaches with high volumes of sand can reduce wave energy more effectively than narrow sandstarved beaches. During storms, sand eroded from the beach is also carried off the shore to form sand bars in the shallow nearshore zone. As waves break on these nearshore sand bars, they lose wave energy and arrive at the beach with less power, mediating further erosion. 40 0  Ecosystems

Processing of imported organic matter, such as phytoplankton and marine macrophyte wrack, by intertidal consumers and microbial communities is an important ecosystem function of beach ecosystems that contributes to nearshore nutrient cycling (Dugan et al. 2011). Beaches were described as “great digestive and incubating systems” (Pearse et al. 1942). Wrack inputs and processing represent an important nutrient and energetic linkage between the marine and intertidal beach and dune environments (Dugan et al. 2011, Dugan and Hubbard 2010) with implications for conservation and management of adjacent coastal ecosystems. Nutrient cycling through the processing of wrack on beaches is mediated by intertidal invertebrates that quickly consume stranded drift macrophytes, breaking them down to particles (Lastra et al. 2008) that are then available to interstitial meiofaunal and microbial communities where turnover and remineralization are rapid (Griffiths et al. 1983, Koop et al. 1982, Koop and Lucas 1983, Dugan et al. 2011). Dense populations of amphipods are estimated to consume approximately 50% of the 42 kg m-1 month-1 of fresh kelp wrack (Macrocystis) deposited on a southern California beach during the summer (Lastra et al. 2008). Recent results suggest that processing and subsequent mineralization of wrack on beaches may be a source of nutrients to nearshore primary producers such as seagrasses and kelps (Dugan et al. 2011). Wrack also traps wind-blown sand and can promote the formation of hummocks and embryo dunes and provides nutrients that support the colonization of native dune plants (Dugan and Hubbard 2010, Nordstrom 2012). Shorebirds exemplify wildlife that use California beaches as wintering and migration habitat. The abundant intertidal invertebrates of beaches provide prey for a remarkably rich and abundant assemblage of wintering and migratory shorebirds, with more than thirty species observed at a single beach in southern California (Hubbard and Dugan 2003) (see Figure 20.10). Many species spend the majority of each year (about eight months) on California beaches, particularly in the south and central regions, leaving to migrate to breeding grounds for a few months in spring and returning by late summer or early fall. Sanderlings (Calidris alba) make up the majority of shorebirds (more than 50%) on many California beaches (e.g., Hubbard and Dugan 2003, Neumann et al. 2008, Nielsen et al. 2013) and can often be seen in large flocks running back and forth in the swash zone (see Chapter 20 cover figure and Figure 20.10). A variety of other shorebird species are commonly encountered on California beaches (Hubbard and Dugan 2003, Neumann et al. 2008, Nielsen et al. 2013). Although only a few species are resident breeders (Western snowy plover, killdeer [Charadrius vociferous], and American black oystercatcher [Haematopus bachmani]), annual means exceeding 100 birds km-1 have been observed on southern and central California beaches (Hubbard and Dugan 2003, Dugan and Hubbard in prep.). These values are among the highest ever reported globally for shorebirds on temperate beaches (Hubbard and Dugan 2003). Shorebird use of beaches is characterized by high spatial and temporal variability, including strong seasonal patterns (see Figure 20.3). Peak values for single censuses can exceed 1,000 birds km-1, but the abundance of shorebirds on California beaches can vary by over an order of magnitude among beaches and across seasons and years (see Figure 20.3) (Hubbard and Dugan 2003, Revell et al. 2011, Colwell and Sundeen 2000, Shuford et al. 1989, Neumann et al. 2008, Dugan and Hubbard in prep.). Collectively, the role of beach ecosystems in supporting win-

tering shorebirds on the California coast may be more important than is generally appreciated. The fact that populations of many species of shorebirds are declining in North America (Bart et al. 2007, Morrison et al. 2001, 2006) highlights the need to conserve coastal habitat and resources required by these wildlife, including sandy beach ecosystems. The uppermost zones of beaches located between the average reach of high tides and the toe of the vegetated foredune are particularly important for wildlife that breed on beaches. Nesting birds and mammal rookeries require beach and/or dune habitat relatively undisturbed by human activity, sparsely vegetated, and not regularly swept by waves during the nesting or pupping season. The threatened Western snowy plover and California least tern (Sterna antillarum browni) nest on open coast and sheltered beaches (Lehman 1994, Page et al. 1995) using distinctly different strategies to protect their shallow nests and chicks. Least terns nest in colonies on beaches and protect their nests and chicks by mobbing predators. In contrast, snowy plovers (and killdeer) depend on crypsis to hide their shallow nest scrapes and chicks (see Figure 20.10). Within a few hours of hatching, the precocial chicks of snowy plovers (and killdeer) permanently leave the nest to follow their parents, finding their own food (Page et al. 1995). These tiny chicks are particularly dependent on wrack-associated amphipods, isopods, and insects. The uppermost intertidal zones of open sandy beaches represent critical spawning habitat for beach-nesting fish, such as the California grunion (Leuresthes tenuis) and day and night smelt species (Figure 20.11). These fish surf up the beach to bury their eggs in the sand near the high tide line during spring high tides. After incubating in the moist, warm sand for two weeks or more, the eggs hatch as the waves hit the nests during subsequent spring high tides (e.g., Thompson 1919, Smyder et al. 2002, Martin 2015). A diversity of pinnipeds, including northern elephant seals, California sea lions, northern fur seals, and harbor seals, pup and raise their young on sandy beaches in rookeries throughout the state (LeBoeuf and Bonnell 1980). Northern elephant seals breed in sizable colonies on the mainland beaches at Año Nuevo and Piedras Blancas and on many beaches of the northern Channel Islands. Small harbor seal rookeries are also present on the mainland coast and islands. California sea lions and northern fur seals breed in large beach colonies on San Miguel Island at the mouth of the Santa Barbara Channel (northern Channel Islands). Finally, fish and invertebrates of sandy surf zones and the low intertidal zones of beaches provide prey resources for a wealth of seabirds. For many of these seabirds, beaches also serve as roosting habitat—​a function that is important and underappreciated in seabird conservation. Safe roosting areas on isolated beaches may be particularly essential for coastal diving birds, such as cormorants and pelicans, that need to dry their feathers regularly.

Human Impacts and Influences A number of widespread human activities, including watershed and coastal structures and management practices and recreation, can significantly affect habitat and community structure, biodiversity, and function of beach ecosystems (Defeo et al. 2009). In southern California the largescale human alteration of beaches includes a 50% reduction in sediment supply from damming rivers (Orme et al. 2011,

Griggs 2005a), armoring of 27% of the coastline (Griggs 1998, 2005b), beach filling totaling more than 200 million cubic meters of sand (Orme et al. 2011), and grooming of 45% of the beaches (Dugan et al. 2003) (Figure 20.12). Recreational use of California beaches is intense; for example, in the Santa Monica Bay area and the nearby coastline of north Los Angeles County, between 50 million and 60 million visits to beaches are made annually (Dwight et al. 2007). Driving in the intertidal zone, including the recreational use of ORVs (off-road vehicles) at designated beaches and the widespread use of public safety vehicles, crushes and kills beach invertebrates (Schooler et al. in prep) and wildlife. The ecological disturbance of sandy beach ecosystems by coastal management practices can be remarkably intense in urban areas of southern California. For example, mechanical beach grooming and raking with heavy equipment—​a practice that removes all wrack and disturbs the sand of the beach to a depth of more than 6 inches—​is conducted as often as twice a day on beaches of Santa Monica Bay in Los Angeles, making beaches the most frequently disturbed ecosystem in the state. Some beaches have been groomed regularly for more than forty years. The loss of wrack subsidies from beach grooming is associated with substantial and widespread alteration in macrofauna community structure of sandy beach ecosystems in southern California (Dugan et al. 2003). This has resulted in a significant reduction in prey resources available to shorebirds over more than 160 kilometers of the California coast (Dugan et al. 2003). The mechanical disturbance of beach grooming also kills grunion eggs as they incubate in the sand near the high tide line (Martin et al. 2006). Beach grooming is also associated with loss of dune vegetation and habitats and their buffering function in southern California (Dugan and Hubbard 2010). Beach ecosystems are strongly affected by active interventions or coastal defense efforts that attempt to prevent shoreline retreat or change. Approaches to coastal defense can be soft (beach fills or nourishment) or hard (armoring) (Komar 1998), both with substantial ecological and physical consequences for beach ecosystems (e.g., Griggs 2005, Peterson et al. 2006 and 2014, Dugan and Hubbard 2006, Dugan et al. 2008). Societal responses to beach erosion and shoreline retreat have relied heavily on coastal armoring for centuries (Charlier et al. 2005, Nordstrom 2000). During the past century approximately 130 miles of California’s coast were armored, and coastal armoring increased 400% between 1971 and 1992 (Griggs 1998), a trend expected to accelerate. These shoreline alterations have reduced the width of the beach over large stretches of coastline in California (Revell and Griggs 2007). Shoreline retreat and erosion coupled with coastal armoring causes a disproportionate reduction (and in many cases the complete loss) of dry upper beach and high tide line habitat relative to wet and saturated lower beach habitats (Dugan and Hubbard 2006, Dugan et al. 2008 and 2012). This directly eliminates the zone of wrack deposition and retention, removing habitat and food for wrack-associated species (Dugan et al. 2008, Jaramillo et al. 2012). Lower species richness and abundance of shorebirds (2 and 3.7 times lower, respectively) have been reported on armored beaches by Dugan and Hubbard (2006) and Dugan et al. (2008). The abundance of roosting seabirds also declined strongly in front of armoring structures (gulls 4.8 times, seabirds 3.3 times lower) (Dugan et al. 2008, 2012). Soft coastal defense in the form of beach filling has been conducted on a massive regional scale for years in southSandy Beaches   401

A

B

C

D

E

F

FIGURE 20.12 Human impacts to beach ecosystems are many and occur on a variety of scales, frequencies, and intensities. Photos: (A) Eduardo Jaramillo, (B, F) Jenifer Dugan, (C–​E) David Hubbard.

Widespread habitat loss is caused by coastal armoring with engineered seawalls and revetments, including (A) seawalls in the intertidal zone (Summerland, California) and (B) revetment/seawall combinations such as this large structure in Pacifica, California. (C) Beach filling or nourishment can provide beach space for recreation, but this intensive disturbance defaunates the intertidal zone, reducing prey resources for birds and fish for extended periods (Goleta Beach, California). (D) Likewise, mortality of intertidal animals caused by seasonal scraping of the intertidal zone to create winter sand berms seaward of homes and infrastructure can be substantial (Carpinteria, California). (E) The widespread practice of beach grooming, raking, and sifting with heavy equipment to remove trash and wrack is conducted weekly on many beaches and up to twice a day on popular beaches in Santa Monica Bay, disturbing beach ecosystems more frequently than any agricultural activity (Carpinteria, California). (F) Vehicle driving in the intertidal zone, including the recreational use of ORVs, occurs at selected beaches, and the widespread use of public safety vehicles, crushes and kills beach invertebrates and wildlife (Oceano Dunes, California).

ern California (Flick 1993, Orme et al. 2011) with little to no scientific evaluation of the direct or cumulative ecological effects on beach ecosystems (Peterson and Bishop 2005). Despite a dearth of information from California, the ecological impacts of beach filling on beach biota are severe, often resulting in 100% mortality of resident fauna (Speybroek et al. 2006) with lasting effects propagating up the food web to shorebirds (Peterson et al. 2006). Recovery of important invertebrate species can take years (Peterson et al. 2000, 2006, 2014). Use of fill sediments finer or coarser than the native beach sand causes even greater and longer lasting ecological impacts to biota (Peterson et al. 2006 and 2014, Speybroek et al. 2006, Viola et al. 2014). Sea level rise and other predicted effects of climate change, including increased storminess, are expected to intensify pressures on beach ecosystems by increasing rates of shoreline erosion and retreat and degrading habitat (Nordstrom 2000, Slott et al. 2006), especially where coastal land uses and development constrain shoreline evolution and retreat. The direction of responses of beach ecosystems to sea level rise may be similar to that observed in episodic El Niño Southern Oscillation (ENSO) events (Revell et al. 2011) although the time scale will differ greatly. The habitat loss, fragmentation, and alteration resulting from climate change carry profound ecological implications, as beaches become narrower, steeper, and coarser and as once continuous stretches of sandy beach are interrupted by submerged coast or drowned beaches. Although beaches are often assumed to be robust, disturbance-adapted ecosystems, strong and lasting negative responses suggest that beaches are in fact sensitive to anthropogenic pressures and impacts and sometimes slow to recover 402  Ecosystems

(McLachlan et al. 1996, Hubbard et al. 2014, Peterson et al. 2014). Bounded by land and sea, sandy beach ecosystems are increasingly squeezed between the impacts of coastal development and the manifestations of climate change in the sea (Schlacher et al. 2007, Nordstrom 2000, Defeo et al. 2009, Dugan et al. 2010). Human alterations limit the ability of beach ecosystems to adjust to changes in shoreline stability (Clark 1996), sea level rise, and erosion caused by climate change.

Conservation and Restoration Strategies Conserving beaches first requires recognizing beaches as ecosystems in coastal policy and management. The fact that a number of sandy beaches are now included as part of a broad network of marine protected areas (MPAs) across the state, protecting harvested species, such as clams and fish, in many regions is a major step forward. However, the majority of these MPAs extend only up to the mean high tide line, meaning many of the highly mobile beach biota, including a wide range of intertidal invertebrates as well as nesting California grunion, regularly use areas that are outside the MPA boundaries. The lack of restrictions on impacts, such as beach grooming, vehicle use, or beach filling, in current MPA regulations is also of concern for some beaches. Nonetheless, the beaches now in MPAs could be used to evaluate responses of beach ecosystems to increased protection from disturbance. Changes in disturbance directly linked to management practices and recreation, such as beach grooming practices and vehicle use, offer low-cost opportunities for conserva-

tion in California and elsewhere in the world. For example, an approach that designates beach areas to be left undisturbed by grooming from their landward boundaries to the sea could provide much needed restoration opportunities on developed coastlines. This approach, if interspersed with groomed areas, could conserve biodiversity in designated areas while allowing traditional recreational and tourism to continue in others. As sea level rises, beaches with enough space to evolve and retreat landward combined with a sufficient sand supply will be able to adjust to changing water levels and maintain ecosystem integrity. Where retreat is constrained by resistant cliffs or armoring and infrastructure, beach ecosystems and their biota and functions will disappear as sea level rises. This is of particular concern for endemic biota already restricted to bluff-backed beaches in some littoral cells (Hubbard et al. 2014). Promoting the use of ample setbacks for new coastal development and identifying locations and opportunities where infrastructure can be removed as part of “managed retreat” to allow beaches to evolve and migrate landward can increase opportunities to maintain and conserve the diversity and ecosystem function of beaches as sea level rises. Moving or abandoning infrastructure to allow room for coastal retreat is already part of the management strategy for California State Parks. One of the best California examples of the use of managed retreat to restore open coast beach and dune ecosystems is a 2010 project at Surfer’s Point near the mouth of the Ventura River, in Ventura County (http://surferspoint.org/). The Surfer’s Point project removed a rock revetment and relocated a parking lot and bike path to give the beach 20 meters of space to migrate landward. The project also restored dunes and coastal strand vegetation in the area where the parking lot was removed.

Summary Composed of sand and biota that are constantly moving across and along the shoreline, sandy beaches are among the most dynamic ecosystems in the world. Dominating open coastlines of California, beaches are iconic assets highly prized for recreation and coastal economies. Less appreciated is the fact that beaches are ecosystems that harbor unique biodiversity, support productive food webs, and provide irreplaceable ecosystem functions and services to society. These include filtering vast volumes of seawater and buffering the land from storm waves. Although their characteristic, highly mobile burrowing animals are often invisible to a casual visitor, California beaches can be hotspots of intertidal biodiversity. Subsidies of kelp and phytoplankton from other marine ecosystems to beach food webs fuel intense biological productivity capable of supporting a high abundance of wildlife, rapid processing of organic inputs, and high rates of nutrient remineralization. However, ongoing disturbance and escalating threats to beach ecosystems pose formidable challenges in California and elsewhere. Groomed beaches in urban areas are subject to the most intense disturbance regime of any ecosystem in the state. Beaches are increasingly trapped in a “coastal squeeze” between urbanization and effects of sea level rise from climate change. Societal responses to beach erosion and retreat rely largely on “soft” (beach filling) or “hard” (shoreline armoring) engineering that both affect the biodiversity, food webs, and functioning of beaches as ecosystems. The wide-

spread but unsupported assumption that beach ecosystems recover very rapidly from all forms of disturbance is used to justify numerous management actions. Increased recognition and understanding of sandy beaches as vulnerable and threatened ecosystems are needed to promote the conservation and protection of these dynamic ecosystems on the edge of a warming and rising sea.

Acknowledgments We dedicate this chapter to Betsy Hubbard for decades of inspiration and encouragement. We thank Karina Nielsen and an anonymous reviewer for their helpful comments on the manuscript. We are grateful to Karen and Doug Martin of Pepperdine University, Shane Anderson, Callie Bowdish and Nicholas Schooler of the University of California at Santa Barbara, and Dan Ayres of the Washington Department of Fish and Wildlife for allowing us to use their photographs in this chapter. Financial support was provided by the National Science Foundation’s Long-Term Ecological Research program (OCE-1232779) (JED) and by the California Sea Grant Project R/MPA-24 through the California Ocean Protection Council.

Recommended Reading Griggs, G., K. Patsch, and L. Savoy. 2005. Living with the changing California coast. University of California Press, Berkeley, California. McLachlan, A., and A. C. Brown. 2006. The ecology of sandy shores. Academic Press, Elsevier, London, UK. Nordstrom, K. F. 2000. Beaches and dunes of developed coasts. Cambridge University Press, Cambridge, UK. Pilkey, O. H., W. J. Neal, J. T. Kelley, and J.A.G. Cooper. 2011. The world’s beaches: A global guide to the science of the shoreline. University of California Press, Berkeley, California. Pilkey, O.H, and J. A. G. Cooper. 2014. The Last Beach. Duke University Press. Durham, North Carolina.

Glossary Allochthonous  Originating from outside the system. Bedload  The portion of the total sediment in transport that is carried by intermittent contact with the seabed by rolling, sliding, and bouncing. Berm  A raised ridge of sand found at high tide, or storm tide marks on a beach. Berm crest  The highest portion of the ridge of sand near the high tide line on a beach. Bottom-up effects  In reference to food webs, influences on the system that are driven by nutrient supply or productivity rather than top-down such as predators. Breakwater  A human-made armoring structure designed to absorb the energy of the waves before they reach the shoreline. Broadcast spawning  Releasing of gametes (sperm and eggs) into open water for external fertilization with no subsequent parental care. Chlorophyll-a  A pigment that is essential for photosynthesis; extracts of chlorophyll-a from water samples can be used as estimates of phytoplankton quantities and productivity. Coastal strand vegetation  The plant community between the high tide line and the foredune. Sandy Beaches   403

Crenulate spit  The term describing a spit that forms at the end of a beach with a wavy or scalloped shoreline form. Cross-shore  In a direction perpendicular to the shoreline; also termed “shore-normal.” Crypsis  The ability of an organism to avoid observation or detection by other organisms using methods including camouflage, nocturnal or subterranean lifestyle, transparency, and mimicry. Cusps  Rhythmic alongshore features of soft sandy shorelines that consist of repeated arc-shaped patterns of horns pointing seaward alternating with concave bays. Cusps represent a combination of constructive and destructive coastal processes. The spacing of bays and horns in cusps is related to wave energy and is often fairly uniform with intervals of 20 to 60 meters. Megacusps associated with strong waves and rip circulation with spacing of >200 meters can form on some beaches. A cusp horn is the pointed seaward projections of sand on the shoreline that separate cusp bays (the concavity in the shoreline between cusp horns). Diatom  Unicellular phytoplankton that are enclosed within a cell wall made of silica (hydrated silicon dioxide) called a frustule. Diatoms are among the most common types of phytoplankton and are important primary producers within the food chain. Direct development  Lacking any planktonic larval stages, direct developing species have very low dispersal potential during larval development. The reproduction of resident individuals is thus important in maintaining populations at a given location. These larvae are also known as “crawl-away larvae,” since forms with this type of development may have larvae crawl away from the brooding female or the egg mass. Dissipative  Refers to surf zones and beach systems with waves that break far from the intertidal zone and dissipate their force progressively along wide surf zones that are flat in profile. Dissipative beaches tend to be wide and have fine sediments and flat slopes. Downcoast or upcoast  Refers to the location of a beach with respect to the prevailing littoral current and a feature that affects sand transport and supply. These features can include a rocky point, a sand source or sink, or a barrier, such as a harbor or a groin. El Niño  Also termed the El Niño Southern Oscillation or ENSO, this is expressed as a band of anomalously warm ocean water temperatures and atmospheric anomalies that periodically develop off the Pacific coast of South America at intervals of two to seven years. Its effects often extend to North America and can be associated with strong climatic effects in California, including altered storm tracks, increased storm intensities, higher sea levels, and increased precipitation. Foredune  A dune ridge that runs parallel to the shore of an ocean, lake, bay, or estuary. In active dune systems the foredunes are located closest to the sea or other body of water. Groin  A human-made armoring structure that is placed perpendicular to the shoreline to retain sand by stopping or slowing the littoral transport of sand along the shoreline. High tide strand  The driftline or tidal high water mark, a shoreline feature where the deposition of buoyant debris and the boundary of damp sand marks the highest extent of tides and wave run-up. Infauna  Organisms that live within the sediments rather than on the surface. Jetty  A human-made armoring structure that extends across the beach from the mouth of a harbor, lagoon, or river into deeper water. Jetties are often placed in parallel pairs and stabilize entrance channels needed for shipping.

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Laminar  Describes flow in which fluid moves in parallel layers, without disruption among the layers; typically this occurs at lower velocities than more turbulent flow patterns. Littoral current  A current located in the littoral zone (an indefinite zone extending seaward from the shoreline to just beyond the surf zone) that generally moves parallel to the shoreline. This current is generated by waves breaking at an angle to the shoreline and is also called the longshore current. Littoral transport  The movement of sedimentary material by waves and currents in the littoral zone. Commonly used as synonymous with longshore transport. Often expressed as a rate of volume per year. Log spiral  A type of beach that develops in the shelter of a prominent headland. These logarithmic spiral shapes in map view may also be called “half heart” or “crenulate” or “headland bays.” Longshore  In a direction parallel to the coast. Macrophyte  Aquatic plant large enough to be seen without magnification that grows in or near water. Mean high tide  The average elevation of the high tides. Mega-cusp  Embayments in the shoreline (usually several hundred meters long) associated with rip currents. Megalopa/postlarvae  The final developmental stage of a decapod crab before it becomes a juvenile crab. This stage follows the larval stages. Morphodynamic scale  A description of the shape of a beach resulting from the interaction of the wave regime with the available sand. This scale encompasses beach states across the spectrum from fully dissipative to fully reflective conditions. Mysid  Shrimplike crustaceans in the superorder Peracarida that brood their young. Neap  Designating a tide that occurs just after the first and third quarters of the moon, when there is least difference between high tide and low tide levels. Pacific Decadal Oscillation  A long-lived El Niño–​like pattern of Pacific climate variability with cycles extending over twenty to thirty years that is detected as warm or cool surface waters in the Pacific Ocean, north of 20°N. Palp  An oral appendage in some invertebrates that plays roles in sensory perception, feeding, and locomotion. Pinnipeds  Semiaquatic marine mammals including seals, sea lions, fur seals, and walrus. Plankton  A diverse group of organisms that live in the water column and cannot swim against a current. Phytoplankton refers to photosynthetic organisms; zooplankton refers to animals that live in the plankton for all or part of their lives. Reflective  Refers to beaches with narrow, shoaling surf zones, steep slopes, and coarse sand, with waves break abruptly on the intertidal zone. Rip current  A seaward flow of water from near the shore, typically through the surf line. Rip currents are usually generated by the energy of breaking waves and develop through gaps in sand bars in the surf zone. Runnel  A shallow trough or low point in the beach that runs parallel to the shoreline where water can pool temporarily. On wide flat beaches with moderate wave energy, runnels can form in a series separated by ridges. Runnels can also develop landward of a berm that has a backslope from its crest. Saprophagous  Feeding on dead or decaying organic matter. Scarp  A steep abrupt feature on the beach face that is caused by erosion.

Seabird  Birds that have adapted to life within the marine environment, including gulls, terns, cormorants, pelicans, and petrels. Sediment budget  The balance between sediment inputs and losses from the coastal littoral system. Shorebird  Birds that are members of the order Charadriiformes. These include plovers, stilts, avocets, oystercatchers, and sandpipers. These birds feed by wading in shallow waters or at waterlines and probing into the water or sand for insects, polychaetes, mollusks, and crustaceans. Shorebirds are found frequently on beaches, marshes, wetlands, mudflats, rocky, and inland shores. However, shorebirds are not confined to these areas and many can also be found in open fields and agricultural areas near water sources. Spring tides  Designating a tide that occurs near full or new moons, when there is maximum difference between high tide and low tide levels. These tides occur when the moon and sun are in alignment. Submarine canyon  A steep-sided valley cut into the sea floor of the continental slope, sometimes extending well onto the continental shelf. Supralittoral  Coastal zone above the reach of the tides. Surf zone  The region of breaking waves that forms near the shoreline. Suspension-feeding  Describing animals that obtain their food from material carried in the water column, often by filter-feeding. Swash climate  The characteristics of a beach’s swash zone in which a turbulent layer of water washes up on a beach after a wave has broken, such as length, period, and speed. Upwelling  An oceanographic phenomenon that involves the wind-driven movement of dense, cooler, and usually nutrient-rich water from deeper depths to the ocean surface, replacing the warmer, usually nutrient-depleted surface water. This nutrient-rich upwelled water stimulates the growth and reproduction of primary producers such as phytoplankton. Due to the biomass of phytoplankton and presence of cool water in these regions, upwelling zones can be identified by cool sea surface temperatures (SST) and high concentrations of chlorophyll-a. Water table (or aquifer)  The underground layer of waterbearing unconsolidated sediments (gravel, sand, or silt). Water table outcrop  The location where the water table emerges on the beach face; also called the effluent line. Wrack  Floating material, such as macroalgae and seagrass, that is deposited on the beach.

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California Department of Fish and Game. . Accessed February 2014. ———. 2001b. Spotfin croaker. Pages 230–​231 in W. S. Leet, C. M. Dewees, R. Klingbeil, and E. J. Larson, editors. California’s living marine resources: A status report. Accessed February 2014. Viola, S. M., D. M. Hubbard, and J. E. Dugan. 2014. Burrowing inhibition by fine textured beach fill: Implications for recovery of beach ecosystems. Estuarine, Coastal, and Shelf Science 150:142–148. Warrick, J. 2013. Dispersal of fine sediment in nearshore coastal waters. Journal of Coastal Research 29:579–​596. Wenner, A. M. 1988. Crustaceans and other invertebrates as indicators of beach pollution. Pages 199–​229 in D. F. Soule and G. S. Kleppel, editors. Marine Organisms as Indicators. Springer-Verlag, New York, New York. Wenner, A. M., J. E. Dugan, and D. Hubbard. 1993. Sand crab population biology on the California Islands and mainland. Pages 335–​348 in F. G. Hochberg, editor. Third California Islands Symposium, Recent Advances in Research on the California Islands. Santa Barbara Museum of Natural History, Santa Barbara, California. Willis, C. M., and G. B. Griggs. 2003. Reductions in fluvial sediment discharge by coastal dams in California and implications for beach sustainability. Journal of Geology 111(2):167–​182. Winant, C. D., D. L. Inman, and C. E. Nordstrom. 1975. Description of seasonal beach changes using empirical eigenfunctions. Journal of Geophysical Research 80:1979–​86. . Wright, L. D., and A. D. Short.1980. Morphodynamic variability of surf zones and beaches: A synthesis. Marine Geology 70:251–​2 85. Yaninek, J. S. 1980. Beach wrack: Phenology of a limiting resource and utilization by macroinvertebrates of sandy beaches. MA thesis. University of California, Berkeley, California. Young, A.P., and S. A. Ashford. 2006. Application of airborne LIDAR for seacliff volumetric change and beach-sediment budget contributions. Journal of Coastal Research 22(2):307–​318. ZoBell, C. E. 1971. Drift seaweeds on San Diego County beaches. Nova Hedwigia 32:269–​314.

T WENT Y-ONE

Coastal Dunes PE TER ALPERT

Introduction One of the great attractions of California is its ocean coast, a strip approximately 1,800 kilometers long of terraces and headlands punctuated by coves, bays, estuaries, and accompanying beaches and sand dunes (Griggs et al. 2005, Griggs 2010, California Coastal Commission publications at www .coastal.ca.gov, aerial photographs at www.californiacoast line.org). About 13% of this strip consists of cliffs over 100 meters tall that are relatively resistant to erosion, and another 59% is lower, more erodible cliffs (Runyan and Griggs 2003). Most of the rest is beaches backed by dunes (Cooper 1967; see Chapter 20, “Sandy Beaches”). This chapter focuses on these coastal dunes—​lands a few tens of meters to several kilometers across, made of sand blown inland from the beach, colonized by distinctive communities of plants and animals, and visited for pleasure by people. Other chapters cover the ocean beaches (Chapter 20, “Sandy Beaches”) and rocky intertidal systems (Chapter 18) that lie just seaward of dunes; the estuaries (Chapter 19, “Estuaries: Life on the Edge”), wetlands (Chapter 31), forests (Chapter 26, “Coast Redwood Forests”), grasslands (Chapter 23), shrublands (Chapter 22, “Coastal Sage Scrub,” and Chapter 24, “Chaparral”), farmlands (Chapter 38, “Agriculture”), and cities (Chapter 39, “Urban Ecosys-

tems”) that adjoin dunes on their inland edge; and the dunes on the islands off the coast of California (Chapter 34, “Managed Island Ecosystems”) and in the interior deserts (Chapter 30). Among California’s terrestrial ecosystems, coastal dunes are particularly dynamic. As long as they are partly bare, they move in the wind and shift underfoot. This creates opportunities and problems for organisms, including humans. Plants and animals find open sites to colonize but suffer from burial or excavation. Humans find recreation, but moving dunes can cover homes, roads, and harbors. Both plants and humans respond by stabilizing dunes. In California, people have introduced plants from other continents to hold down dunes, and some of these introduced species have spread so abundantly that coastal dune ecosystems dominated by native species have shrunk to a few.

Geography Dunes occur along the Californian coast in disjunct patches that range in size from less than 1 hectare to over 100 square 409

kilometers (Figure 21.1; Zedler 1962, Cooper 1967). The California Coastal Commission (1987) listed twenty-seven dune fields. The two largest were at Monterey Bay, covering about 105 square kilometers; and at Nipomo, covering 47 square kilometers. Barbour and Johnson (1988) named thirteen major dune localities: 1. Crescent City (likely corresponding mainly to dunes at the mouth of the Smith River and Lakes Earl and Tolowa in Tolowa Dunes State Park). 2. Humboldt Bay (the North and South Spits, including the Lanphere and Ma-le’l Dunes in Humboldt Bay National Wildlife Refuge). 3. Fort Bragg (south of the mouth of Ten-Mile River, largely in MacKerricher State Park). 4. Point Arena (Manchester Beach, including Manchester Beach State Park). 5. Bodega Beach (from the mouth of Salmon Creek south to Bodega Head). 6. Dillon Beach east of the mouth of Tomales Bay. 7. Point Reyes (Kehoe Beach south to North Beach, notably near Abbotts Lagoon, plus low dunes on Limantour Spit). 8. San Francisco (once very extensive, then almost entirely built or planted over, now restored in areas at Bakers Beach and south of Ocean Beach). 9. Monterey (mostly built or planted over but with remnants at the mouth of the Salinas River [Bluestone 1981] and at Marina and Asilomar State Beaches). 10. Morro Bay. 11. The Santa Maria River complex (likely the area between San Luis Obispo Bay and Point Conception, including dunes at Oceano, Oso Flaco, Guadalupe, Nipomo, Pismo Beach, the Santa Maria River, Vandenberg Air Force Base, and the Santa Ynes River). 12. Los Angeles (almost entirely built or planted over, including former low dunes on barrier spits [Engstrom 2006] and remnant areas at El Segundo Dunes). 13. San Diego Bay (almost entirely built or planted over). Additional, smaller areas of relatively intact dunes include those on Gold Bluff Beach north of Prairie Creek in Prairie Creek State Park (a narrow zone of low dunes between beach and tall cliffs), at the mouth of the Little River (a narrow zone), at the mouth of the Mattole River, in Año Nuevo State Reserve at Franklin and North Points, at Coal Oil Point Reserve in Santa Barbara, and at the mouth of the Santa Margarita River in Camp Pendleton Marine Base north of Ocean­ side (a narrow zone). Most of California’s coastal dunes have been highly transformed by human use and introduced plants. Barbour and Johnson (1988) cited Humboldt Bay and Point Reyes as the least altered of the major dune localities north of Big Sur (see Figure 21.1). Areas at Lanphere and Male’l Dunes at Humboldt Bay and near Abbotts Lagoon on Point Reyes, plus a smaller area at Franklin Point in Año Nuevo State Reserve, probably now support the best remnants of native dune vegetation in

Photo on previous page: Native vegetation on partly stabilized dunes at Point Reyes National Seashore. In the foreground are seaside daisy (Erigeron glaucus, in flower), yellow sand verbena (Abronia latifolia, with green, succulent leaves), and Tidestrom’s lupine (Lupinus tidestromii, with silvery leaves, a federally and state endangered species). Photo: Sarah Minnick, courtesy of the National Park Service. 410  Ecosystems

northern California (Figure 21.2), followed by the dunes at the mouths of the Ten-Mile and Salinas Rivers and at Asilomar. Of the major dune localities south of Big Sur, Barbour and Johnson (1988) cited Morro Bay and parts of the Santa Maria complex as the least altered. Peinado et al. (2007) identified these parts as the dunes at Guadalupe, Nipomo, Pismo Beach, and Oso Flaco. However, considerable portions of these dunes have been heavily affected by off-road driving and introduced plants. Along with the rest of the immediate coast of California, coastal dunes enjoy moderate temperatures (see Chapter 2, “Climate”). Mean monthly air temperature at a given site varies by 9°C or less during the year, and mean annual air temperature rises from north to south by only a few degrees, from 11°C to 17°C (Barbour and Johnson 1988). A few nights of frost per year occur on average along the northern and central coast; snow is virtually unknown. This moderation stems in part from relatively low seasonal change in the surface temperature of the ocean near the coast (see Chapter 16, “The Offshore Ecosystem”) and a related prevalence of convection fog in the summer, when the daily maximum temperature can be 15°C higher 10 kilometers inland than at the coastline. Precipitation on coastal dunes does vary greatly with both season and latitude. The entire coast receives much more rain in the winter than in the summer, and the southern coast is much drier than the north. The dry season in summer increases from two months with less than 1.5 centimeters of rain each in Crescent City to five months with less than 0.5 centimeter each in San Diego. Mean total annual precipitation decreases from approximately 165 to 25 centimeters north to south. The central and southern coastal dunes have a Mediterranean-type climate. Finally, wind speed is often high at the coast, particularly in fair weather in spring. Areas of low air pressure over the Pacific in the rainy season cause winds to tend to come from the south and southwest, while high pressure in the dry season causes north and northwesterly winds to predominate. In sum, the climate features wet and dry seasons rather than cold and warm ones, potentially permits year-round activity of many organisms except as limited by drought, and is relatively windy.

Dune Formation and Shape Coastal dunes in California (Wiedemann and Pickart 2004, Griggs et al. 2005) and in general (Nordstrom 2000, McLachlan and Brown 2006, Maun 2009) are formed by sand blown inland from an adjoining beach. Measurements and models of wind-driven transport of sand indicate that grain size, surface moisture, microtopography, and a variety of aspects of air movement are important (e.g., Namikas 2003, Sherman and Li 2012). Brief, high-velocity movements account for a considerable amount of transport (Craig 2000). In California the movement of sand from beach to dunes is just the last step in a long chain of supply. This begins with the erosion of lands in watersheds as far east as the crest of the Sierra Nevada and continues with transport of sediments to the coast in rivers and then along the coast by currents within littoral cells, followed by deposition onto beaches by waves (Peterson et al. 2010; see also Chapter 20, “Sandy Beaches”). Erosion of coastal bluffs also provides some of the sediments that become dunes. However, Runyan and Griggs (2003) estimated that the contribution of bluffs to deposition of sand on beaches at various sites in the state ranged from less than 1% to at least 12%. Soils on

FIGURE 2 1.1 Locations of systems of coastal dunes in California. Data from U.S. Geological Survey and ESRI. Map: P. Welch, Center for Integrated Spatial Research (CISR).

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FIGURE 2 1.2 Native vegetation on dunes at (A) Lanphere and Ma-le’l Dunes, and (B) Abbotts Lagoon, Point Reyes National Seashore. Photos: (A) Andrea Pickart, (B) Peter Baye.

FIGURE 2 1.3 Coastal dune system at Vandenberg Air Force Base. Photo: Andrea Pickart.

Californian coastal dunes are generally dominated by silicates such as quartz and feldspar derived from inland areas (Muhs et al. 2009). Coastal dunes in California are in good part recycled mountains and hills. Californian coastal dunes date from the Pleistocene and Holocene (Cooper 1967, Barbour and Johnson 1988, Pickart and Barbour 2007). The Pleistocene dunes have largely lost their mobility and dune-like features, and it is the Holocene dunes that are generally considered in treatments of dune ecology and that account for the dune systems treated here. These Holocene dunes often overlie other formations. For instance, some dunes in the Guadalupe-Nipomo system were laid atop estuarine deposits approximately 4,300 to 3,500 years ago (Knott and Eley 2006). In other cases, sand has blown up onto low coastal terraces, creating perched dunes as at Franklin Point. Peterson (2006) provides dating and morphostratigraphy of dune sheets in northern and central California, and Peterson et al. (2007) date dunes in central and southern Oregon. Especially in northern and central California, dune systems have a ridge, or foredune, that runs along the inland edge of the beach. If there is a large enough area of wind-driven sand, a system may form numerous transverse dunes, ridges

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that run perpendicular to the prevailing wind; or parabolic dunes, curved ridges whose arms point upwind. Sets of ridges can move downwind more or less in unison, creating a mobile landscape or dune sheet (Figure 21.3) of relatively sheltered and sometimes moist troughs, windward faces where objects are excavated, relatively wind-blown and dry ridge tops, and lee faces where objects are buried. Especially in the wetter climates of northern California, sizable areas upwind of dunes can be eroded down to the water table, forming deflation plains or basins. Small seasonal wetlands, some of which are termed dune swales or slacks, and even small permanent lakes occur in dune systems in California, such as Fen Lake at Ten-Mile Dunes and Oso Flaco Lake at Oceano Dunes. Wind may also hollow out portions of dune ridges, forming blowouts (Hesp 2002). Various sets of terms distinguish dunes at different distances inland, such as nearshore dunes, moving dunes, and back dunes (Pickart and Barbour 2007). Hesp (2007) and Barchyn and Hugenholtz (2012) discuss classification and formation of dunes. Plants modify the shapes and movement of dunes both by holding sand down and by intercepting blowing sand. Worldwide, the most complex dunes may occur where disturbance by wind and stabilization by plants are roughly equal (Doing 1985). One dramatic demonstration of the influence of plants on dune morphology in California has been the remodeling of the foredunes by the introduced European beach grass (Ammophila arenaria) (Figure 21.4). Before this grass was introduced, the backs of beaches in southern California were generally colonized by taprooted, perennial forbs that formed a zone of large mounds of sand. The backs of beaches and foredunes in northern California were colonized by forbs and grasses including the native, rhizomatous American dune grass (Elymus mollis), which formed a low, gradual, partly open ridge (Barbour and Johnson 1988). European beach grass has now replaced American dune grass on foredunes just about everywhere in unrestored areas in California except near Abbotts Lagoon at Point Reyes and on the Lanphere and Ma-l'el Dunes at Humboldt Bay (Wiedemann and Pickart 2004). Concurrently, the gradual foredunes in northern California have been replaced by steeper, taller, more densely vegetated ones that are more effective at trapping sand (Zarnetske et al. 2012). This tends to truncate the back of the beach and cut off sand supply to more

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FIGURE 2 1.4 Foredunes dominated by (A) European beach grass (Ammophila arenaria) at Bodega Marine Reserve (B) native plants such as American dune grass (Elymus mollis) at Abbotts Lagoon. Photos: (A) Jackie Sones, (B) Peter Baye.

inland dunes, which then become more densely vegetated and stable. This change in the shape of foredunes appears due largely to the contrasting responses of European beach grass and American dune grass to burial. European beach grass produces dense, vertical rhizomes and tillers when buried, whereas American dune grass increases lateral spread. The growth response of European beach grass sets up a positive feedback, trapping more sand that induces further vertical growth. Where European beach grass grows on dunes behind the foredune, it can also raise their height; at Bodega Bay the height of dunes with European beach grass increased 4 centimeters per year during the 1900s, while the height of nonvegetated areas decreased at about the same rate (Danin et al. 1998). In Oregon, where European beach grass had spread over all foredunes by 1983 except on one island at the mouth of the Columbia River, a new phase of remodeling has begun (Hacker et al. 2012): another introduced grass, American beach grass (Ammophila breviligulata), is replacing European beach grass. American beach grass has a growth form somewhat intermediate between those of European beach grass and American dune grass, and the foredunes are getting lower again.

Physical Conditions for Organisms Physical factors likely to limit the performance of organisms on dunes in California include burial or excavation by movement of sand, low soil nutrient availability, low water availability, desiccation by wind, high solar radiation, and high salinity in soil and aerosols (e.g., Fink and Zedler 1990, Wilson and Sykes 1999, Pickart and Barbour 2007, Cornelisse and Hafernik 2009). Little work seems to have been done to characterize these factors since about 1990. Pickart and Barbour (2007) reviewed the literature and concluded that many of these factors are important. They note that burial may be particularly so, although burial can have positive effects on some species and negative effects on others (e.g., Zedler et al. 1983, Zhang and Maun 1992, Bonte et al. 2006). Availability of nitrogen on bare dune soils is very low. Cushman et al. (2010) found a mean concentration of 5–​7 µg inorganic nitrogen (N) per gram dry mass of soil and a net mineralization rate of 2 µg N per gram soil per month at 0–​5 centimeters below the soil surface during February

on a stabilized dune at Bodega Bay. Since nutrient levels are generally higher on older dunes and nearer the soil surface, and highest in winter at this site, these values are likely to be maximal. For instance, Alpert and Mooney (1996) reported a mean concentration of 0.5–​1.2 µg inorganic N per gram dry mass of soil and net mineralization of 0.06 µg N per gram per month at 0–​15 centimeters depth in summer on partly stabilized dunes at Franklin Point; concentrations were no more than 0.2 µg per gram at 15–​6 0 centimeters depth. By comparison, Maron and Jefferies (1999) measured net mineralization rates of about 4 and 10 µg N per gram soil per month at 5–​10 centimeters depth away from and under nitrogenfixing shrubs, respectively, in grassland adjacent to dunes at Bodega Bay. Availability of water may not be as low as is sometimes assumed. The upper 20–​100 centimeters of the soil on dunes can be very dry when deeper soil is very moist, forming surprisingly distinct layers. Measurement of the predawn water potential of American dune grass on foredunes in Oregon suggested that soil water potential in its rooting zone never fell below –​0.2 MPa during the driest part of the year (Pavlik 1985), a value unlikely to limit growth of terrestrial plants. However, plants did display evidence of midday water stress such as partial closure of stomata, suggesting that high solar radiation or wind made the habitat effectively dry at certain hours. Rapid drainage and evaporation from the coarsely textured soils on dunes can dry the upper soil quickly and likely makes drought a limiting factor in the establishment of seedlings and the performance of shallowly rooted plants. The dry upper layer of sand can bar movement of water upwards from deeper sand, resulting in dry and wet layers. Recent measurements on dunes in Korea emphasize the high variability between levels of both soil moisture and nutrients at different distances from the sea and in different topographic positions (Kim et al. 2008). On beaches, salt spray and overwash can strongly affect plants (Pickart and Barbour 2007). However, negative effects of high salinity on dunes may be largely limited to the seaward edges of areas of dunes and the times when unusually strong winds blow spray inland or unusually high waves overtop foredunes. Marine aerosols transfer sodium (Na) and other cations to the leaves of shrubs on dunes and thence potentially to soil (Clayton 1972), but inputs of Na, potassium (K), magnesium (Mg), and calcium (Ca) in salt spray can

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roughly equal losses from leaching at 1 to 2 meters in depth in the soil. There appear to be no published studies on the intensity or frequency of fire on coastal dunes.

Macroorganisms Species Most of California’s dunes are dominated by perennial, herbaceous to shrubby plants. Trees dominate some stabilized dunes from Mendocino County north and near Monterey. The fully resident animals on dunes are mostly invertebrates, a few rodents, and possibly rabbits. However, dune forests have additional vertebrate residents, vertebrates from adjacent habitats seek food on dunes, and dunes and beaches serve mammals as relatively open avenues of travel (Moore 2002). A few species of birds such as the western snowy plover (Charadrius nivosus nivosus) nest, and some marine mammals such as northern elephant seal (Mirounga angustirostris) rest where dunes meet beaches. One photogenic mammalian use of coastal dunes in California is by Roosevelt elk (Cervus canadensis roosevelti) that travel between inland meadows and the ocean at Prairie Creek State Park. Dunes also support annual plants, bryophytes, lichens, and macrofungi, including some distinctive species of mosses and lichens (Danin et al. 1998, Glavich 2003, Cushman et al. 2010). The first species of plant described from California was the beach and dune species pink sand verbena (Abronia umbellata), collected in 1786 at Monterey Bay and named a few years later by a British botanist based on plants grown in a garden in Paris (Beidleman 2006). Many species of herbaceous plants and some species of shrubs and insects and other invertebrates live only on dunes. The natural communities on dunes thus are very different from those in the adjacent coastal forests, grasslands, and shrublands. Plant traits associated with growth on dunes include succulence, trailing stems, clonal growth via rhizomes or stolons, large taproots, and pale color of leaves or stems due to dense hair or thick wax (see Figure 21.5). The distinctiveness of the natural communities on dunes generally diminishes with distance from the ocean and with the length of time that dunes have been covered and stabilized by plants—​t hat is, with the degree to which a dune habitat has become more like the adjacent habitat. Several semipopular guides to the dune plants of California (e.g., Munz 1964, Dawson and Foster 1982), and many scientific publications on dune vegetation at individual sites and along the coast (e.g., as reviewed in Barbour and Johnson 1988 and Pickart and Barbour 2007) are available. Listings of endangered, threatened, and rare species by the California Natural Diversity Database (CNDDB, dfs. ca.gov/biogeodata/cnddb) and the California Native Plant Society (CNPS, rareplants.cnps.org) suggest that a number of plant species of conservation concern are essentially limited to coastal dunes in California. Federally endangered or threatened species are Menzies’s wallflower (Erysimum menziesii, family Brassicaceae), beach layia (Layia carnosa, Asteraceae), Nipomo Mesa lupine (Lupinus nipomensis, Fabaceae), and Tidestrom’s lupine (Lupinus tidestromii, Fabaceae). Additional, state-listed endangered or threatened plant species are surf thistle (Cirsium rhothophilum, Asteraceae, also found in coastal bluff scrub [Zedler et al. 1983]) and beach spectacle

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pod (Dithyrea maritima, Brassicaceae [Aigner 2004]). Additional species listed by CNPS as endangered, threatened, or rare include Nuttall’s acmispon (Acmispon prostratus, Fabaceae, formerly Lotus nuttallii), coastal goosefoot (Chenopodium littoreum Chenopodiaceae), round-headed Chinese houses (Collinsia corymbosa, Scrophulariaceae), and sand-loving wallflower (Erysimum ammophilum, Asteraceae). These do not include subspecific taxa, species that are rare in California

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FIGURE 2 1.5 Some distinctive dune plant species. Photos: (A, E) Jackie Sones, (B–​C) Andrea Pickart, (D, F–​H ) Peter Baye. A Yellow sand verbena (Abronia latifolia) at Bodega Marine Reserve. B Beach bur (Ambrosia chamissonis) at Lanphere and Ma-l’el Dunes. C Beach morning glory (Calystegia soldanella) at Ten-Mile Dunes. D Menzies’ wallflower (Erysimum menziesii) at Abbotts Lagoon. E Silky beach pea (Lathyrus littoralis) at Wrights Beach, Sonoma

County. F Beach layia (Layia carnosa) at Abbotts Lagoon. G Tidestrom’s lupine (Lupinus tidestromii) at Abbotts Lagoon. H Douglas’ bluegrass (Poa douglasii) at Abbotts Lagoon.

but common in Oregon, or species that appear to have primary habitats besides dunes. At least two species of moths are known only from coastal dunes in California. Powell (1976) found a new, flightless species (Areniscythris brachypteris, Scythrididae) on dunes at Oso Flaco (Powell 1976) and a new, winged species in the genus Lithariapteryx (Heliodinidae) on other dunes (Powell 1991). The globose dune beetle (Coelus globosus, Tenebrionidae) occurs only on foredunes from northern California to Baja California. Subspecies of insects confined to California coastal dunes include a hairy tiger beetle (Cicindela hirticollis gravida, Cicindelidae; Nagano 1980), a satyrid butterfly (Coenonympha tullia, Nymphalidae; Porter and Mattoon 1989) Belkin’s dune tabanid fly (Brennania belkini, Tabanidae), the El Segundo flower-loving fly (Rhaphiomidas terminatus terminatus, Mydidae), and the El Segundo blue butterfly (Euphilotes battoides allyni, Lycaenidae; Arnold 1983). The last is known only from the El Segundo Dunes and is federally listed as endangered. The California coastal dune beetle (Hyperaspis annexa, Coccinellidae; Gordon 1985) is confined to California but not to the coast. Two other unusual groups of animals found on Californian coastal dunes are the stiletto flies (Therevidae; Holston 2005) and the legless lizards (Anniella spp., Anniellidae; Kuhnz et al. 2005, Papenfuss and Parham 2013), but neither group appears to contain any taxa restricted to dunes.

Patterns of Vegetation The vegetation of coastal dunes in California tends to vary on three scales. First, over hundreds of kilometers from north to south, the vegetation of foredunes shifts from dominance by grasses to dominance by forbs; and the vegetation of more inland, stabilized dunes shifts from forest to shrubland. These patterns are likely tied to climate. Recent, extensive studies of dune vegetation types along the Pacific Coast from Alaska to Baja California (Peinado et al. 2007, 2011) suggest two main latitudinal transitions. Between about Cape Mendocino (40°N) and the Chetco River (42°N) in southern Oregon, a southward transition from temperate to Mediterranean vegetation is marked by the disappearance of dune forests. At about 37°N, near Santa Cruz, there is a second southward transition marked partly by a shift from herbaceous to shrubby species. This transition was also noted by Barbour et al. (1976). On a second scale, over hundreds of meters, vegetation changes from dominance by herbs near the beach to dominance by woody species inland. For example, foredunes at the mouth of the Salinas River are dominated by trailing, herbaceous perennials and subshrubs, whereas rear dunes are dominated by shrubs (Bluestone 1981). This pattern is likely tied both to levels of stress and disturbance and to succession. Using ordinations, Holton and Johnson (1979) found that dune scrub plant communities at Point Reyes correlated with distance inland and with soil organic matter, nitrogen content, and particle diameter, suggesting that drought and nutrient stress affected community composition. Over a decade, Williams and Williams (1984) observed changes in the vegetation of dunes at Morro Bay that included the replacement of some non-nitrogen-fixing species by legumes. The occurrence of mosses on the soils of some stabilized dunes (e.g., Danin et al. 1998, Cushman et al. 2010) probably depends on earlier colonization by plants that better tolerate unstabilized sand.

The restriction of certain species to the more sheltered parts of dunes is observed on other continents and seems more pronounced where climate is harsher (Doing 1985). On a third scale, plant community composition differs over meters from the crests to the hollows of individual dunes. For instance, plant species composition differs between the windward slope, the top, and the lee side of dunes at Morro Bay (Williams and Potter 1972, Williams 1974). Total numbers of individuals and species there are highest on the lee side, but some individual species are most abundant on the windward slopes or tops of dunes. These patterns are likely associated with differences in stress and disturbance. The various scales of patterns in Californian coastal dune vegetation have been reviewed by Barbour and Johnson (1988) and Pickart and Barbour (2007) and incorporated into the treatment of the plant communities of California by Sawyer et al. (2009) and into treatments of dune vegetation of the Pacific Coast and North America (Barbour et al. 1976, 1981, 1983).

Population and Evolutionary Ecology The relative geological and climatic similarity of coastal dunes in California combine with their high disturbance levels, disjunct distribution, and sharp environmental differences from adjacent habitats to create a distinctive landscape for population dynamics and microevolution. This landscape is not unlike a very long, linear archipelago of numerous, small to tiny islands. One likely consequence is that the degree of genetic differentiation among populations of a dune-restricted species will be determined largely by the ability of the species to disperse between dune systems. Accordingly, two dune species likely to have low longdistance dispersal rates, the flightless beetle Coelus ciliatus (Chatzimanolis and Caterino 2008a) and the trapdoor spider Aptostichus simus (Bond et al. 2001), show strong genetic differentiation between central and southern California populations. The dune beetle Cercyon fimbriatus shows high genetic diversity within populations yet very little effect of distance on genetic diversity between populations (Chatzimanolis and Caterino 2008b). This suggests high rates of long-distance dispersal, though the rates appear not to have been independently measured. Capacity for long-distance dispersal also seems to vary greatly among plant species on dunes. A few dune plants appear to have dispersed between California and Chile. Coast strawberry (Fragaria chiloensis) is considered native to both California and Chile but not to areas in between; the species is also considered native to Hawaii. Long-distance dispersal from California to South America followed by speciation is the likely origin of two Chilean species of Gilia (Morrell et al. 2000). On the other hand, the fitness of experimentally planted populations of a widespread beach and dune herb, beach evening primrose (Camissoniopsis cheiranthifolia), did not change across the current northern range limit, suggesting that other factors such as dispersal may limit its range (Samis and Eckert 2009). Some of the populations of this species have small flowers and predominant self-pollination (Dart and Eckert 2013), a derived characteristic associated with insular populations with little incoming dispersal. A second likely consequence of the sharp and relatively consistent environmental differences between dunes and inland habitats is selection for coastal races of inland spe-

Coastal Dunes   415

cies. For example, plants of common monkey flower (Mimulus guttatus) tend to be perennials on the coast and annuals inland, and the coastal plants tolerate salt spray better (Lowry et al. 2009). The quantitative trait loci (~genes) involved in this salt tolerance affect fitness at the coast but not inland, suggesting local adaptation to coastal conditions without a trade-off against fitness in inland habitats. Coastal and inland populations are reproductively isolated from each other by selection against immigrants and by differences in time of flowering (Lowry et al. 2008). Selection appears driven by seasonal drought inland and by salt spray on the coast, and limited gene flow might prevent the evolution of races adapted to both types of habitat (Hall et al. 2010). A different genetic mechanism, polyploidy, appears associated with spread of the native herb, yarrow (Achillea millefolium) onto dunes (Ramsey 2011). Differences between tolerance of salinity by coastal dune and freshwater lakeshore subspecies of American searocket (Cakile edentula) are consistent with selection for salt tolerance on beaches and foredunes (Boyd and Barbour 1986). California poppies (Eschscholzia californica) show genetically-based differences between native coastal and inland populations in California and similar differences between introduced coastal and inland populations in Chile (Leger and Rice 2007). Published studies on the population dynamics of dune organisms appear very few. Herbarium records indicate that the ranges of two common dune and beach plants, beach evening primrose and pink sand verbena, have been stable over the past hundred years (Samis and Eckert 2007). The density of plants within populations was not greater nearer the centers of their ranges, but beach evening primrose showed higher seed production nearer the center. In the monocarpic surf thistle, mortality on dunes in Santa Barbara County was 41% over two years, but mean lifespan appeared to be about five years and population size was fairly stable (Zedler et al. 1983). Population size can fluctuate greatly and rapidly in yellow bush lupine (Lupinus arboreus), a large shrub of coastal dune and grassland (Strong et al. 1995). Dune habitat might indirectly affect the population dynamics and structure of plants via selection for clonal growth, reproduction by means of vegetative offspring that remain attached to the parent. Connected plants within a clone can often share resources such as water, photosynthates and nutrients (e.g., Alpert and Mooney 1986), enabling a parent plant to support an offspring if it is buried or while its roots reach down to sand that remains moist. In some clonal dune plants such as coast strawberry, successful reproduction from seeds is rarely observed while rates of vegetative reproduction are often high. This can lead to a population of intermingled clones (Alpert et al. 1993). In outcrossing (versus self-pollinating) species, clonal structure can limit seed production. For example, clonal diversity within a 10 square meter area strongly predicted seed set in beach spectacle pod (Aigner 2004).

Community Ecology EFFECTS OF PL ANTS

Two salient physical features of coastal dunes are their instability and low soil nitrogen content. Two major effects of

416  Ecosystems

plants on dunes are to stabilize the soil and increase nitrogen availability. As an example of the second effect (Alpert and Mooney 1996), concentrations of total nitrogen in the soil at a depth of 0–​15 centimeters on a partly stabilized dune at Franklin Point were three to four times higher under shrubs of yellow bush lupine and coastal sagewort (Artemisia pycnocephala) than in open areas adjacent to shrubs; concentrations at a depth of 15–​45 centimeters were about two times higher under shrubs than in the open. As in other systems, nitrogenfixing plants such as species in the Fabaceae, the pea family, can have relatively large effects. Concentrations of inorganic nitrogen (most of the nitrogen available for uptake by plants) and net rates of mineralization of nitrogen to inorganic form were approximately two to three times higher under yellow bush lupine, which fixes nitrogen, than under coastal sagewort, which does not (Alpert and Mooney 1996). Neither species of shrub affected total soil phosphorus. Effects may be smaller on older, stabilized dunes where plants have already raised overall nitrogen availability. In soil at a depth of 0–​5 centimeters on a stabilized dune at Bodega Bay (Cushman et al. 2010), concentrations of inorganic nitrogen were approximately two times higher under shrubs of chamisso bush lupine (Lupinus chamissonis) and mock heather (Ericameria ericoides) than away from shrubs. Net nitrogen mineralization rate was about four times higher under the nitrogen-fixing chamisso bush lupine and two times higher under the nonnitrogen-fixing mock heather than away from shrubs. Some dune plants strongly reduce light availability for other plants. Light availability for photosynthesis near the soil surface was 70% lower 20 centimeters inside the edge of the canopies of shrubs of yellow bush lupine than in the open, and 90% lower 20 centimeters under shrubs of coastal sagewort than in the open at Franklin Point (Alpert and Mooney 1996). Shrubs of yellow bush lupine thus created reciprocal patchiness of light and nitrogen availability, with low-light, highnitrogen patches under shrubs and high-light, low-nitrogen patches between shrubs. Plants can have multiple, opposing effects on availability of water on dunes. At Franklin Point during the summer dry season (Alpert and Mooney 1996), soil water content was approximately 30% higher under shrubs than in the open at a soil depth of 0–​15 centimeters. In contrast, water content at 15–​45 centimeter depths was about 20–​40% lower under shrubs than in the open. Shrubs likely caused this pattern by decreasing evaporation from the soil surface and taking up water in their rooting zones below the surface. After a century, soils on dunes forested with trees in Golden Gate Park in San Francisco had 40–​80 Mg ha-1 organic carbon whereas an unstabilized dune had 5 Mg ha-1 (Amundson and Tremback 1989), suggesting that forestation greatly increased waterholding capacity. However, water repellency was high on the planted dunes and absent on the unstabilized one, suggesting that availability of water in the soil after a small amount of precipitation might be higher on the unstabilized dune. McBride and Stone (1976) likewise correlated concentration of organic carbon and water-holding capacity with apparent time since stabilization by plants on dunes on the Monterey Peninsula. Environmental patchiness created by plants can in turn affect establishment and growth of other plant species. Here again, net effects are often composed of positive and negative

A

B

FIGURE 2 1.6 Coast strawberry (Fragaria chiloensis): (A) a population of clones on the North Spit of Humboldt Bay; (B) clones growing between open sand, where light availability is high but nitrogen is low; and underneath yellow bush lupine (Lupinus arboreus), which creates complementary, low-light and high-nitrogen patches at Franklin Point, Año Nuevo State Reserve. Photos: (A) Andrea Pickart; (B) Peter Alpert.

components. For instance, embryo dunes formed by sea sandwort (Honckenya peploides), a common species at the inland edge of beaches in Oregon, promote emergence of seedlings of American dune grass. However, seedling growth is lower on than off these dunes, possibly due to burial by accumulating sand and relatively high salinity (Gagné and Houle 2001). Net effects depend on the species or even the form of species causing the effects and the species responding to them. The prostrate but not the erect form of coyote bush (Baccharis pilularis) facilitates establishment of yellow bush lupine at Bodega Bay (Rudgers and Maron 2003). However, areas under prostrate shrubs have lower richness and abundance of other plants than areas under erect shrubs, possibly due to low light or differences in soil surface temperature or litter depth (Crutsinger et al. 2010). On a stabilized dune at Bodega Bay, density and aboveground dry mass of different herbaceous species were variously higher, lower, or no different under than away from shrubs (Cushman et al. 2010). The clonal herb, coast strawberry, can reciprocally transport photosynthates and nitrogen between connected plants through stolons (Friedman and Alpert 1991). Plants appear able to mine the low-light, high-nutrient patches created by yellow bush lupine by first sending photosynthates along stolons from plants in the open to connected plants under shrubs. These shaded plants can then grow, take up nitrogen, and transport it to plants in the open (Figure 21.6). Resourcesharing capacity was higher in clones from dunes than in clones from grasslands, suggesting selection for sharing on dunes in response to high resource patchiness (Alpert 1999, Roiloa et al. 2007). Interactions between plants and the physical environment on dunes generate a diversity of habitats for animals. The nine species of stiletto flies at the Nipomo Dunes occur in three plant assemblages—​one each on active dunes, in dunes stabilized by shrubs, and in shrubby wetland areas (Holston 2005). Nesting of a native leafcutter bee is associated with certain plants on some dunes in northern California and appears to depend upon persistence of intermediate successional states of vegetation (Gordon 2000). At Point Reyes, shading by vegetation promotes abundance of the burrows used for reproduction by two species of tiger beetle (Cicindela; Cornelisse and Hafernik 2009).

The two different species are associated with different grain sizes and ranges of soil moisture, pH, or salinity. Dense stands of European beach grass or dune rush (Juncus lescurii) on dunes at Point Reyes provide nesting sites for the deer mouse (Peromyscus maniculatus); at one site the mouse concentrated its activities in these stands even though abundance of its food plants was relatively low there (Pitts and Barbour 1979). In contrast, European beach grass appears to shelter the introduced purple veldtgrass (Ehrharta calycina) from a different herbivore, the blacktailed jackrabbit (Lepus californicus; Cushman et al. 2011).

INTER ACTIONS BET WEEN PL ANTS AND ANIMALS AND MICROBE S

Herbivory can substantially reduce plant growth and reproduction on dunes. Rodents consumed approximately 65–​85% of yellow bush lupine seeds placed on dunes at Bodega Bay, compared to 5–​55% in adjacent grassland (Maron and Simms 1997). The seed bank of the species on dunes was only about 3% as large as in grassland. Excluding rodents from dunes increased emergence and establishment of adults from sown seeds several-fold over three years (Maron and Simms 2001). Predation on seeds after dispersal had no effect on seedling recruitment of cobweb thistle (Cirsium occidentale) on dunes at Bodega Bay, but herbivory on flowers by specialized insects reduced seed production and recruitment by more than 50% (Maron et al. 2002). At a different site, insect damage to seeds of cobweb thistle was low and fungi instead reduced seed production by around 30% in one of two years (Palmisano and Fox 1997), while rabbits reduced growth and delayed the reproduction of established plants. Zedler et al. (1983) estimated that insects destroy approximately 25% of the achenes (fruits) of surf thistle on sand dunes at two sites in Santa Barbara County. Herbivory can modify effects of plants on resource availability and on other dune organisms. Exclosures at Bodega Bay showed that black-tailed deer (Odocoileus hemionus columbianus) decreased seedling growth and adult seed production of chamisso bush lupine. Deer sometimes increased net mineralization of nitrogen or decreased inorganic nitrogen concentrations under the shrub, presumably through effects on

Coastal Dunes   417

the concentration of nitrogen in litter (Warner and Cushman 2002, McNeil and Cushman 2005). These exclosures also showed that presence of black-tailed jackrabbit decreased flowering of grass and abundances of forbs, shrubs, and the four most visible invertebrate herbivores: two snails, a grasshopper, and the tiger moth caterpillar (Huntzinger et al. 2008). The effect on snails appeared due to the decrease in shade that accompanied herbivory by the jackrabbits (Huntzinger et al. 2011). Some herbivores can actually decrease total herbivory. Branches of the willow Salix hookeriana containing the aphid-tending ant Formica obscuripes had more arthropods per leaf shelter but less herbivory than branches without ants (Crutsinger and Sanders 2005). There appear to be almost no published studies of pollination on coastal dunes in California. Vilà et al. (1998) reported a number of generalist pollinators on highway iceplant (Carpobrotus edulis). Removal of highway iceplant and sea rocket flowers did not affect seed set of beach spectacle pod, providing no evidence for competition among these plants for pollinators (Aigner 2004). However, seed set of beach spectacle pod was very low at a site with almost no pollinators, suggesting that pollination can sometimes limit reproduction of dune plants. Published research on soil microbes on Californian coastal dunes also appears scarce. Studies from dunes elsewhere indicate that microbe abundance increases as plants colonize and stabilize dunes. Near Ensenada in Baja California, Siguenza et al. (1996) found less than 1% colonization by mycorrhizal fungi of red sand verbena (Abronia maritima) in mobile dunes but up to 80% colonization of six species in fixed dunes. Analysis of phospholipid fatty acids in soil along transects from the shoreline to inland shrubland in Australia suggested that total microbial biomass and proportional abundance of fungi were highest in the shrubland (Yoshitake and Nakatsubo 2008). Near the sea, microbial biomass increased with soil organic matter and was not correlated with soil salinity. Some dune plants on Cape Cod in Massachusetts depend on mycorrhizae to survive (Gemma and Koske 1997). Even though unvegetated dune sites there appear to lack spores of mycorrhizal fungi, uninoculated plantings of native species can eventually be infected, presumably by dispersal of spores from outside a site.

Ecosystem Services Coastal sand dunes around the world provide a variety of products and functions that contribute to human well-being and can in theory be sustainably used (Barbier et al. 2011). The main products, or social and economic goods, that humans extract from coastal dunes are probably sand and its component minerals such as silica and feldspar, followed by fodder for livestock and then by wild foods such as fruits and mushrooms. The two main functions, or social and economic services, of coastal dunes are likely protection of property and structures from damage by waves and soil erosion; and recreation in the forms of walking, sunbathing, off-road driving, playing, swimming or wading in dune ponds, photography, horseback riding, solitude, and viewing of animals and plants. In addition, dunes catch and purify water for human use, maintain biodiversity that underlies various goods and services, provide opportunities for education and research, and sequester carbon. The quantitative values to humans of most of these goods and services of dunes have been little studied, but estimates of people’s willingness to pay for ero418  Ecosystems

sion control and recreation suggest that they are substantial ecosystem services (Barbier et al. 2011). Despite the apparent unavailability of published studies of the values of the ecosystem services of Californian coastal dunes, one can reasonably hypothesize about their relative importance. Recreation, both for day use and for overnight camping, is very likely the most highly valued service. Many dunes are easily accessible from cities, and the mild climate of the immediate coast makes recreation attractive yearround and offers escape from inland summer heat. Mining of sand and grazing on dunes are probably now very limited in California. Dunes likely protect estuaries, harbors, and some built-up areas near river mouths from storms. However, cities have largely eliminated the dunes that might protect them.

Effects of Humans Humans have had major effects on the ecology and probably on the ecosystem services provided by nearly all of California’s coastal dunes. Major causes have been conversion to residential, industrial, and other land uses; and intentional introduction of plant species to stabilize the sand. Land use has largely eliminated dunes near major cities including San Francisco, Los Angeles, and San Diego; about one-third of the city of San Francisco sits atop a Holocene dune system that was once one of the largest in the state. Effects of introduced species are reviewed below. Other effects of humans on dunes include vegetation loss and destabilization by foot traffic and off-road vehicles. Disturbance by recreational use can have measurable effects on plants and animal communities and is locally important. For example, 5–​2 0% of variation in the community compositions of arthropods on the El Segundo Dunes was explained by level of disturbance (Mattoni et al. 2000). Abundance of the burrowing legless lizard Anniella pulchra was lower in more disturbed soils on central coast dunes (Kuhnz et al. 2005). However, total plant cover did not differ between sites with unrestricted and restricted foot traffic in a study of seventeen sites in California (Tobias 2013), and restriction of visitor use for two years had only slight effects on the abundance and diversity of native plants and animals at Fort Funston in San Francisco (Russell et al. 2009). Driving on dunes can clearly denude them. Effects of driving on beaches are less clear; at Assateague National Seashore in Virginia and Padre Inland in Texas, beach driving did not appear to cause net loss of sediment from the beach and dune system but did reduce the size of dunes (Houser et al. 2013). Beach grooming can inhibit the formation of dunes in southern California even though it increases transport of sand by wind (Dugan and Hubbard 2010; see also Chapter 20, “Sandy Beaches”). Livestock grazing can have strong effects on dunes (Moore 2002). The introduction of sheep to San Miguel Island contributed to dune destabilization; dunes there have begun to restabilize following the removal of sheep and other introduced animals (Erlandson et al. 2005; see also Chapter 34, “Managed Island Ecosystems”). Published studies of grazing on dunes along the mainland California coast appear lacking, but studies from northern Europe suggest that livestock can have mixed effects on dunes in cool, moist climates like those of northern California. Autocorrelation of trampling by cattle and humans on Flemish dunes has led to local extinc-

tion of specialized arthropods (Bonte and Maes 2008). Reintroduction of livestock to dunes in North Wales after decades of exclusion increased plant diversity on dune ridges but not in slacks (Plassmann et al. 2010).

A

Introduced and Invasive Species Introduced species are those brought into a new region intentionally or incidentally by humans; invasive species are those that have spread into a new habitat and that have had negative effects on species already there (see Chapter 13, “Biological Invasions”). Gaertner et al. (2009) concluded that introduced plant species are generally associated with lower numbers of native species on dunes in Mediterranean-type climates. In California, a few intentionally introduced, perennial plants have become highly invasive on coastal dunes. Foremost are European beach grass, introduced from Europe to stabilize dunes; and highway iceplant, a succulent introduced from southern Africa as a horticultural plant and to stabilize dunes and roadsides (Figure 21.7). These introduced, invasive species contrast with a number of introduced plant species that have become widespread on dunes in California but appear so far to have had no substantial effects on existing species or ecological processes. Examples include the succulent, perennial American sea rocket (Cakile edentula). European beach grass was first introduced to the Pacific Coast of the U.S. in 1868 to stabilize dunes in San Francisco for the future Golden Gate Park (Lamb 1898 as cited in Pickart and Barbour 2007). It has since occupied nearly all the foredunes and much of the rest of the dunes from central California to the mouth of the Columbia River. On the North Spit of Humboldt Bay, cover of European beach grass increased more than fivefold over fifty years ending in 1989; spread on inland dunes was mainly due to plantings, whereas spread on foredunes was less dependent on planting (Buell et al. 1995), perhaps partly due to the ability of rhizome fragments to survive for several days in saltwater (Baye 1990 as cited in Pickart and Barbour 2007) and thus disperse via the ocean. Although not widespread in southern California so far, European beach grass can survive along the entire Californian coast and has become abundant at some southern sites (Pickart and Barbour 2007). Invasiveness of European beach grass on California dunes appears linked to having a growth form unlike that of native species and possibly to having a relatively high tolerance of stress. No native plant species forms such dense stands on foredunes. Rhizomes of European beach grass show more vertical and less lateral growth than rhizomes of the common native large grass of foredunes, American beach grass, especially when nitrogen availability is low, as well as greater survivorship and capacity for vegetative reproduction (Pavlik 1983). In one field study European beach grass maintained higher midday water potential, turgor pressure, and stomatal conductance than American beach grass despite showing less seasonal adjustment of solute accumulation and osmotic pressure (Pavlik 1985). Invasiveness of European beach grass does not appear to depend on interactions with soil microbes (Kowalchuk et al. 2002, Beckstead and Parker 2003). The direct negative effects of European beach grass on other dune plants are likely mediated largely by competition for light. One indirect negative effect is mediated by a positive effect of European beach grass on the deer mouse (Peromyscus maniculatus); the mouse nests and shelters in stands of Euro-

B

C

FIGURE 2 1.7 Three of the most highly invasive introduced species on coastal dunes in California. Photos: (A) courtesy of Lorraine Parsons, National Park Service; (B–​C) Andrea Pickart. A European beach grass (Ammophila arenaria) near Abbotts Lagoon. B Highway iceplant (Carpobrotus edulis) at Samoa Dunes National

Recreation Area. C Purple veldtgrass (Ehrharta calycina) at Guadalupe-Nipomo Dunes.

pean beach grass and preys on the seeds of other plants such as European sea rocket (Boyd 1988) and Tidestrom’s lupine (Dangremond et al. 2010). The latter is an endangered dune endemic at increased risk of extinction because of apparent competition with the invader. European beach grass directly affects the federally threatened western snowy plover by reducing the extent of its nesting habitat at the backs of beaches (Zarnetske et al. 2010). European beach grass has negative effects on various dune invertebrates, either due to Coastal Dunes   419

direct effects or by displacing native plants. Dunes dominated by European beach grass lack stiletto flies at Nipomo Dunes (Holston 2005). Whereas diversity of dune arthropods increases with cover of native plant species at sites from Sonoma to San Luis Obispo Counties, it declines with cover of European beach grass (Slobodchikoff and Doyen 1977). Invasiveness of highway iceplant likewise appears linked to its ability to form very dense stands, high tolerance of stress, and effective dispersal. Comparison of highway iceplant with a dune native in Europe suggests that drought tolerance of seedlings, dispersal by rabbits, and allelopathy all increase the ability of highway iceplant to spread on dunes (Novoa et al. 2012). Highway iceplant spreads much more than another introduced species in the same genus, Chilean sea fig (Carpobrotus chiloensis); dispersal of the former’s seeds by native mammals that eat the fruits could be partly responsible. Proportions of the two species’ seeds in scat of mule deer (Odocoileus hemionus) and black-tailed jackrabbit suggest preferential consumption of highway iceplant fruits, and passage through animals increased germination of seeds of highway iceplant but decreased germination in Chilean sea fig (Vilá and D’Antonio 1998). Dispersal of highway iceplant seeds by deer and rabbits likely aids its spread (D’Antonio 1990) even though herbivory limits its establishment (D’Antonio 1993). Hybridization can sometimes promote invasiveness; data from chloroplast DNA, allozymes, and plant morphology all indicate that hybridization results in gene flow from Chilean sea fig to highway iceplant (Schierenbeck et al. 2005). In southern Europe, plant-soil interactions inhibit initial establishment of highway iceplant (de la Pena et al. 2011); however, once established, the species alters soils so as to be more favorable to its own growth and less favorable to at least some native plants. Even after removal, highway iceplant can inhibit colonization by native plants through persistent effects on soil chemistry (Conser and Connor 2009). Over a decade, Williams and Williams (1984) observed changes in dune vegetation at Morro Bay that included displacement of natives by spread of highway iceplant or Chilean sea fig. Competition with highway iceplant for water decreases water potential and shoot growth in native mock heather and leads it to root more deeply (D’Antonio and Mahall 1991). Molinari et al. (2007) review the multiple, strong effects of highway iceplant on dunes. Two other intentionally introduced, highly invasive perennials on dunes along parts of the coast are purple veldtgrass (Figure 21.7) and yellow bush lupine. Purple veldtgrass was introduced from southern Africa to control erosion and provide forage. The grass is said to be very abundant on dunes at Vandenberg Air Force Base and could prove a major invasive species on dunes in southern California. No work on its invasion ecology in California seems to have been published yet. Yellow bush lupine is a nitrogen-fixing shrub native to coastal dunes in central California. It was intentionally introduced to northern California to increase soil fertility and stabilize dunes and has become locally abundant (Pickart et al. 1998b). Dunes are also subject to invasion by the introduced, European annual grasses that have spread widely in California’s coastal grasslands (see Chapter 23, “Grasslands”). On a stabilized dune at Bodega Bay, the most abundant introduced species was the invasive annual grass known as ripgut brome (Bromus diandrus; Lortie and Cushman 2007). Abundance of these grasses is positively associated with nitrogen availability, and introduction of yellow bush lupine to Humboldt Bay has facilitated the spread of introduced annual grasses on dunes there (Pickart et al. 1998b) . 420  Ecosystems

Management Strategies Strategies for managing coastal dunes worldwide include (Doody 2013):

. regulation of development, such as housing, that obliterates dunes

. decisions about which areas to maintain in a semi. . . . .

natural state for agricultural uses such as forestry and grazing encouragement or restriction of recreational uses of natural areas compensation for erosion through addition of sand stopping sand from moving onto adjacent areas such as roads and harbors restoring dunes that have been devegetated by human use or covered by the spread of introduced species maintaining populations of rare or endangered species

All of these are of some importance in California except perhaps allocation of lands to forestry and grazing, which seem little practiced on coastal dunes in the state. Since two main effects of humans on coastal dunes in California have been development and plant introductions, two logical main areas of management are protection of remaining dunes from development and removal of introduced plants. A key law governing development on coastal dunes in California is the California Coastal Act of 1976 (Division 20 of the California Public Resources Code, www.coastal.ca.gov), established pursuant to the Federal Coastal Zone Management Act of 1972 (Title 16 U.S.C. 1451-1454). The California Coastal Act designates a coastal zone whose terrestrial portion generally extends 1,000 yards (914 meters) inland from the mean high tide line. The extent is less in developed urban areas and more, up to five miles (8 kilometers) or to the first major ridgeline, in “significant coastal estuarine, habitat, and recreational areas.” The act thus applies to nearly all of the state’s remaining dunes. The act declares that “it is necessary to protect the ecological balance of the coastal zone and prevent its deterioration and destruction.” Furthermore, “existing developed uses, and future developments that are carefully planned and developed consistent with the policies of this division, are essential.” Among the basic goals of the act are to ensure public access to the coast, to maintain prime agricultural land in production, and to protect “environmentally sensitive habitat” and “scenic and visual qualities.” The act creates a California Coastal Commission to implement it as the designated state planning and management agency for the coastal zone. The commission exercises this charge in large part by evaluating, and either granting or denying permission for, any activity in the coastal zone covered by the act. The act provides in a general way for the use of science to inform management, in part by instructing the commission to consult with social, physical and natural scientists on a wide variety of coastal issues, especially “coastal erosion and geology, marine biodiversity, wetland restoration, the question of sea level rise, desalination plants, and the cumulative impact of coastal zone developments.” The act does not refer specifically to dunes, but one of its stated aims is to reduce the risk of beach erosion, and the commission has included dunes in areas designated as environmentally sensitive habitat. Whether as a result of the act or by elimination through development of the dunes outside preserves, the relatively

intact portions of major dune systems along the California coast now mostly lie within federal, state, or local preserves of some kind. This moots purely commercial development and emphasizes the issue of how to balance recreational use and ecological conservation. In some cases, such as the use of dunes by animals of special conservation concern, human use and ecological conservation may be mutually exclusive. For instance, Lafferty (2001) concluded based on work near Santa Barbara that protection of nesting by western snowy plovers requires that all access be prohibited to stretches of beach where birds are nesting and that dogs in adjacent stretches be on leash during the nesting season. These recommendations are being followed at a number of locations including the Coal Oil Point Reserve of the University of California at Santa Barbara, where populations of plovers have increased (Snowy Plover Program 2014). In other cases, management can help reconcile use and conservation. Addition of sand to beaches, also known as beach nourishment or replenishment, is common around the world and in some places in California (Clayton 1991, Speybroeck et al. 2006). Nordstrom et al. (2011) suggest promoting natural features on heavily used shores by adapting techniques such as beach nourishment to more closely mimic natural landforms and by restricting some activities such as driving and raking. Removing wrack from the seaward slope of the foredune can reduce scouring of the foredune and increase transport of sand inland (Jackson and Nordstrom 2013). Fencing at the seaward base of the foredune can also reduce scouring but tends to decrease transport of sand. In a survey of beaches and dunes on three continents, McLachlan et al. (2013) suggested using one index of conservation value and another of recreation potential to guide management for conservation, recreation, or both. While some of the net effects of management can probably be judged by informal inspection, others—​such as effects on populations of endangered species—​ are better measured with quantitative monitoring (Pickart et al. 2000).

Restoration Like some other habitats where frequent natural disturbance maintains early successional stages with minimal soil development, the unstabilized and partly stabilized portions of coastal dunes appear relatively amenable to restoration of native vegetation. Essential steps appear to include curtailment of driving or other human uses that remove plants, and restoration of a natural regime of sand movement. If native plants remain at a site, they sometimes spread by themselves. If not or to speed restoration, native plants can be grown in a local nursery and transplanted (Trent et al. 1983). Since areas without plants may lack mycorrhizal fungi, transplants already infected with fungi may grow better than uninfected transplants (Gemma and Koske 1997). On dunes along California’s northern and central coasts, restoring the physical environment for native vegetation now generally begins with removal of introduced plants (Pickart and Sawyer 1998). These are most often European beach grass and highway iceplant. One notably successful set of restorations of dunes covered by European beach grass has taken place at Humboldt Bay using only manual pulling and digging (Pickart 2013). For instance, one project used intensive, paid labor in each of three years followed by annual maintenance mainly by volunteers; topography and vegetation

twenty-five years later were similar to conditions on nearby, intact dunes. These projects also provide a nice example of cooperation between agencies and nongovernmental organizations (Deblinger and Jenkins 1991). Alternative methods of removing European beach grass and highway iceplant are to use heavy equipment to uproot and bury plants or to spray them with herbicides such as glyphosate or imazapyr. An ongoing series of projects near Abbotts Lagoon in Point Reyes National Seashore has yielded promising results with either mechanical removal or herbicides, sometimes abetted by hand pulling, removal of dead plants, or transplanting natives (Abbotts Lagoon Coastal Dune Restoration Project 2014). Mechanical removal of European beach grass in this work was relatively expensive and required digging up to 3 meters deep. Mechanical burial has also been used to eradicate rugosa rose (Rosa rugosa) from dunes in Denmark (Kollmann et al. 2011). Removal of introduced plants from dunes can help restore communities of native animals as well as of native plants. At Fort Funston in the Golden Gate National Recreation Area, active dune restoration has increased the cover and number of species of native plants and the abundance and diversity of native animals (Russell et al. 2009). Removal of European beach grass in Oregon and Washington has increased abundance of both the western snowy plover and native dune plants (Zarnetske et al. 2010). Sampling at six sites in northern and central California showed higher abundance and diversity of arthropods on foredunes from which European beach grass had been removed than on foredunes dominated by the grass (Doudna and Connor 2012). Restoration of dunes occupied by introduced shrubs of yellow bush lupine at Humboldt Bay requires not just removing the shrubs but also reversing their effects on the soil and other plants (Pickart et al. 1998a). Yellow bush lupine increases nitrogen availability and amount of plant litter and favors the spread of introduced, annual grasses. Pickart et al. (1998b) were finally able to remove and decrease reestablishment of this suite of introduced plants by clearing yellow bush lupine with a brush rake, removing litter and duff with a plow blade, and laying down weed mats for two years. Research in Italy indicates that the numbers of introduced and of native plant species on dunes can be controlled by different natural and human factors (Carboni et al. 2010), which might suggest further strategies for restoration.

Future Scenarios The remaining coastal dunes in California are largely protected from development, since they lie mostly within reserves, recreational areas, or military bases. Their most immediate threat is probably the continued spread of introduced invasive plants. Even though the most far-spreading invasives are no longer intentionally planted within reserves and are widely agreed to be undesirable on natural dunes, eradicating these species is difficult and controlling them requires continuing effort. A key conservation measure would be to focus this effort on the roughly ten areas in California where native communities of dune plants and animals are well developed and fairly intact. In the longer term, loss of coastal dunes to climate change appears likely. The National Research Council (2012) has projected a rise in sea level of about 0.9 meter by 2100 along the coast of California south of Cape Mendocino. Coastal Dunes   421

CO AS TA L

CALIFORNIA COASTAL COMMISSION

EM YST S NE U Regulation, D

Regulation

development

PUBLIC LAND MANAGERS

HUMAN VISITORS

Disturbance

Recreation

Facilitation, competition

OCEAN

BEACH

Aerosol deposition Sand and propagule transport

PLANTS

PRIVATE LAND Construction sites, burial, protection HOLDERS

Restoration, introduction of invasive species

Land conversion, introduction of invasive species

Pollination, dispersal, consumption

Water and nutrient uptake, burial and unburial

Food, habitat creation Litterfall, stabilization, sand trapping

SOILS

ANIMALS

Habitat creation

Food supply to nonresident animals Consumption by nonresident animals, propagule transport

COASTAL FOREST, GRASSLAND, SHRUBLAND, WETLAND

Key

Sand transport

Font type

Example

AGENT

Bold, all caps

ANIMALS

Current process

Regular

Disturbance

Former process

Italics

Land conversion

FIGURE 2 1.8 Key components of coastal dune systems in California.

­Revell et al. (2011) used a downscaled regional global climate model, statewide data on coastal geology, and two methods of estimating flood elevations to predict erosion distances of 170 to 600 meters on coasts backed by dunes in California given a 1.4 meter rise in sea level. Although effects of climate change on the frequency or intensity of storms along the California coast are uncertain (National Research Council 2012), wave height has increased along the Pacific Coast of the U.S. over the past twenty-four years, consistent with increases in the North Pacific and Multivariate ENSO Indices over the past century that may reflect climate change (Seymour 2011; see Chapter 2, “Climate”). Increases in sea level and wave height could interact with effects of introduced species on dune morphology to flood large areas of dunes with saltwater. A model of dune overtopping as a function of storm intensity and sea level rise showed that the reduction in foredune heights associated with replacement of European beach grass by American beach grass in Oregon could triple the area subject to flooding by saltwater washing over the foredune (Seabloom et al. 2013). A model of succession on dunes in Texas suggests that sea level rise would particularly lead to losses of late-successional plants (Feagin et al. 2005). Because beaches supply the sand for dunes, changes in beach erosion will also influence the future of dunes. Hapke et al. (2009) showed a recent increase in net beach erosion, finding that only 40% of beaches along approximately 750 kilometers of the California coast had undergone net erosion over the past 120 years, but that 66% had shrunk over the past 25 years. This could be caused by an increase in gross erosion associated with climate change or by a decrease in gross deposition caused by engineering. Slagel and Griggs (2008) estimated that dams have reduced the transport of sand down 422  Ecosystems

rivers to the ocean by 5%, 31%, and 50% respectively in northern, central, and southern California. Human construction can also erect barriers to longshore transport of sand (see Chapter 20, “Sandy Beaches”). For example, erosion of Ocean Beach in San Francisco is associated with both reduced sediment supply from San Francisco Bay and an exposed sewage outflow pipe (Barnard et al. 2012). Other potential effects of climate change are more speculative. Higher temperatures throughout the state might increase visitation to beaches but shift use to activities with relatively low impact such as bathing (Coombes and Jones 2010). Although no clear expectations have emerged about effects of climate change on ocean currents, changes to the southerly, longshore flow that characterizes the northern and central California coast could reduce fog and increase temperatures. Where space exists on their inland edge, dune systems might respond to higher seas by migrating inland rather than disappearing. Psuty and Silveira (2010) offer a conceptual model of accretion and transport of sand based on observations of dunes on Fire Island in New York that suggests that sea level rise can sometimes lead to more positive sediment budgets and the displacement of dune systems inland without net loss of the system. However, development on the inland side of dunes could cause a “coastal squeeze” (Schlacher et al. 2007), with dunes caught between a rising, stormier sea and homes or businesses. To adapt management of dunes to climate change in California, one could thus avoid development where shorelines are retreating (Defeo et al. 2009). One could also counter erosion with appropriate beach nourishment (Brown and McLachlan 2002) and by avoiding the reduction of inputs of sand to beaches, such as by dams along rivers (Slagel and Griggs

2008). The coastal dunes of California are a geologically ephemeral fringe. However, they embody natural dynamics that cover most of the state, from erosion in the Sierra Nevada and coastal mountains to currents and storms in the ocean; and social dynamics that involve most of the state’s people, through coastal development, engineering of rivers, recreation, and concern for other species. For both research and management, the coastal dune ecosystem should be viewed as comprising open, dynamic, coupled natural and human systems (Figure 21.8).

Summary Coastal sand dunes are recycled land—​washed down from inland mountains in rivers, carried alongshore by currents, washed ashore by waves, and blown in from beaches to form a moving landscape of ridges and hollows. Dunes have distinctive plants and invertebrates, some rare or confined to the state. Plants that colonize bare dunes stabilize them, at which point other plants appear and the main factors controlling vegetation shift from movement of sand to availability of resources such as water, nitrogen, and light. Coastal dunes offer the people of California recreation, scenery, and attractive sites for construction. However, the movement of sand dunes can also damage property and infrastructure. People have responded by protecting some dunes, building over many of the dunes close to cities, and stabilizing dunes with introduced plants. Dunes in northern California now are largely covered with introduced plants; dunes in southern California have been mostly built over; and most of the state’s relatively intact dunes lie within public reserves. Further development of coastal dunes in most of the state is regulated by the California Coastal Commission, created by the California Coastal Act of 1976. Management of remaining dunes requires a balance of human use and conservation, restoration of dunes laid bare by use or covered with introduced plants, and control of introduced plants where communities of native plants and animals persist. Restorations have been notably successful at several sites in northern California. In the long term, the future of dunes is linked to climate change. Rising sea level is expected to displace dunes inland, likely causing net loss.

Acknowledgments I thank the California Department of Parks and Recreation and the U.S. Park Service, particularly the staffs of Point Reyes National Seashore and Año Nuevo State Reserve as well as the University of California Bodega Marine Laboratory and Reserve for facilitating the research that gave me familiarity with the coastal dunes of California. Peter Baye, Carla D’Antonio, Lorraine Parsons, Andrea Pickart, and Jackie Sones kindly provided photographs; and Carla D’Antonio contributed helpful comments on an earlier version of the chapter. I am also very grateful to Hal Mooney and Erika Zavaleta for their editorial leadership, guidance, and patience.

Recommended Reading California Coastal Commission. 2003. California coastal access guide. University of California Press, Berkeley, California. Doody, J. P. 2013. Sand dune conservation, management, and restoration. Springer Verlag, New York, New York.

Griggs, G., K. Patsch, and L. Savoy, editors. 2005. Living with the changing California coast. University of California Press, Berkeley, California. Pickart, A. J., and M. G. Barbour. 2007. Beach and dune. Pages 155–​ 173 in M. G. Barbour, T. Keeler-Wolf, and A. A. Schoenherr, editors. Terrestrial vegetation of California. Third edition. University of California Press, Berkeley, California.

Glossary Allozymes  Enzymes that occur in organisms in variant or alternate forms as a result of coding by different alleles (versions or alternative forms of a gene). Apparent competition  A mutual though often asymmetrical negative effect of two species on each other caused, not by direct interaction or consumption of a common resource, but instead by an increase in a predator or parasite. Blowout  A portion of a dune ridge hollowed out by wind. Deflation plain  An area upwind of a dune that wind has eroded down to the water table. Derived  Of a trait of an organism, present in some taxa but not in their common ancestor. Dune sheet  A set of dune ridges that form a continuous area of sand. Forbs  Plants that are herbaceous and not grasses or like grasses; includes plants commonly called wildflowers. Foredune  The most seaward dune of a dune system, adjacent to and paralleling the beach. Introduced  Of a species, brought into a new region by human action. Invasive  Of a species, entering into a habitat where it was not previously present and having negative effects on species already there. Mineralization  Conversion from organic to inorganic form, as of nitrogen from amino acids to ammonium or nitrate, the main forms in which nitrogen is taken up by plants. Monocarpic  Refers to plants that reproduce (fruit) only once and then die, without necessarily having a short lifespan. Ordination  Any of various statistical methods for arranging items in an abstract space such that more similar items are nearer each other. Parabolic dune  A curved dune ridge whose ends point upwind. Polyploidy  The condition of containing more than two paired sets of chromosomes. When polyploidy occurs in plants, it can yield a new species if the polyploid individuals cannot reproduce with their diploid ancestors and counterparts. Quantitative trait loci  The regions of DNA that are associated with the genes for a particular trait. Slack  A seasonal wetland in a dune system. Stolon  A horizontal stem that grows at or just below the soil surface and that can form new shoots and roots at stem nodes; distinguished from rhizomes mainly by the location of rhizomes well below the soil surface. Stomata  A pore in the epidermis of a plant that can be opened and closed, controlling movement of gasses in and out of the plant. Succession  A progressive change in species composition in a place over time. Transverse dune  A dune ridge oriented perpendicular to prevailing winds. Coastal Dunes   423

Water potential  A measure of the free energy of water; water tends to moves from where water potential is higher to where it is lower, so the water potential in the roots of a plant must be lower than in the adjacent soil for the plant to take up water.

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T WENT Y-T WO

Coastal Sage Scrub ELSA E . CLEL AND, JENN IFER L . FUN K , and ED ITH B. ALLEN

Introduction From a distance, coastal sage scrub might not appear the most compelling of California ecosystems; short in stature, it lacks the majesty of a redwood forest—​but up close, coastal sage scrub is a feast for the senses. This ecosystem is known for its fragrance; it is often (though not necessarily) dominated by shrubs in the genera Artemisia and Salvia, and visitors are frequently overwhelmed by the heady scent of sage in the air. Coastal sage scrub is also renowned for both its beauty—​t he texture and colors when in bloom are reminiscent of a coral reef—​as well as its spectacular diversity. Coastal sage scrub includes some two hundred species of forbs throughout its range (Skinner and Pavlik 1994) as well as a diverse array of associated animal species of conservation concern (Diffendorfer et al. 2007). The Mediterranean climate region of California is considered one of the global biodiversity hotspots (Myers et al. 2000), and coastal sage scrub with its interspersed forblands and riparian areas has a greater concentration of rare species than any other California ecosystem type. This has made coastal sage scrub one of the most endangered ecosystems in the U.S. (Rubinoff 2001); hence, restoration and conservation are important aspects of this chapter. Coastal sage scrub is found in coastal zones between San Francisco and Baja California and, despite its name, extends

up to 100 kilometers inland in some areas (Figure 22.1). Drought-deciduous shrubs tend to dominate coastal sage scrub, while evergreen shrubs are more common in adjacent chaparral communities; this distinction has led to the colloquial description of coastal sage scrub as “soft chaparral” (Mooney and Dunn 1970, Harrison et al. 1971). In the northern range coastal sage scrub occurs at low elevations often intermixed with grassland, chaparral, and/or oak woodland (Kirkpatrick and Hutchinson 1980; Figure 22.2). At the southern extent of its range coastal sage scrub is found both at the coast and at higher elevations, between zones of desert scrub and higher-elevation chaparral (Mooney and Harrison 1972). The complexity of this mosaic suggests that disturbance, soils, and/or topographic effects play a major role in determining distributions, as opposed to purely climate.

Major Ecosystem Services Provided by This System FOOD AND FOR AGE

Historically, coastal sage scrub and interspersed forblands provided food for indigenous native Californians as docu429

mented by 1770s Spanish explorers who were invited to consume seeds of native wildflowers such as Salvia columbariae that occurred in great abundance (Minnich 2008). European settlers grazed cattle and sheep in coastal sage scrub, but dense shrublands were not prized grazing lands. Rather, interspersed forblands and the occasional perennial grasslands of southern California were the most productive grazed areas in lowlands. Coastal sage shrubs could be easily removed mechanically or by burning, and shrublands were thus converted to exotic annual grassland to improve forage (Burcham 1957, Robinson et al. 1993). Nevertheless, the short growing season and semiarid climate proved too harsh for a sustained grazing industry, and most large-scale grazing operations ended by the 1930s (Robinson et al. 1993). The recent decline of coastal sage scrub has been postulated by some researchers as failure to recover from grazing, but many sites that were mapped as annual grassland in the 1930s have been able to recover native shrub cover (Minnich and Dezzani 1998). By the time these early botanical surveys were conducted, most land that could be cultivated for agriculture had been converted. The vegetation type map (VTM) survey of the 1930s (Wieslander, http://vtm.berkeley.edu/about/) showed that most level valley bottoms in coastal sage scrub landscapes had been planted to annual grains or were grazed and converted to exotic grassland, while slopes remained occupied mainly by shrubs. Today much abandoned agricultural land is undergoing conversion to urban development (Chen et al. 2010), but other areas are being restored to coastal sage scrub for biodiversity protection (Allen et al. 2005, Marushia and Allen 2011).

URBAN RECRE ATION AND BIODIVERSIT Y SERVICE S

Because of its location in coastal areas with a desirable climate and low topographic relief (relative to chaparral, which is often found on steeper slopes), the major modern-day use of coastal sage scrub land is urban development. Some 25% of California’s coastal sage scrub has been converted to urban and suburban development (Figure 22.3). Shrubs provide important erosion control on slopes adjacent to urban development; following heavy winter rain events, slopes occupied by exotic grasses are more susceptible to erosion than slopes occupied by shrubs (Gabet and Dunne 2002). Current important ecosystem services of coastal sage scrub include “natural beauty” (Huntsinger and Oviedo 2014) and recreational opportunities for nearby urban dwellers as well as biodiversity maintenance in this highly diverse ecosystem type. Some one hundred listed and sensitive animal species occur in southern California coastal sage scrub, including more than thirty bird species, thirty-plus mammals, over twenty-five herptiles, and several invertebrates (Riverside County Transportation and Land Management Agency 2015, San Diego Multi­species Conservation Plan 2015). Conserving the lands in which these species occur requires major efforts by local and federal agencies, including land acquisition to increase the extent of reserves, land management for invasive species control, wildfire control, recreation management, and maintenance of rare species populations (Scott et al. 2006, Barrows et al. 2005).

Photo on previous page: Santa Monica Mountains National Recreation Area, off of Deer Creek Road. Photo: Jessica D. Pratt. 430  Ecosystems

Major Physical Features and Controls over Distribution The coastal zones where coastal sage scrub occurs have a Mediterranean-type climate with a winter rainy season and a prolonged summer drought. Rainfall across the range averages 250–​450 mm yr-1, and while found on a range of aspects, coastal sage scrub is most common on steep, south-facing slopes (Kirkpatrick and Hutchinson 1980). Coastal sage scrub soils tend to be relatively nutrient-rich; Westman (1981b) surveyed soils in sixty-seven coastal sage scrub sites and found total nitrogen of 0.15%, two to three times greater than measurements taken in California chaparral. Other major nutrients also tend to be relatively abundant, with 24–​2 8 mg/g extractable phosphorus and 141–​246 mg/g potassium in Venturan and Riversidean sage scrub (Westman 1981b, Padgett et al. 1999). It is unclear whether high soil nutrients result from high nutrient turnover rates by quickly decomposing deciduous litter, or whether coastal sage scrub species have high nutrient requirements and hence soil nutrient availability is an underlying predictor of their distribution. However, coastal sage scrub tends to be found more commonly on argillaceous soils with high clay content as compared with shallow siliceous soils where chaparral may be more common (Kirkpatrick and Hutchinson 1980). This suggests that inherently high cation exchange capacity could explain some of the higher soil nutrient availability in coastal sage scrub. In Mediterranean-climate regions such as this, fine-textured soils can retain moisture closer to the surface during the rainy season, promoting drought-deciduous shrubs with shallow root systems. Correspondingly, roots of coastal sage scrub shrubs are concentrated in shallow soil layers to 1.5 meters, whereas chaparral has species with the potential for roots to extend to several meters depth (Hellmers et al. 1955, Harrison et al. 1971, Wood et al. 2006). The current distribution of coastal sage scrub is limited most by urbanization and land-use change. An estimated 10–​15% of original coastal sage scrub remains intact (Westman 1981a). These figures are based on an estimated 1 million hectares covered by coastal sage scrub before EuropeanAmerican settlement (Küchler 1977), although more recent estimates suggest that coastal sage scrub covered less area (632,800 hectares in Fenn et al. 2010). The original extent of coastal sage scrub is uncertain based on Minnich’s (2008) reanalysis of vegetation descriptions from the travel diaries of mid-eighteenth-century Spanish explorers. In areas where Küchler mapped coastal sage scrub in the Los Angeles and Riverside–​Perris Plains, explorers described extensive fields of wildflowers rather than shrubs. Minnich interprets these as a vegetation type that has been little recognized in California, annual forblands, which occur throughout California in areas with low precipitation (less than 30 centimeters). Coastal sage scrub occurred on hillsides but less on valley bottoms, according to reinterpreted descriptions. Bottomlands were extensively converted to agriculture or intensive grazing before the 1930s VTM survey (Wieslander 1935), so few modern-day accounts of annual forblands exist. The hydrologic characteristics of soils and plant water relations that would limit coastal sage scrub from colonizing finer-textured, deeper soils of valley bottoms still need to be explored. Altered fire regimes, nitrogen deposition and subsequent conversion to exotic annual grassland threaten the integrity of coastal sage scrub (Talluto and Suding 2008, Keeley et al. 2005, Eliason and Allen 1997, Minnich and Dezzani 1998). The impacts of these human activi-

FIGURE 2 2.1 Distribution of coastal sage scrub in California. Data from U.S. Geological Survey, Gap Analysis Program (GAP). Map: Parker Welch, Center for Integrated Spatial Research (CISR).

A

B

C

D

California Coastal Sage Scrub by Protection Status Biodiversity protected (15.2%) Multiple use (19.8%) No protection mandated (10.7%) Unprotected (54.0%)

FIGURE 2 2.2 (Above) Coastal sage scrub is a diverse plant community dominated by short-statured shrubs but with significant variation in species composition across its range. Photos: (A) Todd Keeler-Wolf, (B) Justin Valliere, (C) E. B. Allen, (D) Elizabeth Wolkovich. From north to south, four major variants of coastal sage scrub occur: A Diablan (January 2007 in eastern Alameda

County); plants include Artemisia california and Salvia mellifera interfacing with Quercus douglasii woodland in background.

Monterey Biodiversity Protected DIABLAN(105,355, ha), Multiple Use (134, 796 ha), No Protection Mandated (72,552 ha), Unprotected (366,967 ha)

B Venturan (December 2012 at Deer Creek); plants

Sources: California Department of Forestry and Fire Protection (2007) Protected Areas Database of the United States (2012) San Luis Obispo Counties of Orange, Riverside, and San Diego (2013)

C Riversidian (Lake Skinner); plants include Keckiella

VENTURAN

RIVERSIDIAN

Los Angeles

N

0

100 Kilometers

200

DIEGAN

San Diego

include Salvia apiana, Artemisia californica, Hazardia squarosa, Opuntia littoralis, and Avena fatua, with the evergreen shrub Malosma laurina in background. antirrhinoides and Malacothamnus fasciculatus in foreground, Eschscholzia californica in open spaces between shrubs, and exotic annual grasses in areas with shrubs historically cleared for grazing; in background, upper slopes of Black Mountain covered by mixed chaparral. D Diegan (April 2007, facing east in the Sweetwater

area of San Diego National Wildlife Refuge); plants include Eriogonum fasciculatum, Artemisia californica, and Malosma laurina.

FIGURE 2 2.3 (Left) Protected area status of coastal sage scrub in California (data from U.S. Geological Survey 2013, ), with the four major variants of coastal sage scrub outlined in black (after Westman 1983).

ties on coastal sage scrub is the focus of later sections in this chapter.

Principal Organisms The shrubs that comprise the majority of coastal sage scrub community biomass include many drought-deciduous or semievergreen species (Table 22.1). These shrubs are typically less than 2 meters in height with relatively shallow root systems, soft stems, and often thin, deciduous leaves (Holland and Keil 1995). A few common evergreen species occur, which survive in dry coastal environments via deep roots. In the

understory and in open areas between shrubs is a great diversity of annual and perennial forb species as well as several common, native, perennial grass species. Several species of cactus occur in the southern areas of coastal sage scrub; the most common is Opuntia littoralis. At a local level, community composition is influenced by aspect, soil texture, and soil depth; south-facing slopes, coarser-textured soils, and shallower soils create more xeric conditions and favor drought-tolerant shrubs like Eriogonum fasciculatum, Artemisia californica, and Salvia apiana along with higher abundance of Opuntia species (DeSimone and Burk 1992). Because community composition and species dominants can change during succession following a disturbance, these factors do not consistently pre-

TABLE 2 2 .1 Common coastal sage scrub native and exotic plant species, and their responses to fire

Scientific name

Common name

Origin

Functional type

Response to fire

Amsinckia menziesii

Common fiddleneck

Native

Annual forb

Reseeder

Cryptantha intermedia

Common cryptantha

Native

Annual forb

Reseeder

Artemisia californica

California sage

Native

Deciduous shrub

Reseeder

Encelia californica

California brittlebush

Native

Deciduous shrub

Reseeder

Eriogonum fasciculatum

California buckwheat

Native

Deciduous shrub

Reseeder

Salvia apiana

White sage

Native

Deciduous shrub

Resprouter

Salvia leucophylla

Purple sage

Native

Deciduous shrub

Reseeder

Salvia mellifera

Black sage

Native

Deciduous shrub

Reseeder

Heteromeles arbutifolia

Toyon

Native

Evergreen shrub/tree

Resprouter

Malosma laurina

Laurel sumac

Native

Evergreen shrub/tree

Resprouter

Rhamnus californica

California buckthorn

Native

Evergreen shrub/tree

Resprouter

Rhus integrifolia

Lemonadeberry

Native

Evergreen shrub/tree

Resprouter

Acmispon glaber

Deer weed

Native

N-fixing subshrub

Reseeder

Eschscholzia californica

California poppy

Native

Perennial forb

Reseeder

Dichelostemma capitatum

Blue dicks

Native

Perennial forb (corm)

Resprouter

Opuntia littoralis

Coastal prickly pear

Native

Succulent

Resprouter

Marah macrocarpus

Wild cucumber

Native

Vine

Resprouter

Brassica nigra

Black mustard

Exotic

Annual forb

Generally positive

Centaurea melitensis

Maltese star-thistle

Exotic

Annual forb

Generally positive

Erodium cicutarium

Redstem filaree

Exotic

Annual forb

Generally positive

Avena fatua

Wild oat

Exotic

Annual grass

Generally positive

Bromus hordeaceus

Soft brome

Exotic

Annual grass

Generally positive

Bromus madritensus spp rubens

Red brome

Exotic

Annual grass

Generally positive

Vulpia myuros

Rat tail fescue

Exotic

Annual grass

Generally positive

Medicago polymorpha

Bur clover

Exotic

Annual n-fixing forb

Generally positive

Cynara cardunculus

Artichoke thistle

Exotic

Perennial forb

Generally positive

Hirschfeldia incana

Short-podded mustard

Exotic

Perennial forb (short-lived)

Generally positive

Source: Malanson and O’Leary 1982. Note: Reseeders may be weak resprouters if not burned to ground level; strong resprouters can also reseed.

Coastal Sage Scrub   433

FIGURE 2 2.4 Exotic annual grasses that have senesced late in the growing season on a previously burned hillside in Orange County, surrounded by mature coastal sage scrub. Exotic grass litter is highly flammable and helps spread fire, fueled by Santa Ana winds in many parts of southern California. Photo: Jennifer Funk.

dict species presence or absence within particular coastal sage scrub habitats. Four regional associations of coastal sage scrub are often described based on differences in community dominants: Diablan, Venturan, Riversidian, and Diegan (c.f. Westman 1983, building on Kirkpatrick and Hutchinson 1977, Axelrod 1978; see Figure 22.2, spatial distribution in Figure 22.3). Diablan coastal sage scrub extends from Mount Diablo in the north to just south of San Luis Obispo and is sometimes termed “northern coastal scrub” (Ford and Hayes 2007). Venturan and Diegan types include coastal areas extending from Santa Barbara to San Diego and San Diego to Baja California, respectively. Riversidian sage scrub includes inland areas from Ventura south to Baja California and is also called inland sage scrub. Within these types many authors have proposed sage scrub associations based on geographic boundaries, community composition, or dominant species (Westman 1983, Sawyer and Keeler-Wolf 1995). One type that is consistent across these variants is alluvial sage scrub that occurs wherever rivers flow through coastal sage scrub and contains some shrub and forb species not found in other coastal sage scrub variants (Sawyer and Keeler-Wolf 1995). Coastal sage scrub communities across this range exhibit high beta-diversity; while only two to five shrub species usually dominate any given locality (Minnich and Dezzani 1998), strong regional variation yields different dominant species dominating in each zone. Describing this regional variation in vegetation composition has been the focus of excellent prior reviews (e.g., Rundel 2007). Variation in community dominants might be driven partly by environmental gradients. For example, an analysis of forty-three habitat variables found summer evaporative stress—​in particular, the temperature of the warmest month as opposed to total or seasonal rainfall—​to be the best predictor of species distributions across sixty-seven coastal sage scrub sites (Westman 1981b). Species richness also varies regionally within coastal sage scrub. Richness is positively correlated with distance inland and soil pH, and negatively correlated with soil nutrient content (Keeley et al. 2005b). 434  Ecosystems

Community composition changes following fire as some shrubs and cactus resprout and a diverse mix of herbaceous annual and perennial species appear (see Table 22.1; O’Leary and Westman 1988). One of the most dominant species following fire is Acmispon glaber (previously Lotus scoparius), a keystone nitrogen-fixing shrub that replenishes nitrogen lost in the fire and provides nutrient-rich leaves for herbivores (Rundel 2007). Yucca whipplei is also common in areas characterized by frequent fire, as its fibrous main stem makes it particularly fire-resistant (Rundel 2007). Invasive species can be common in coastal sage scrub, particularly following disturbances such as fire (Figure 22.4). Most invasive species are annual grasses and forbs introduced from other Mediterranean-climate ecosystems. These species germinate early in the growing season, have shallow, fibrous root systems, and produce a thick layer of litter that persists until the next growing season (Bartolome 1979, Eliason and Allen 1997, Wainwright et al. 2012). Several animal species that occur in coastal sage scrub habitat are listed as threatened, endangered, or of conservation concern by state and federal agencies (CDFG 2013). For example, the California gnatcatcher (Polioptila californica) is a federally threatened species and requires coastal sage scrub habitat. Modeling studies show that this species is particularly threatened by climate change due to projected shifts in the future range and extent of coastal sage scrub (Preston et al. 2008). As discussed later in this chapter, the abundance and diversity of animal species is strongly linked to vegetation dynamics and their responses to fire and human disturbance.

Ecosystem Characteristics: Coastal Sage Scrub versus Chaparral Many researchers have quantified ecosystem characteristics of coastal sage scrub for the purpose of comparing them to chaparral, as a way of contrasting communities dominated by similar life forms (shrubs) but differing patterns of growth and leaf longevity (drought-deciduous versus evergreen). Primary production in coastal sage is generally lower than evergreen chaparral and varies among sites—​255 g m-2 yr-1 at one site (Gray and Schlesinger 1981), 355 g m-2 yr-1 at another (Gray 1982)—​a nd also among years with varying precipitation (e.g., from 250–​600 g m-2 yr-1) (Vourlitis et al. 2009). The lower productivity of coastal sage scrub compared with chaparral is partially explained by lower precipitation and the shorter growing season of drought-deciduous shrubs (Gray 1982). However, consistent with their location closer to the high-return end of the leaf economic spectrum (Wright et al. 2004), deciduous coastal sage scrub shrubs can produce two to three times more biomass during the peak spring growing season than co-occurring evergreen species (Gray and Schlesinger 1983). The lower production of coastal sage scrub relative to chaparral has been investigated in relation to a number of physiological differences, particularly nutrient-use efficiency (NUE) and drought tolerance. The short-lived leaves of droughtdeciduous shrubs tend to have higher N content but lower NUE than leaves of evergreen chaparral shrubs (Gray 1983). This is likely a key difference given that fertilization experiments suggest the production of coastal sage shrubs is nitrogen-limited (Padgett and Allen 1999, Vourlitis 2012). Drought tolerance and overall water relations in coastal sage scrub are strongly influenced by rooting depths. Roots of coastal sage

0.8

Native Exotic

0.6

NDVI

0.4

0.2

July 30

July 3

Mar. 20 Mar. 29 Apr. 9 Apr. 17 Apr. 27 May. 8 May. 20

Feb.18 Mar. 2

Jan. 21 Feb. 3

0 Nov. 16 Nov. 27 Dec. 6 Dec. 20

scrub species are often shallower than chaparral species (e.g., Harrison et al. 1971, Poole and Miller 1975). Poole and Miller (1975) found drought-deciduous species had shallower roots, higher rates of transpiration, and earlier onset of summer water stress than evergreen species characteristic of chaparral. While co-occurring S. mellifera (coastal sage) and Ceanothus megacarpus (chaparral) reached similarly low seasonal water potential (Gill and Mahall 1986), S. mellifera suffered greater embolism and loss of conductivity from drought stress (Kolb and Davis 1994). Similarly, Jacobsen et al. (2008) found that stem xylem was more resistant to cavitation in a suite of chaparral shrubs than in coastal sage shrubs; consequently, chaparral shrubs maintained physiological processes at lower water potentials relative to coastal sage shrubs (Jacobsen et al. 2008). The Mediterranean-type climate results in a distinct phenology of coastal sage scrub, where most growth and nutrient dynamics are constrained by the prolonged summer drought. While most shoot and leaf production occurs during the winter rainy season, A. californica and other common droughtdeciduous coastal sage scrub dominants (S. leucophylla and S. mellifera) have seasonally dimorphic leaves: following leaf drop of the larger leaves, smaller leaves are produced on auxiliary shoots that persist through the summer (Westman 1981c, Gray and Schlesinger 1981). Subsequently, A. californica flowers between April and December, with more northern populations flowering thirty to fifty days earlier than southern populations (Pratt and Mooney 2013). Soil nitrate availability in coastal sage scrub is also influenced by the timing of rainfall and subsequent plant uptake. Soils generally have low soil nitrate levels during the wet season, presumably due to plant uptake (Padgett et al. 1999, Liu and Crowley 2009), and higher levels during the summer dry season, with a peak of availability following the first winter rains (Padgett et al. 1999, Vourlitis et al. 2009) when accumulated N from mineralization and dry deposition is flushed into the soil. Similarly, organic nitrogen and carbon at three coastal sage scrub sites were lower during the wet season and higher during the dry season, reflecting soil moisture limitation of microbial activity (Liu and Crowley 2009). Biogeochemical cycles in coastal sage scrub can also be strongly influenced by invasion into these systems. For instance, Wolkovich et al. (2010) found that experimental addition of exotic annual grass litter increased carbon and nitrogen pools in coastal sage scrub by approximately 20%, due to both increased production and decreased rates of litter decomposition. The increase in production likely resulted from increased soil water availability under the exotic litter layer (Wolkovich et al. 2009b), although invasive grasses can also accelerate water loss from soils via high transpiration rates (Wood et al. 2006). The decreased decomposition rates observed with litter addition might have resulted from reduced UV photodegradation; this abiotic avenue for decomposition is important in other areas of California and declines in magnitude with increased shading under deeper litter layers (Henry et al. 2008). As already noted, the phenology of common invaders is distinct from that of native coastal sage scrub species (Cleland et al. 2013), with potentially important implications for the timing of ecosystem carbon uptake. A comparison of NDVI (normalized difference vegetation index, a measure of canopy greenness) shows a longer and less pronounced period of growth for plots dominated by native S. mellifera versus a shorter period of rapid growth for plots dominated by herbaceous exotic species (Figure 22.5; Ellen Esch unpublished data).

Date FIGURE 2 2.5 Normalized Difference Vegetation Index (NDVI), a

measure of canopy greenness, collected during the 2012–​2 013 growing season at the Santa Margarita Ecological Reserve. Data were collected over plots dominated by native shrubs (Salvia mellifera) or exotic herbaceous species (Hirschfeldia incana, Centauria melitensis, and several exotic annual grasses). Source: Ellen Esch, unpublished data.

Key Community Interactions Native and Exotic Species Interactions Disturbances such as fire, grazing, and nitrogen deposition facilitate invasion by exotic species, and once established, exotic species can outcompete native coastal sage scrub shrub species for limiting resources (Schultz et al. 1955, Davis and Mooney 1985, Eliason and Allen 1997, Cione et al. 2002, Yelenik and Levine 2010). Roots of exotic annual grasses and native shrub seedlings occupy the same soil region (Figure 22.6); thus annual grasses have the strongest competitive effects on native shrubs at the seedling stage both via water and through their effects on shading (Figure 22.7; Eliason and Allen 1997, Yelenik and Levine 2010). Competition for water can be alleviated in larger native shrubs by deeper roots; however, shallow-rooted exotic grass species can preferentially use precipitation before it reaches deeper roots of native coastal sage scrub shrubs (Eliason and Allen 1997, Wood et al. 2006). While exotic grasses compete effectively for light and water, competition for nutrients appears to be more even between native and exotic species. For instance, Padgett and Allen (1999) found that exotic grasses and native shrub seedlings displayed similar positive responses to nitrogen fertilization, and native species can outcompete annual exotic species for nitrogen under low-nutrient conditions (Zink and Allen 1998, Yoshida and Allen 2001). Some shrubs appear to compete particularly well with exotic species (Eriogonum fasciculatum, Hazardia squarrosa, Baccharis pilularis); how these species outperform exotic species is an open and interesting question (Rundel 2007). In a new twist on the invasion process, commercial cultivars of Pinus radiata (an endangered species native to the California central coast) have invaded northern coastal sage scrub and are associated with reduced species richness and cover in areas they invade (Steers et al. 2013). Coastal Sage Scrub   435

FIGURE 2 2.6 Root profiles of native shrubs (light brown) and exotic annual grasses (darker brown). Roots of establishing shrubs occupy the same shallow layers as roots of exotic annual grasses and place them in direct competition for water and nutrients. Exotic annual grasses also contribute to a significant litter layer aboveground, with multiple potential effects on establishing shrub seedlings including shading of shrub seedlings and greater soil moisture retention for deeper-rooted, mature shrubs. Illustration: Julie E. Larson.

In addition to competition, exotic grasses affect native shrubs through modification of the abiotic environment. By decreasing evaporation and consequently increasing soil moisture, invasive grass litter increased growth of the native shrub A. californica (Wolkovich et al. 2009b). Such beneficial effects of exotic grass litter might be restricted to adult shrubs, as other studies have found that exotic grass litter suppresses native shrub seedling establishment through competition for light (Eliason and Allen 1997, DeSimone and Zedler 1999; see Figure 22.7). Exotic grasses can alter abiotic factors in other ways that benefit native plant growth in coastal sage scrub habitat. Yelenik and Levine (2011) found that A. californica displayed higher growth on soils influenced by the exotic grass Avena barbata. However, A. californica failed to colonize grassland during the four-year study period, leading the authors to conclude that processes other than plantsoil feedbacks, such as competition, were more important in native species recovery. The mechanism for the positive plant-soil feedback observed in this study was not investigated and could involve effects of A. barbata on soil nutrients, structure, or biota.

herbaceous species under shrubs (Bartholomew 1970). Seed predation has a stronger impact on native seedling recruitment than seedling herbivory, although this varies among shrub species because species with large or multiple-seeded propagules are more attractive to granivores (DeSimone and Zedler 1999). Changes in plant community composition can influence shrub- and ground-dwelling arthropods in coastal sage scrub habitat. Higher productivity in invaded coastal sage scrub increased the abundances of shrub-dwelling arthropod herbivores and predators associated with the grazing food web but reduced the abundance of ground-dwelling arthropods associated with the detrital food web (Wolkovich et al. 2009a; Wolkovich 2010). The positive effect of invasion on shrubdwelling arthropods was linked to overall increases in shrub biomass, while the decline in ground-dwelling arthropods was likely caused by decreased temperatures and reduced foraging ability under the dense litter accumulated by exotic annual grasses.

Plant-Animal Interactions

Soil fungal communities in coastal sage scrub habitats are taxonomically and functionally diverse—​roughly twenty times more taxonomically diverse than coastal sage scrub plant communities (Karst et al. 2013). Fungi can be beneficial, neutral, or detrimental to plant growth. Mordecai (2012) found that soil moisture negatively influenced seed germination rate in both coastal sage scrub and grassland systems, likely through positive effects of soil moisture on fungal growth. Seed germination rate was lower in coastal sage scrub than in grassland and was unaffected by the presence of exotic grass

Biotic disturbances created by small mammals influence shrub seedling recruitment during interfire periods (DeSimone and Zedler 1999). Specifically, herbivores and pocket gophers create gaps between plants within coastal sage scrub stands and in adjacent grasslands, and these gaps facilitate seedling establishment. More rodents are found in coastal sage scrub than neighboring grassland because shrubs provide better shelter, and rodents increase native seedling survival by grazing on 436  Ecosystems

Soil Biota

Aboveground dry mass (g plant-1)

1.0 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0 0

25

100

200

500

Grasses m-2 FIGURE 2 2.7 Mean dry mass of California sagebrush (A. californica) seedlings grown in competition with exotic annual grasses planted at varying densities. Source: Redrawn from Eliason and Allen 1997.

thatch, suggesting that exotic grass invasion into coastal sage scrub might not reduce seed survival by promoting fungal growth. Working across nine sites that varied in nitrogen deposition, Egerton-Warburton and Allen (2000) found that the richness and diversity of arbuscular mycorrhizal (AM) fungi declined as nitrogen deposition increased. This increase was accompanied by shifts in AM community composition with important ramifications for plant community composition and performance. In a study of the fungal communities associated with native (A. californica) and exotic (B. rubens) components of coastal sage scrub, Sigüenza et al. (2006) found that while native species were colonized primarily by AM fungi, with reductions in colonization under high nitrogen availability, exotic species were instead associated primarily with a fine AM endophyte that promoted growth regardless of nitrogen availability. Pink-pigmented facultative methylotrophic (PPFM) bacteria are thought to benefit plants by excreting growth hormones that positively influence seed germination and root growth and by excreting osmoprotectants that protect plants from desiccation. In return, PPFM bacteria use simple carbon compounds, such as methanol, generated by plants. In a survey of annual and perennial coastal sage scrub species, Irvine et al. (2012) found that PPFM bacteria are more abundant in the root zones of annual plants, likely due to greater carbon production rates by actively growing roots. As many of the most aggressive invaders in coastal sage scrub are annual species, PPFM bacteria might benefit invasive species over native species in coastal sage scrub and adjacent grassland systems.

Variation of the System in Time and Space Temporal Variation in Plant Community Composition

!

Temporal variation in coastal sage scrub plant diversity is strongly tied to fire cycles, with diversity peaking in the first two years following a fire, particularly on south-facing slopes (O’Leary 1990). Historically, fire is more common in coastal sage scrub than in neighboring chaparral communities. As natural sources of ignition are rare in coastal areas, fire-return intervals likely were on the order of thirty to one hundred years (Rundel 2007). The drought-deciduous shrubs that dominate coastal sage scrub communities are less fire-

resistant than shrubs in chaparral communities because they lack woody root crowns or burls. The ability of shrubs to resprout following fire depends on plant age and fire intensity. Older stands can burn more intensely than younger stands, and few coastal sage scrub shrub species resprout following intense fire (Westman et al. 1981, Malanson and O’Leary 1982), particularly older plants (Keeley et al. 2005a). However, two suffrutescent species (A. glaber and Helianthemum scoparium) are favored by high-intensity fire (Keeley et al. 2005a), which triggers the germination of their dormant seeds. Postfire succession involves a shift in composition among three regeneration modes: obligate resprouters, facultative seeders, and obligate seeders (Keeley et al. 2006). Obligate resprouters regenerate vegetatively, have deep roots that tap into groundwater, are drought sensitive, lack a dormant seed bank, and have low seedling recruitment, especially in low precipitation years. Facultative seeders are capable of both vegetative regeneration and postfire germination of dormant seed banks. Obligate seeders do not regenerate vegetatively following fire and instead rely on dormant seed banks in the soil. In the first five years following fire, facultative seeders are most abundant (60–​70% cover), while obligate resprouters (20–​25% cover) and obligate seeders (0–​5 % cover) comprise a smaller fraction of the community (Keeley et al. 2006). Seedlings of obligate seeders (e.g., annual and perennial herbaceous species) are most abundant the first year following fire and can disappear if intervals between fires are long (Keeley et al. 2005b, 2006). Many coastal sage scrub shrubs produce copious seeds in the first few years following fire, but this levels off after the canopy closes (Keeley et al. 2005). Closing the canopy can ameliorate drought stress and promote seedling establishment of obligate resprouters, which take longer to establish following fire (Keeley 1992). Several obligate resprouting subshrubs such as Encelia californica and Hazardia squarrosa are not subject to these limitations and often display numerous seedlings resulting from seeding of resprouted plants in the second year following fire (Keeley et al. 2006). For many other shrub species, few seedlings might establish in the first year following fire because of limited seed dispersal from neighboring, unburned areas (Minnich and Dezzani 1998, Rundel 2007). In larger burned areas with no adjacent unburned area, the introduction of seed will come primarily from resprouted individuals (Keeley et al. 2005b). This can open a window of opportunity for exotic species with superior dispersal traits to establish and preempt resources prior to native species establishment (Keeley et al. 2005; Keeley and Brennan 2012), initiating a type-conversion from shrubland to grassland. Fire can affect the structure of animal communities (Figure 22.8); species that require open habitat often increase in abundance following fire, at the expense of species that require mature, shrub-dominated habitat (small mammals [Brehme et al. 2011], birds [Mendelsohn et al. 2008], herpetofauna [Rochester et al. 2010]). Furthermore, Diffendorfer et al. (2012) found that mammal species composition following fire was more strongly influenced by habitat heterogeneity + and vegetation composition than by fire intensity or distance to adjacent unburned areas. Fire can also lead to declines in species diversity (herpetofauna [Rochester et al. 2010], ants [Matsuda et al. 2011]), although for some groups the shift in community composition following fire is not accompanied by changes in diversity (e.g., small mammals [Brehme et al. 2011, Diffendorfer et al. 2012]). Coastal Sage Scrub   437

(H)

(B) (C)

(A)

Closed canopy

(D)

(E)

(G)

Fire

(I)

Open canopy (J)

(F)

(L) (K)

FIGURE 2 2.8 Composition of the animal community in coastal sage scrub shifts following fire with the loss of shrub cover. Closed canopies provide habitat for (A) bushtit (Psaltriparus minimus), (B) wrentit (Chamaea fasciata), (C) Anna’s hummingbird (Calypte anna), (D) western toad (Bufo boreas), (E) California mouse (Peromyscus californicus), and (F) pacific treefrog (Pseudacris regilla). Postfire, open canopies favor a different suite of species, including (G) lazuli bunting (Passerina amoena), (H) horned lark (Eremophila alpestris), (I) western fence lizard (Scheloporus occidentalis), (J) kangaroo rat (Dipodomys sp.), (K) California vole (Microtus californicus), and (L) common side-blotched lizard (Uta stansburiana). Illustration: Julie E. Larson.

Transitions between Coastal Sage Scrub and Grassland Although coastal sage scrub can transition between grassland and chaparral or oak woodland as part of a “shifting landscape mosaic” (Callaway and Davis 1993), coastal sage scrub is not generally thought of as simply a successional stage on a trajectory to another vegetation type. Using aerial photographs and vegetation maps, researchers have documented transitions between coastal sage scrub and grassland over multiple decades. Historical records are unclear about how much land was occupied by shrubland, grassland, or forbland prior to the arrival of Franciscan missionaries in the late 1700s (Minnich 2008). However, in the last century, conversions of coastal sage scrub to exotic annual grassland and vice versa have been influenced by fire frequency, grazing pressure, and nitrogen deposition. For example, Talluto and Suding (2008) resampled 232 plots established in the VTM survey in the early 1930s and found that exotic grasses were more dominant in plots with increased fire frequency. While nitrogen availability is widely understood to promote invasion 438  Ecosystems

(Huenneke et al. 1990, Davis et al. 2000), Talluto and Suding (2008) found that nitrogen deposition was positively correlated with grass cover only in plots with low fire frequency in Diegan sage scrub. However, a higher proportion of Riversidean sage scrub was converted to exotic grassland in areas with high nitrogen deposition (Minnich and Dezzani 1998, Fenn et al. 2010). Once exotic annual grasses establish, they can promote increased fire frequency, which maintains their dominance (see Figure 22.5). Annual grasses can increase the fuel load within ecosystems as their dry biomass accumulates (D’Antonio and Vitousek 1992, Keeley et al. 2005a). Repeated, low-intensity fires favor exotic annual grasses for two reasons. First, exotic seed bank survivorship is high when fire intensity is low; and second, a short fire-return interval can inhibit the regeneration of native woody species, particularly obligate seeding shrubs (Keeley and Brennan 2012). Shrub colonization into grassland can also be limited by a depleted seed bank, low dispersal from neighboring areas, and competition from grasses (Stylinski and Allen 1999, Cione et al. 2002). Cases have been documented of conversion from grassland

1.5

A

500

LRR total flowers

Total flowers (x1000)

600

400 300 200 100 0

B

1.2 0.9 0.6 0.3 0

SD 32.9

SM 34.1

CAM 35.5

Latitude (South

SC GG 36.9 37.8

North)

GG SC 0.29 0.31

CAM SD 0.39 0.41

SM 0.48

Precipitation CV

FIGURE 2 2.9 (A) Number of flowers per shrub grown under high (closed symbols) or low (open symbols) rainfall, for California sagebrush (A. californica) collected at five locations. From south to north: SD=San Diego, SM=Santa Monica, CAM=Cambria, SC=Santa Cruz, GG=Golden Gate National Recreation Area. (B) The log response ratio (LRR) of the flowering response to increased water availability is strongly predicted by the coefficient of variation (CV) of annual precipitation at these five sites. Source: Based on Figures 2c and 3c in Pratt and Mooney 2013.

to shrubland (Freudenberger et al. 1987, Callaway and Davis 1993, DeSimone and Zedler 2001). For example, an analysis of vegetation from 1946 to 1955 at Starr Ranch in south Orange County, California, found a shift from native perennial grassland to shrubland (DeSimone and Zedler 2001). These authors found that four woody species (A. californica, A. glaber, E. fasciculatum, and S. apiana) colonized grassland by producing large seed crops, recruiting in gaps (particularly those created by pocket gophers), and displaying rapid seedling growth and low susceptibility of seedlings to herbivory. In contrast, in northern coastal scrub, Baccharis pilularis (coyote brush) invades adjacent grasslands in the absence of gaps created by disturbance (Hobbs and Mooney 1986). Shrub establishment in these northern coastal scrub systems tends to be episodic, associated with high rainfall years (Williams et al. 1987), while in southern coastal sage scrub DeSimone and Zedler (2001) found that woody species establishment was relatively insensitive to annual precipitation. Thus the drivers of coastal sage scrub and grassland transitions vary across the range of this vegetation type.

Human Impacts from the Postcolonial Era to Present Climate Change Southern California will likely be a hotspot of future climate change (Diffenbaugh et al. 2008), experiencing both drier and more variable conditions (Cayan 2008, Schubert et al. 2008, see Chapter 2, “Climate”). Dynamic vegetation models predict that shrublands in southern California (including both coastal sage scrub and chaparral) will decline in extent in response to drier conditions and greater fire frequency and will be replaced by grasslands (Lenihan et al. 2003, 2008). Models predicting the shift in vulnerable species across the landscape with future climate change illustrate the additional challenge of the highly fragmented nature of coastal sage scrub habitat. For instance Hannah et al. (2012) identified four areas of California requiring additional land protection to create corridors to facilitate range shifts of vulnerable species. One of these sites spans Santa Barbara and Ventura

Counties where Salvia leucophylla, a key coastal sage scrub shrub species, will likely decline without these additional protections. Animal populations in coastal sage scrub will likely be strongly affected by climate change via shifts in the abundances of food sources and nesting habitat. Two federally listed coastal sage scrub animals, the Quino checkerspot butterfly and the California gnatcatcher, were predicted to have smaller suitable habitat areas under higher temperature and diminished areal extent of coastal sage scrub (Preston et al. 2008). New populations of Quino checkerspot butterflies have been observed in higher-elevation (cooler and moister) chaparral, with interspersed forblands supporting larval host plants as predicted by global warming models (Preston et al. 2012). Variation in rainfall associated with El Niño cycles might also play a role; Morrison and Bolger (2002) found that rufous-crowned sparrows (Aimophila ruficeps) had higher reproductive success in high rainfall (El Niño) years due to both higher food abundance and reduced nest predation by snakes. Relatively little experimental work has evaluated coastal sage scrub responses to climate change. A recent common garden study collected cuttings of A. californica (California sage, one of the key foundational species of coastal sage scrub) from populations across its range and found that populations from more variable climates also showed the greatest growth responses to high versus low rainfall (Figure 22.9; Pratt and Mooney 2013). Local adaptation among populations thus could play a key role in coastal sage scrub responses to climate change. In contrast to the dearth of experimental work on coastal sage scrub responses to climate change, a great deal of research has addressed the roles of urbanization and agriculture, increasing fire frequency and atmospheric nitrogen deposition on coastal sage scrub loss and decline.

Urbanization and Agriculture The greatest impacts on coastal sage scrub both historically and today are urbanization and agriculture. Valley bottoms in much of California have rich, deep soils suited to irrigated agriculture and/or some dryland small-grain production. HisCoastal Sage Scrub   439

torically, irrigated citrus and avocado orchards and vineyards were planted both in valleys and on hillsides. However, agriculture is in decline as urbanization increases (Chen et al. 2010). Urban development is expanding into abandoned agricultural fields (e.g., Orange County, which was notable for its citrus industry in former coastal sage scrub habitat, no longer produces commercial citrus), and some previously pristine coastal sage scrub areas are also being developed for housing. A statewide analysis comparing the original 1930s VTM (Wieslander 1935) coastal sage scrub acreage to the 2006 vegetation map (Hollander 2007) shows a loss of 6.1% of coastal sage scrub to agriculture and 19.4% to urban development in that period (Table 22.2). The area of coastal sage scrub converted to agriculture prior to the 1930s is unknown. Of the remaining, undeveloped land previously occupied by coastal sage scrub, 15.3% has converted to exotic annual grassland and 44% remains shrubland (33% coastal sage scrub plus 11% chaparral). In Riverside County, 71% of remaining undeveloped coastal sage scrub has converted to exotic annual grassland based on a more detailed California Department of Fish and Wildlife 2006 vegetation map (Evens and Klein 2006). Type conversion could be related to abandonment from agriculture, but Riversidean sage scrub could be more highly invaded than other coastal sage scrub subassociations because it has lower shrub cover, providing larger interspaces for invasion. It also has hotter summers that cause greater shrub leaf fall, further reducing canopy cover and enabling invasion at the onset of fall rains. As described below, it also has greater levels of nitrogen deposition. Most commercial grazing in coastal sage scrub declined greatly by the 1930s (Robinson et al. 1993); these lands were used for agriculture, abandoned and allowed to recover, or eventually urbanized. Grazing has persisted to the present in some areas, and recent grazing was related to local extirpation of Quino checkerspot butterflies (Preston et al. 2012). Native forbs declined relative to exotic grasses and forbs under domestic grazing in annual forblands (Kimball and Schiffman 2003). Grazing has also been blamed for the spread of invasive species. Fecal analyses have demonstrated that domestic grazing animals move seeds (Malo and Suarez 1995), but seeds also move along roadsides and other corridors (Zink et al. 1995). Domestic grazing might have hastened dispersal, but with widespread human disturbances in coastal sage scrub and possibly native animal dispersers, these seeds would have arrived eventually (Minnich and Dezzani 1998, Minnich 2008). By whatever means invasive species arrive, their productivity is assured in relatively nutrient-rich coastal sage scrub soils, and atmospheric nitrogen inputs further stimulate productivity.

Nitrogen Deposition The critical load of anthropogenic nitrogen deposition of 10 kg N ha-1 yr-1 is exceeded in 33% of coastal sage scrub land area, a higher proportion than any other California ecosystem type (Figure 22.10; Fenn et al. 2010). A critical load is the value of a pollutant input rate above which negative ecosystem consequences occur. In coastal sage scrub, negative impacts of elevated nitrogen include increased exotic grass productivity causing increased fine fuels for fire (Fenn et al. 2010). The amount of nitrogen added increased exotic grass biomass to the fire threshold value of greater than 1 ton ha-1 440  Ecosystems

TA B L E 2 2 . 2 Remaining coastal sage scrub vegetation and conversion to agriculture and urban development in 2006 relative to the 1930s vegetation type mapping (VTM) survey

Map year

Cover type

Hectares

Percentage of original VTM

1930s VTM

Original VTM coastal sage scrub

715,132

100.0

2006

Converted to agriculture

43,401

6.1

Converted to urban

138,914

19.4

Converted to exotic grassland

109,350

15.3

Remaining coastal sage scrub

234,494

32.8

Classified as chaparral

81,457

11.4

Classified as woodland, forest

81,256

11.4

Classified as other types

26,260

3.7

Source: Hollander 2007. The 1930s VTM survey is at http://vtm .berkeley.edu. Note: Areas that were classified as chaparral, woodland, or forest in 2006 may have experienced vegetation shifts since the 1930s; or differences may be due to map scaling errors, as in the case of classification to other types (barren, desert, or water). Analysis does not include the Channel Islands, which have approximately 30,000 hectares of coastal sage scrub.

in all but the driest years, while unfertilized plots had lower productivity (Fenn et al. 2003). By contrast, forbs disarticulate in the dry season and are relatively poor carriers of fire across the landscape (Brooks et al. 2004). Elevated nitrogen also reduced richness of native forbs along a nitrogen deposition gradient from 6 to 20 kg N ha-1 yr-1, with a steep drop from sixty-seven to thirty-seven species per site at 10 kg N ha-1 yr-1 and a further drop to sixteen species at 20 kg N ha-1 yr-1 (Fenn et al. 2010). Reduced richness and percent infection of AM fungi occurred along the same gradient, although with a gradual decline rather than a threshold (Egerton-Warburton and Allen 2000). The shift in AM species composition promoted reduced mutualism for the native shrub A. californica, but inoculum from a high-nitrogen-deposition soil was still beneficial for the exotic grass Bromus rubens (Sigüenza et al. 2006). This implies that the poor establishment of A. californica seedlings observed in mixtures with exotic grasses is exacerbated by less effective mycorrhizal inoculum of soils experiencing high nitrogen deposition. An eight-year nitrogen fertilization experiment in coastal sage scrub showed that nitrogen addition increased shrub production only in years when rainfall exceeded 45 cm yr-1 (Vourlitis 2012), but shrub relative abundance shifted to greater dominance by A. californica within five years (Vourlitis et al. 2009). Other changes to the nitrogen cycle under elevated nitrogen include increased plant, soil, and microbial nitrogen levels (Sirulnik et al. 2007b, Vourlitis and Fernan-

Coastal Sage Scrub CL Exceedance kg N ha -1 yr -1 N < 7.8 7.8 < N < 10 N > 10 Monterey

DIABLAN

San Luis Obispo VENTURAN Sources: Nitrogen Deposition: BCOE-CERT (UCR) Terrain: SCAS (OSU)

RIVERSIDIAN

Center for Conservation Biology Angeles N UC - Riverside, Los September 2009

0

100

200

Kilometers

DIEGAN

San Diego

dez 2012), and increased rate of nitrogen but not carbon mineralization (Sirulnik et al. 2007a, Vourlitis and Zorba 2007, Vourlitis et al. 2007). The implication of the latter is that increased nitrogen absorbed by plants will cycle back to the soil as mineral nitrogen, but there will be no change in soil carbon, resulting in decreased soil C:N. Overall, legislative efforts to curb nitrogen deposition below critical loads should be encouraged. Those areas below critical loads of 10 kg N ha-1 yr-1 will be the best candidates for conservation reserves to maintain current diversity and ecosystem functioning.

Change in Fire Frequency Estimating the historic change in fire frequency in coastal sage scrub is difficult because historical fire frequencies are unknown and estimates of current frequencies are sparse. Native Californians might have used fire as a tool to convert coastal sage scrub to forb- and grasslands (Keeley 2002). Spanish explorers of the mid-1700s reported extensive and frequent, sometimes annual, Indian burns in the Los Angeles Basin (Minnich 2008). One reason for this could have been that fire promoted the dominance of native annual forbs, as seeds of some species were a staple food. Estimates of midto late-twentieth-century fire return intervals are twenty years in Venturan sage scrub of the Santa Monica Mountains (Westman and O’Leary 1986) and twenty-five years in Riversidean sage scrub of the Box Springs Mountains (based on aerial photography, Minnich unpublished). Currently, the fire return interval in the Box Springs Mountains in Riverside is only about ten years (Minnich unpublished), possibly a result of increased exotic grass productivity from nitrogen

FIGURE 2 2.10 Critical loads (CL) of anthropogenic nitrogen (N) deposition in coastal sage scrub, which increase exotic grass productivity and reduce richness of native forb and arbuscular mycorrhizal species. The higher value of 10 kg N ha-1 yr-1 is based on modeled CMAQ (Congestion Monitoring and Air Quality) data; the lower level of 7.8 is based on empirical data (data from Fenn et al. 2010).

deposition (Fenn et al. 2010) and anthropogenic ignitions from local development. At the military reserve Camp Pendleton, many coastal sage scrub sites had two to eight fires in a twenty-six-year period in the late twentieth century (Fleming et al. 2009). A recent analysis shows that fire frequency has increased throughout coastal sage scrub vegetation (see Chapter 3, “Fire as an Ecosystem Process”). The difficulty with determining fire return interval in coastal sage scrub is that it is commonly interspersed with chaparral, such that landscape-scale analyses easily confound the two vegetation types. Focused analyses of larger-scale coastal sage scrub landscapes are needed to determine fire return interval. Fire has direct effects on biogeochemistry as well as indirect effects from altered plant-soil feedbacks of invasive species that colonize after fire. While we are not aware of studies that examine how frequent fire affects soil nutrients in coastal sage scrub, a southern California perennial grassland (in a matrix of coastal sage scrub) that was burned at sevenyear intervals showed no changes in soil carbon or in total or extractable nitrogen (Allen et al. 2011). Vegetation with low fuel loads such as coastal sage scrub burn at relatively low temperatures, thus preserving soil nutrients (Allen et al. 2011). While stocks of nitrogen and carbon in plant tissue are depleted by fire, soil nitrogen and carbon can remain little changed, although soil mineral nitrogen is often ephemerally increased immediately postfire because of deposited mineralized plant nitrogen. This pattern was observed in burned chaparral that experienced no long-term impacts of fire on soil nutrient cycling (Fenn et al. 1993). Stocks of nitrogen and carbon in vegetation typically recover rapidly during succession following fire in semiarid shrublands (Allen et al. 2011). However, when coastal sage scrub has been type-­ Coastal Sage Scrub   441

converted to annual grassland under frequent fire, long-term and persistent impacts to soils can be expected such as elevated soil nitrogen and increased rates of nitrogen cycling (Dickens et al. 2013). These can be restored once shrubs have reestablished. Soil total nitrogen and carbon are not changed by exotic grass invasion into coastal sage scrub, although their spatial distribution is more homogeneous (Dickens et al. 2013).

Fragmentation: Impacts on Plants and Animals Fragmentation of coastal sage scrub by urban development has caused declines in both plant and animal species. The number of native plant species decreased and exotic species increased in coastal sage scrub patches isolated up to eightysix years ago in San Diego County (Alberts et al. 1993). While size was the major determinant of species richness, average native species loss over all patch sizes was approximately 43%. The overall richness of native coastal sage scrub plant species was higher than exotic invasive species richness, but native and exotic species richness were positively correlated at all scales of observation from 1 to 400 m 2 in the Santa Monica Mountains (Sax 2002). Conversely, cover of exotic species can be much higher than of native species in some coastal sage scrub fragments, impacting habitat quality for wildlife. Crooks and Soulé (1999) proposed that habitat fragmentation reduces the abundance of top-level predators like coyotes (Canis latrans), leading to an increase in mesopredators (e.g., gray fox [Urocyon cinereoargenteus], Virginia opossum [Didelphis virginianus], domestic cat [Felis sylvestris catus]) and a consequent decrease in bird diversity. However, Patten and Bolger (2003) argued that trophic cascades are more complex in coastal sage scrub habitat. Animals other than mesopredators, such as avian predators and snakes, contribute to nest predation. Additionally, raptors and snakes can consume mesopredators. Thus interactions among predatory groups complicate our understanding of how fragmentation affects food webs and, consequently, bird diversity. Coastal sage scrub habitat loss to exotic species or development has negatively affected dozens of mammal, bird, and reptile species (reviewed in Keeley and Swift 1995). Working across a gradient of exotic plant cover, Diffendorfer et al. (2007) found that animal species richness decreased with increasing exotic plant cover only for some groups (e.g., small mammals and birds) in some years. Overall, species richness did not change along the exotic plant gradient. Instead, animal taxa favoring native woody shrubs were replaced with those favoring grassy, exotic-dominated habitat. Urban development over the past decade has drastically reduced the population of San Diego cactus wren in Los Angeles County, even in protected wildlife areas on the Palos Verdes Peninsula (Cooper et al. 2012). In contrast, habitat fragmentation increases the abundance of native and non-native spiders with corresponding increases in taxonomic diversity of nonnative spiders (Bolger et al. 2008). Increased spider density in smaller fragments may arise because of higher primary productivity, as increased edge area results in higher water flow into fragments, or because of more favorable microsites due to shading of non-native tree species. Edge effects also impact birds and mammals of coastal sage scrub, because of either poor habitat quality caused by dominance of invasive plants at edges or other impacts that reduced their abundance at edges (Kristan et al. 2003). 442  Ecosystems

Management and Restoration The major management issues for coastal sage scrub arise from exotic species invasions that have caused vegetationtype conversion, frequent fire that prevents natural successional processes from recovery of invaded coastal sage scrub, fragmentation of coastal sage scrub habitat from urban development (Soulé et al. 1988, Bolger et al. 1997), and accompanying declines of rare and listed plant and animal species. The direct loss of habitat to urbanization has prompted development of habitat conservation plans (HCPs) mandated under the Endangered Species Act to identify and protect remaining critical habitat and is reviewed below. In this section we focus on restoration of protected habitat fragments that have been degraded by invasive species, the most pervasive conservation problem for coastal sage scrub. Some abandoned agricultural lands are also being restored to coastal sage scrub, and these have the combined challenges of invasive species and lack of a native seed bank. Invasive species affect soil chemical and microbial properties, but these feedbacks have not been the major factor limiting restoration. Negative feedbacks of soils conditioned by invasive grasses did not limit field establishment of coastal sage scrub shrubs (Yelenik and Levine 2011). Conversely, native shrubs can have different effects on mineralization rates, promoting or inhibiting exotic grass colonization (Yelenik and Levine 2010). Phenologies of native shrub and invasive grass were major determinants of soil microbial activity and rates of nitrogen mineralization, but invaded soils recovered once native shrubs were reestablished (Dickens et al. 2013). The low impact of exotic grasses on biogeochemical processes is surprising given the many decades they have dominated some areas of coastal sage scrub. In fact, soil total nitrogen and carbon did not change with invasion, indicating that soil chemical and biological properties are resilient to invasion (Dickens et al. 2013). Conversely, invaded vegetation is itself resilient and requires active restoration including invasive plant control and, often, seeding or planting of native species. Control of abundant exotic annuals requires seed bank reduction, as annual plants require seed banks to persist. The exotic soil seed bank in historically grazed Riversidean sage scrub was greater than 10,000 seeds m-2, while native species had only 400 seeds m-2 (Cox and Allen 2008a). In an area of high nitrogen deposition and frequent fire, no native seed bank was left at all (Cione et al. 2002). Similarly, areas abandoned from agriculture have little or no native seed bank (Allen et al. 2005). Methods for exotic seed bank control include fire both before and after seed dispersal (Gillespie and Allen 2004, Cox and Allen 2008a, Cox and Allen 2008b), solarization (soil heating with plastic) (Marushia and Allen 2010), herbicides and dethatching (Cione et al. 2002, Allen et al. 2005), and mowing and mechanical control (DeSimone 2006). Timed grazing to remove plants before they go to seed appears to be an obvious technique but is seldom used in coastal sage scrub both because it is difficult to time grazing during a short growing season and because of public and agency perceptions that domestic grazing animals are undesirable on public lands (Sulak and Huntsinger 2007). Drought is part of the natural cycle of coastal sage scrub and can deplete the seed banks of exotic invasives with shortlived seed banks. For instance, the exotic annual grass Bromus rubens is subject to local seed bank extirpation following several dry years in the desert (Salo 2004), and seed bank deple-

tion has been observed in coastal sage scrub (Minnich 2008). Restoration ecologists might be able to take advantage of natural fluctuations in exotic seed banks to restore native plants, but considerable work is needed to understand seed bank longevity and susceptibility to drought and other factors. While passive restoration has resulted in stable, nativedominated coastal sage scrub communities in some cases (DeSimone 2013), active restoration might be required in severely disturbed or invaded coastal sage scrub soils, where natural succession might not effectively restore native vegetation (Stylinski and Allen 1999). Revegetation success after exotic species control has been variable. Removal of invasive grass species led to an increase of exotic forb species, especially Erodium spp., rather than of seeded native shrub species, although native forb species increased somewhat (Cox and Allen 2008b). This suggests that shrubs are part of a competitive hierarchy in order of decreasing aggressiveness from exotic annual grasses to exotic forbs, native forbs, and native shrub seedlings (Cox and Allen 2011), and that exotic forbs as well as grasses might need to be controlled for successful restoration. Managers might take advantage of the early establishment phenology of invasive forbs and grasses and control them early in the season, before they can negatively impact native species (Wainwright et al. 2012). Mulch addition and litter removal have both been used to control exotic species and establish native species, with variable outcomes. Mulch immobilizes excess soil nitrogen and has been used for experimental restoration of nitrogenimpacted coastal sage scrub (Zink and Allen 1998, Cione et al. 2002). High litter cover from exotic grasses increased growth of mature native shrubs through increased soil moisture (Wolkovich et al. 2010), while litter reduction either improved or had no significant effect on establishment of native plants from seed (Allen et al. 2005, Cox et al. 2008b, McCullough and Endress 2012). A more impractical aspect of mulch addition or removal is that it can only be used for relatively small-scale restorations and not for extensive invaded landscapes. Fire, herbicides, grazing, and mowing to reduce exotic grass cover are more useful for large-scale restoration. Restored as well as natural coastal sage scrub vegetation might be unstable under the constant threat of invasion from exotic species (Allen et al. 2005, Cox and Allen 2008b), although the long-term success of coastal sage scrub restoration efforts varies considerably. An experimental removal study showed that exotics return to pretreatment levels within four to five years without continued management (Allen et al. 2005). A factor that promotes invasion resistance is the closed shrub canopy in Venturan and Diegan sage scrub, compared to the more invasible open canopy of Riversidean sage scrub (Westman 1983). Once invasive grasses dominate shrub interspaces, the threat of fire can increase, making imperative the need for grass control. The real test of coastal sage scrub restoration will come when a restored site has burned and is able to recover naturally by succession (Bowler 2000). Restored, burned Riversidean sage scrub reverted to exotic grassland (Cione et al. 2002, Allen unpublished), while burned, restored Diegan sage scrub recovered (Margot Griswold unpublished). Passive restoration was successful after invasive plant control in more mesic Diegan sage scrub (DeSimone 2013). Given the high degree of invasion and continual efforts to restore coastal sage scrub in some parts of its range, this vegetation can be considered a highly managed, hybrid ecosystem (as defined by Hobbs et al. 2009, maintaining some native species but also having persistent novel components).

The major plant diversity lies in the native forbs—​as many as two hundred species throughout the range of coastal sage scrub (Skinner and Pavlik 1994), but these are the most difficult to restore. In many areas today, particularly those close to human population centers, coastal sage scrub has an overstory of native shrubs with an understory dominated by exotic annuals. Managers often apply site-specific treatments to increase populations of sensitive and listed plant and animal species. These include mowing grass and even removing shrubs to maintain Stephen’s kangaroo rat (Dipodomys stephensi) habitat (Price et al. 1994, Kelt et al. 2005), restoration plus artificial burrows for burrowing owls (Athene cunicularia), and exotic plant control coupled with seeding host plants (especially Plantago erecta) for Quino checkerspot butterfly habitat (Marushia and Allen 2011). The resulting habitat often is not pristine and is continually invaded by existing and even new invasive species; but with concerted management, rare species can be maintained.

Future Scenarios Because of its location on the coast and relatively gentle terrain, coastal sage scrub vegetation has been converted to a greater extent than most other vegetation types in California. The maximum potential development scenario for southern California shows conversion of most coastal sage scrub vegetation except for on steep slopes (Landis 2006). However, these scenarios are likely to be ameliorated due to coastal sage scrub’s high concentrations of endangered, threatened, and sensitive species, which render many areas subject to HCPs under the federal Endangered Species Act (ESA) or natural communities conservation plans (NCCP) under State of California regulations. Figure 22.3 summarizes the conservation status of coastal sage scrub to date: 15.5% is protected under HCPs, NCCPs, or other conservation reserves; another 19.8% is multiple use such as BLM and Forest Service; 10.7 % is designated as “no protection mandated,” lands such as military or Indian reservations that are primarily wildlands and can be developed but are subject to some protections under the ESA; and 54% is unprotected. Conservation programs are expanding and will acquire lands currently in the unprotected category. Two of the counties with both the fastest rate of urban development and the most coastal sage scrub, San Diego and Riverside Counties, have active HCPs that are acquiring habitat for listed species. For instance, western Riverside County has only 12% coastal sage scrub by land area, but of the approximately 200,000 hectares designated for HCP protection, a disproportionately large share (more than 24,000 hectares) will likely be coastal sage scrub because of the large number of listed species relying on that habitat type. These lands are still being purchased, so the proportion of coastal sage scrub they will eventually contain is unknown (www.rctlma.org/ mshcp). The urbanization of agricultural lands and their loss as buffers with coastal sage scrub has increased the amount of urban-wildland interface (Scott et al. 2006), further necessitating large coastal sage scrub preserves. The Orange County NCCP requires conservation protection of15,000 hectares, a large proportion of which is coastal sage scrub (www.dfg. ca.gov/habcon/). Counties with coastal sage scrub vegetation but without HCPs either have lower development rates and therefore reduced threats to coastal sage scrub species or are under negotiation to develop plans for habitat conservation. Coastal Sage Scrub   443

Habitat connectivity is a major concern for movement, genetic diversity, and protection of coastal sage scrub species and ecosystem functioning (Crooks and Sanjayan 2006), especially because coastal sage scrub is naturally patchy. Undeveloped coastal sage scrub landscapes include a matrix of other vegetation types such as chaparral, native or exotic annual grasslands, and riparian corridors. Planning efforts are under way statewide to maintain corridors between conserved habitat patches, with notable efforts such as the “Tenaja Corridor” linking coastal sage scrub as well as other vegetation types between Riverside and San Diego Counties (Morrison and Reynolds 2006). Emphasis is also on maintaining large patches of coastal sage scrub to reduce edge effects, but success in this effort will depend on which currently unprotected areas can be maintained as conservation reserves. Beyond conservation efforts, the future hope for coastal sage scrub diversity and functioning lies in the large number of restoration efforts. Information on websites and published literature has expanded greatly for coastal sage scrub restoration in the past two decades; these activities are in part supported by legally mandated mitigation funds for off-site development (Bowler 2000). Efforts to improve air quality (see Chapter 7, “Atmospheric Chemistry”) will reduce the longdistance impacts of pollutants such as nitrogen and ozone on this sensitive vegetation type. While restoration efforts do not always result in predisturbance conditions (DeSimone 2013), mitigation has given rise to conservation reserves that are open to public scrutiny and to improved potential for future improvement of diversity and functioning.

Summary Coastal sage scrub is a taxonomically and functionally diverse plant community. The dominant native shrub species tend to be drought-deciduous and more shallowly rooted than the evergreen species in neighboring chaparral communities; as a result, a long history of ecophysiological studies has compared the life-history strategies of species that sometimes co-occur at the ecotones between these two ecosystem types (see Chapter 24, “Chaparral”). Coastal sage scrub is also among the most threatened ecosystems, facing numerous challenges for land management and conservation. Its inherently coastal distribution means that this ecosystem faces some of the greatest pressures associated with anthropogenic land-use change in California as well as other environmental changes associated with large human population centers—​in particular, nitrogen deposition from fossil fuel production, accelerated fire regimes, and invasion by exotic species. Exotic annual grasses are especially problematic and often defy attempts at restoration. Recent research in coastal sage scrub has focused on environmental conditions that favor native shrub establishment over the growth of exotic annual grasses, opening windows into our understanding of how factors such as fire, seasonality (phenology), and interactions with soil microbial communities structure species interactions in Mediterranean-type ecosystems.

Acknowledgments We thank Sandra DeSimone and Elizabeth Wolkovich for valuable comments on this chapter. We are very grateful to Robert Johnson of the Center for Conservation Biology at the 444  Ecosystems

University of California at Riverside for analysis of spatial data and assistance with maps.

Recommended Reading Ford, L. D., and G. F. Hayes. 2007. Northern coastal scrub and coastal prairie. Pages 180–207 in M. G. Barbour, T. Keeler-Wolf, and A. A. Schoenherr, editors. Terrestrial vegetation of California. Third edition. University of California Press, Berkeley, California. Minnich, R. A. 2008. California’s fading wildflowers: Lost legacy and biological invasions. University of California Press, Berkeley, California. Rundel, P. W. 2007. Sage scrub. Pages 208–​228 in M. G. Barbour, T. Keeler-Wolf, and A. A. Schoenherr, editors. Terrestrial vegetation of California. Third edition. University of California Press, Berkeley, California.

Glossary Arbuscular mycorrhizal (AM) fungi  A mutualistic association between certain species of fungi and roots of land plants, whereby the fungus penetrates the root cortical cells, allowing the plant to transfer carbohydrates to the fungus, and in turn soil nutrients and water are supplied to the plant. Burl  A widening of the base of a shrub just at or below the soil surface where resources can be stored and largely used to fuel regrowth following biomass loss due to fire. Drought-deciduous  Having short-lived leaves that are shed annually during the summer drought. Endophyte  A microbe (bacteria or fungus) that lives in an endosymbiotic relationship with a plant without causing harm and often having benefit to the plant. Facultative  Meaning that the phenomenon is capable of occurring but is not required, often referring to plant regeneration capacity: the opposite of “obligate.” Litter  The accumulation of dead plant material on the soil surface, “litter” can refer to either herbaceous material or woody debris. Obligate  Equivalent to “required” or “necessary.” Often referring to regeneration capacity of plants, the opposite of “facultative.” Suffrutescent  Being slightly woody at the base but not having woody tissue throughout the support structure of a plant. Generally refers to the woody base of subshrubs, which have otherwise nonwoody stems.

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from plant litter in exotic annual grasslands. Soil Biology and Biochemistry 39:24–32. Sirulnik, A. G., E. B. Allen, T. Meixner, M. E. Fenn, and M. F. Allen. 2007b. Changes in N cycling and microbial N with elevated N in exotic annual grasslands of southern California. Applied Soil Ecology 36:1–9. Skinner, M.W., and B. M. Pavlik. 1994. CNPS inventory of rare and endangered vascular plants of California. California Native Plant Society, Sacramento, California. Soulé, M. D., D. T. Bolger, A. C. Alberts, R. Sauvajot, J. Wright, M. Sorice, and S. Hill. 1988. Reconstructed dynamics of rapid extinctions of chaparral-requiring birds in urban habitat islands. Conservation Biology 2:75–​92. Steers, R. J., S. L. Fritzke, J. J. Rogers, J. Cartan, and K. Hacker. 2013. Invasive pine tree effects on northern coastal scrub structure and composition. Invasive Plant Science and Management 6:231–​242. Stylinski, C. D., and E. B. Allen. 1999. Lack of native species recovery following severe exotic disturbance in southern Californian shrublands. Journal of Applied Ecology 36:544–​554. Sulak, A., and L. Huntsinger. 2007. Public land grazing in California: Untapped conservation potential for private lands? Working landscapes may be linked to public lands. Rangelands 29:9–​12. Talluto, M. V., and K. N. Suding. 2008. Historical change in coastal sage scrub in southern California, USA, in relation to fire frequency and air pollution. Landscape Ecology 23:803–​815. USGS GAP Analysis Program. 2013. Protected areas database of the United States. . Accessed July 2013. Vourlitis, G. L. 2012. Aboveground net primary production response of semi-arid shrublands to chronic experimental dry-season N input. Ecosphere 3:art22. Vourlitis, G. L., and J. S. Fernandez. 2012. Changes in the soil, litter, and vegetation nitrogen and carbon concentrations of semiarid shrublands in response to chronic dry season nitrogen input. Journal of Arid Environments 82:115–122. Vourlitis, G. L., and G. Zorba. 2007. Nitrogen and carbon mineralization of semi-arid shrubland soil exposed to long-term atmospheric nitrogen deposition. Biology and Fertility of Soils 43:611–615. Vourlitis, G. L., S. C. Pasquini, and R. Mustard. 2009. Effects of dryseason N input on the productivity and N storage of Mediterranean-type shrublands. Ecosystems 12:473–​488. Vourlitis, G. L., G. Zorba, S. C. Pasquini, and R. Mustard. 2007. Chronic nitrogen deposition enhances nitrogen mineralization potential of semiarid shrubland soils. Soil Science Society of America Journal 71: 836–842. Wainwright, C. E., E. M. Wolkovich, and E. E. Cleland. 2012. Seasonal priority effects: Implications for invasion and restoration in a semi-arid system. Journal of Applied Ecology 49:234–​241. Westman, W. E. 1983. Xeric Mediterranean-type shrubland associations of Alta and Baja California and the community/continuum debate. Vegetatio 52:3–​19. ———. 1981a. Diversity relations and succession in Californian coastal sage scrub. Ecology 62:170–​184. ———. 1981b. Factors influencing the distribution of species of Californian coastal sage scrub. Ecology 62:439–​455. ———. 1981c. Seasonal dimorphism of foliage in Californian coastal sage scrub. Oecologia 51:385–​388. Westman, W. E. and J. F. O'Leary. 1986. Measures of resilience: the response of coastal sage scrub to fire. Vegetatio 65:179–189. Westman, W. E., J. F. O’Leary, and G. P. Malanson. 1981. The effects of fire intensity, aspect, and substrate on postfire growth of Californian coastal sage scrub. Pages 151–​179 in N. S. Margaris and H. A. Mooney, editors. Components of Productivity of Mediterranean Regions, The Hague, The Netherlands. Wieslander, A. E. 1935. A vegetation type map of California. Madroño 3:140–144. Williams, K., R. J. Hobbs, and S. P. Hamburg. 1987. Invasion of an annual grassland in northern California by Baccharis pilularis ssp. consanguinea. Oecologia 72:461–​465. Wolkovich, E. M. 2010. Nonnative grass litter enhances grazing arthropod assemblages by increasing native shrub growth. Ecology 91:756–​766. Wolkovich, E. M., D. A. Lipson, R. A. Virginia, K. L. Cottingham, and D. T. Bolger. 2010. Grass invasion causes rapid increases in ecosystem carbon and nitrogen storage in a semiarid shrubland. Global Change Biology 16:1351–​1365.

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Veneklaas, and R. Villar. 2004. The worldwide leaf economics spectrum. Nature 428:821–​827. Yelenik, S. G., and J. M. Levine. 2011. The role of plant-soil feedbacks in driving native-species recovery. Ecology 92:66–​74. ———. 2010. Processes limiting native shrub recovery in exotic grasslands after non-native herbivore removal. Restoration Ecology 18:418–​425. Yoshida, L. C., and E. B. Allen. 2001. Response to ammonium and nitrate by a mycorrhizal annual invasive grass and native shrub in southern California. American Journal of Botany 88:1430–​1436. Zink, T., and M. Allen. 1998. The effects of organic amendments on the restoration of a disturbed coastal sage scrub habitat. Restoration Ecology 6:52–​58. Zink, T. A., M. F. Allen, B. Heindl-Tenhunen, and E. B. Allen. 1995. The effect of a disturbance corridor on an ecological reserve. Restoration Ecology 3:304–310.

T WENT Y-THREE

Grasslands VALER IE T. E VINER

Introduction Much of California’s original grassland habitat has been lost to both changes in hydrology and in urban and agricultural development. Even with this extensive habitat loss, more than 10% of California’s land area remains covered by grasslands today (Corbin et al. 2007a, Barbour et al. 2007). These remaining grasslands are among California’s most altered ecosystems (Corbin et al. 2007a, Jantz et al. 2007). Non-native plant species comprise more than 90% of plant cover in most grassland areas, with many sites below 1% native cover (Bartolome et al. 2007). Even in their non-native dominated state, California’s grasslands are a tremendous diversity hotspot, averaging more than fifty plant species per 30 x 30 meter area (Heady et al. 1992) and providing habitat for nearly 90% of state-listed rare and endangered species (Skinner and Pavlik 1994) and seventy-five federally listed plants and animals (Jantz et al. 2007). These grasslands provide many ecosystem services critical for adjacent agricultural and suburban/urban areas. Almost all of California’s surface water passes through grasslands and oak woodlands (Tate et al. 1999). The grasslands provide high infiltration rates that attenuate storm events, leading to gradual release of storm water to streams (Lewis 1968,

Dahlgren et al. 2001). This reduces flood risk while also maintaining streamflow into the dry season. Grasslands can also improve water quality by filtering pathogens, nutrients, and sediments, serving as effective buffer strips between agricultural and urban uplands and streams (Tate et al. 2006, Atwill et al. 2006). California’s grasslands contribute significantly to regional carbon storage through their large spatial extent and high quantity of carbon storage per unit area (Silver et al. 2010). Grasslands also support many of the pollinators needed in California’s crop systems (Chaplin-Kramer et al. 2011). Direct economic benefits of these grasslands include their provisioning of 75% of the state’s livestock forage (Corbin et al. 2007a, CCCC 2009, Cheatum et al. 2011). Because 88% of California grasslands are privately owned (Jantz et al. 2007), their conservation and restoration depend largely on private land owners and the ways they balance management for livestock production, biotic diversity, and ecosystem services (SRDC 2006, Barry et al. 2006, FRAP 2010, Ferranto et al. 2011). Managing California’s grasslands can be challenging because their structure and function are influenced by multiple, interacting controllers. This produces heterogeneous 449

community and ecosystem dynamics across space and time (Huenneke and Mooney 1989). Grasslands are distributed across a broad range of precipitation regimes; from the highrainfall coasts to drier inland valleys (Figure 23.1). Grassland structure and function vary across this precipitation gradient and with high temporal variability in weather across seasons and years (see Chapter 2, “Climate”). While precipitation patterns are the strongest controller of California grassland dynamics, within the confines of weather patterns grassland structure and function also respond to human management and interactions with soil type, topography, non-native plants, small mammals, insects, microbes, livestock, wild herbivores, and disturbance regimes (Figure 23.2) (Bartolome et al. 2007). California’s Mediterranean climate makes its grasslands distinct from other North American grasslands, which are exposed to a temperate climate where temperature drives the seasonality of plant growth (Corbin et al. 2007b). In contrast, seasonality of precipitation largely governs ecosystem processes in California’s Mediterranean climate, with most plant production occurring during the cool, wet winters and with little plant activity during the hot, dry summers (Corbin et al. 2007b). California’s grasslands differ from other Mediterranean grasslands across the globe because of the stable longterm dominance of annual species in most areas of California. While annual species play an important role in other Mediterranean grasslands shortly after disturbances, successional dynamics eventually lead to domination by perennial species (Rice 1989). Domination of annuals likely makes California’s grasslands more sensitive to fluctuations in abiotic and biotic controllers and is likely the reason behind the need for persistent management to meet many conservation and production goals (Bartolome et al. 2007, Malmstrom et al. 2009). California’s grasslands are experiencing further changes due to shifts in management and the environment, including nitrogen deposition, altered weather patterns, non-native species introductions, and altered grazing and fire regimes. California’s grasslands are also threatened by further land use change. Successful management of these grasslands, particularly in a changing environment, will require site- and regionspecific approaches (Bartolome et al. 2007) because sites from different climate regimes and soil types respond differently to weather variability (George et al. 1988) and management (Bartolome et al. 2007).

Primary Factors Controlling the Distribution of California’s Grasslands California’s 5,640,400 hectares of grassland (Bartolome et al. 2007) are most commonly found in well-drained areas below 1,200 meter elevation (Heady 1977), across a wide diversity of soils (Jackson et al. 2007) and across a broad precipitation gradient ranging from 12 to 200 centimeters per year (Bartolome et al. 2007). Many of the herbaceous species that dominate open grasslands are also key components of other California ecosystems (Bartolome et al. 2007), including oak savannas (Chapter 25, “Oak Woodlands”), shrublands (Chapters

Photo on previous page: California wildflowers in bloom at the North Table Mountain Ecological Reserve, Butte County. Photo: ­Valerie Eviner.

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22, “Coastal Sage Scrub,” and 24, “Chaparral”) and deserts (Chapter 30, “Deserts”). California’s grasslands experience a Mediterranean climate with a mismatch in the timing of ideal temperature versus moisture conditions for plant growth (Figure 23.3). Moisture limits plant growth in the hot and dry summers, while temperature and light limit growth during the cool, wet winters. Ideal periods for plant growth are thus restricted to short periods in the fall and spring (Evans and Young 1989, Bartolome et al. 2007). The growing season begins with the first significant rains (greater than 1.5 centimeters within a week) (Chiariello 1989) and continues until soil moisture declines in the spring (exact timing depends on amount and timing of precipitation as well as the water-holding capacity of soil). When the system dries in the spring, most early- to midphenology annuals set seed and senesce, avoiding the hot, dry summer conditions. Some summer annual species (e.g., tarweeds [Hemizonia, Madia] and wild lettuce [Lactuca]) do grow through the hot, dry summer, using their deep taproots to access moisture (Chiariello 1989, Bartolome et al. 2007). Native perennial grasses often begin growth early in the fall, sometimes even before the rains begin, and can grow later into the summer than most annuals. However, most native perennial grasses do experience aboveground senescence in the summer (Bartolome et al. 2007). This highly seasonal climate, with soil moisture limiting plant growth for four to eight months out of the year (Bartolome et al. 2007), results in stable grasslands even at total annual precipitation levels that, if evenly distributed through the year, would support woody-dominated species in temperate climates. In areas with precipitation patterns that can support either herbaceous or woody species, soils that are fine-textured tend to be dominated by grasslands (Tyler et al. 2007). While environmental conditions shape the distribution of many grasslands, some of California’s grasslands were formerly woody-dominated and were converted through burning, cutting, and herbicide applications (Tyler et al. 2007). The distribution of grasslands across broad precipitation gradients and soils leads to three key subtypes: interior, coastal, and more localized soil-specific grasslands (Keeler-Wolf et al. 2007) (Figure 23.4).

Interior Grassland The most widespread grassland type is interior grassland (also known as valley grassland, south coastal grassland) (see Figure 23.4a). Interior grasslands tend to be distributed in the Central Valley as well as up to 700 meters into the foothills and coastal hills (particularly in the South Coastal hills and in the interior valleys of the Northern Coast Range) (KeelerWolf et al. 2007). Since the early nineteenth century, nonnative grasses and forbs have dominated these grasslands (Keeler-Wolf et al. 2007). Grass cover dominates, but forb species richness is four times greater than grass richness (Sims and Risser 2000). Some native perennial grasses persist in this system (e.g., purple needle grass [Stipa pulchra], valley wild rye [Elymus triticoides], blue wild rye [Elymus glaucus], and California brome [Bromus carinatus]), but their growth, survival, and seed establishment are limited in the interior grasslands by the fast growth, high density, shading effect, and high water use of the highly competitive non-native species (Corbin et al. 2007a). Interior grasslands extend across a wide mean annual precipitation gradient, producing variations in plant

FIGURE 23.1 Distribution of grasslands in California. Additional grassland not shown here occurs in the understory of oak savannas and woodlands in much of the state (see Chapter 25, “Oak Woodlands”). Data from U.S. Geological Survey, Gap Analysis Program (GAP). Map: P. Welch, Center for Integrated Spatial Research (CISR).

Topography

Broad climate gradients

Soil type

Interactions Weather variations

nvir n ental Environmental changes changes

Ecosystem Ecosy y stem processes processes

Vegetation composition

Management

Other organisms

Disturbance regime

Grasslands (also called prairies) along California’s central and north coasts (ranging from San Luis Obispo to southern Oregon) tend to experience longer, wetter growing seasons than inland areas (Ford and Hayes 2007, Keeler-Wolf et al. 2007) (see Figure 23.4b). In addition, fog inputs mitigate summer moisture limitation and can account for 28% to 66% of root water uptake by perennial grasses in summer (Corbin et al. 2005). These wetter conditions (especially when precipitation is greater than 100 centimeters per year) lead to dominance by native and non-native perennial herbaceous and woody species. Common woody invaders include Scotch broom (Cytisus scopartus), French broom (Genista monspessulana), and gorse (Ulex europeus) (Heady et al. 1992, Ford and Hayes 2007). Overall, native cover is higher in coastal than interior grasslands. The annual non-natives that are common in the interior grasslands are often only minor components of the coastal grasslands or restricted to disturbed areas (Corbin et al. 2007a). While some of these coastal grasslands are stable as grasslands, particularly in the drier sites, others are maintained by disturbance regimes such as burning, livestock grazing, and deer browsing that impede the persistence of woody plants (Ford and Hayes 2007).

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25

10

20

8 15

atio n

6 4 2 0

Aug

Oct

ure rat

10

e mp

Te

5

Aboveground plant growth rate Dec

Feb

April

June

0

Average monthly air temperature (˚C)

Coastal Grassland

12

Pre cip it

community composition. Species such as foxtail brome (Bromus madritensis) and red-stemmed filaree (­ Erodium cicutarium), both exotic, tend to be common in dry sites (less than 25 centimeter mean annual precipitation), while exotic species such as soft chess (Bromus hordeaceus), wild oats (Avena barbata), and broad leaf filaree (Erodium botrys) tend to be more common on wetter sites (65–​100 centimeter mean annual precipitation) (Bartolome et al. 1980).

Average monthly precipitation (cm)

FIGURE 23.2 Controls over the structure and function of California’s grasslands. Broad climate gradients are the primary controller of structure and function, while soil type and topography have impacts within the confines of climate conditions. All three of these state factors affect how weather variations influence affect how weather variations influence interactions among vegetation composition, other organisms (e.g., small mammals, large herbivores, microbes, insects), disturbance regimes, and human management. These interactions determine ecosystem processes (C cycling, N cycling, water dynamics), which feed back to affect these interactive factors. Environmental changes (e.g., climate change, nitrogen deposition) can lead to unique interactions that influence grassland structure and function. Illustration by Valerie Eviner.

Month FIGURE 23.3 Seasonal variations in temperature (dashed line) and precipitation (solid line) drive plant growth rate (gray shaded area), with most production occurring when both moisture and ideal growing temperatures are present. Temperature and moisture data are from the California Irrigation Management Information System (CIMIS), averaged across 1985 through 2005 and across grassland sites, including: Sierra Foothills, San Joaquin Valley, Bay Area, Sacramento Valley, North Coast Valley, South Coast Valley, and Central Coast Valley. The left-side y-axis provides the precipitation scale, while the right-side y-axis represents the temperature scale. Aboveground growth rate is not present on either y-axis, but scales from 0 to 200 g/m2/month. Growth rate data is a seasonal average across experimental sites in the Central Valley and North Coast (Eviner, unpublished data). Source: Figure updated from Biswell 1956.

A

B

C

F

E

D

FIGURE 23.4 Diversity of California grassland types. Photos: Valerie Eviner. A Annual grassland with a mix of grasses, forbs, and legumes (Interior Coast Range, Mendocino County). B Native perennial grassland (Coast Range, Marin County). C More recent invasion of late-season non-native grasses (goatgrass, medusa head) in the foreground (green), invading into naturalized

annual non-native grassland (background, senesced) (Interior Coast Range, Mendocino County). D Alkali grassland (Sacramento Valley floor, Glenn County). E Vernal pool (Sacramento Valley floor, Solano County). F Serpentine grassland (Interior Coast Range, Mendocino County).

Subtypes of Grasslands Determined by Unique Soils Unique soil conditions in both interior and coastal grasslands create relatively small patches of distinctive grassland types, including serpentine, alkali sinks, and vernal pools. Serpentine soils are derived from rock from the earth’s mantle. They tend to be nutrient-poor, with low calcium to magnesium ratios, high levels of heavy metals (particularly nickel), and low water availability. These stressful soil conditions lead to a low-productivity system with sparse, short plants and a high degree of endemism (species that are unique to particular locations). Serpentine sites are usually dominated by diverse native forbs, with less than 10% grass cover (see Figure 23.4f). While serpentine grasslands typically have far fewer nonnative species than surrounding grasslands on nonserpentine

soils, invasion from surrounding grasslands does occur, particularly in areas receiving high amounts of nitrogen deposition (Harrison and Viers 2007). Vernal pools are shallow, seasonal wetlands within a grassland matrix, usually found in shallow depressions with an impermeable soil layer (see Figure 23.4e). While the edges of vernal pools may be dominated by upland grassland species, the pools themselves contain a rich diversity of native and introduced grasses and forbs with composition strongly influenced by depth and duration of flooding (Solomeshch et al. 2007). Alkali sinks (see Figure 23.4d) are also seasonal wetlands but with a high pH and high salinity. These foster a rich community of native and introduced grasses and forbs including a number of endemic, threatened, and endangered plants (Heady 1977, Dawson et al. 2007).

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Variations within Grassland Types: Local Controls over Structure and Function While broad precipitation gradients and soil types can determine the distribution of distinct grassland types, each grassland type also contains considerable spatial and temporal variation within it. Interactions of multiple biotic and abiotic controllers mediate heterogeneity in community and ecosystem dynamics (Huenneke and Mooney 1989) (see Figure 23.1), as described in the following sections.

TOPOGR APHY AND SOILS

California’s grasslands occur on diverse soil types, including five of the twelve soil orders. Even at a local scale, soil can be highly heterogeneous, affecting vegetation composition and growth through differences in soil fertility and water infiltration and storage (Jackson et al. 2007). Species such as soft chess are common on many soil types, but soil type limits the distribution of many other species, including wild oats, medusa head (Taeniatherum caput-medusae), and filaree (Evans and Young 1989). Clay soils have higher water-holding capacity and thus tend to increase plant production. Similarly, deep soils increase plant production by providing water to deep-rooted plants (Reever Morghan et al. 2007). Other soil characteristics, such as pH, nutrient content, soil organic matter and texture, can also influence community composition (Hoopes and Hall 2002). Topography also has strong impacts on local heterogeneity in vegetation composition and production, largely through its impacts on microenvironment (McNaughton 1968, Evans and Young 1989, Heady et al. 1992). For example, south-facing slopes are so much drier and hotter than north-facing slopes that the growing season can be one month shorter on south-facing slopes (Hufstader 1978). North-facing slopes tend to favor species with deeper roots, greater water use and later phenology (Ng and Miller 1980). Thus native perennials and late-season invaders such as goatgrass (Aegilops triuncialis) are more common on north-facing than south-facing slopes (personal observation). Germination rates tend to be higher on north-facing slopes (Evans et al. 1975), but it is not clear whether this is due to environmental conditions at germination or variation in seed characteristics determined by environmental conditions during seed production the previous spring. Topography can also alter the impacts of grazing (Huntsinger et al. 2007) and elevation (Bartolome et al. 2007) on community and ecosystem dynamics.

VARIATIONS IN WE ATHER

As discussed previously, the amount and seasonality of moisture and temperature determine the presence of grasslands, while precipitation gradients structure the distribution of grassland types. Variations in grassland structure and function within a given site are strongly driven by fluctuations in weather patterns within a growing season and across years (Heady et al. 1992, Bartolome et al. 2007, Keeler-Wolf et al. 2007). At a given site, annual precipitation can vary as much as 50 centimeters to 100 centimeters from its long-term mean (Pitt and Heady 1978, Reever Morghan et al. 2007), with high variation particularly associated with El Niño–​Southern Oscillation events (Reever Morghan et al. 2007). Lower rain454  Ecosystems

fall years tend to produce lower plant diversity (Bartolome et al. 1980), but total rainfall does not reliably predict plant production and community composition—​the timing of rainfall is far more important than the annual total (Figure 23.5) (Pitt and Heady 1978, George et al. 2001, Reever Morghan et al. 2007, Suttle et al. 2007). Early fall weather conditions can have large impacts on vegetation composition, mediated through plant germination characteristics. The timing and temperature of initial fall rains can influence the germination of rare plants (Levine et al. 2011) as well as the identity of dominant plants (Pitt and Heady 1979). Alternating dominance among grasses, forbs, and legumes has been frequently observed across years in California’s grasslands (Pitt and Heady 1979, Keeler-Wolf et al. 2007) and has been attributed to variations in weather conditions. An initial flush of germinating rains (at least 1.5 centimeters with a week) stimulates rapid germination of the annual grasses, depleting most of their seedbank (Young and Evans 1989, Chiariello et al. 1989, Bartolome et al. 2007). If precipitation continues throughout the fall, grasses dominate the vegetation throughout the growing season. However, when a germinating rain is followed by a dry fall, the germinated grasses are likely to die. In these years grasslands are dominated by forbs (e.g., filaree) that can survive the fall drought or forbs and legumes that germinate with later rains (Young and Evans 1989, Bartolome et al. 2007, Keeler-Wolf et al. 2007). The response of vegetation composition to rainfall patterns can vary greatly across sites, so that the conditions for a “forb year” are likely to result in more frequent patches of forbs across the landscape, among diverse vegetation patches (Jackson and Bartolome 2002). Precipitation patterns in the winter and spring also affect community dynamics. Extended winter or spring drought enhances clovers (Castilleja, Medicago, Melilotus, Orthocarpus, Trifolium) (Corbin et al. 2007a) and alters seed production (Ewing and Menke 1983). Midwinter droughts are common in California’s grasslands, averaging nineteen days without rain in December through January (Reever Morghan et al. 2007). These midwinter droughts favor perennials over annuals, which are less tolerant of dry conditions during the growing season (Corbin et al. 2007a). Spring precipitation strongly impacts the amount and timing of seed production, but the effects vary by species and ecotype. For example, during dry springs some species flower earlier while others have a later but shorter flowering period (Chiariello 1989). Late-spring and early summer rains can enhance the growth and fecundity of late-season species, such as the non-native yellow starthistle (Centaurea solstitialis) and native tarweeds. These latespring rains are unlikely to affect most annual grasses (Pitt and Heady 1978), which are hard-wired to senesce by early summer even in the presence of ample moisture (Jackson and Roy 1986, Chiariello 1989). However, later-season noxious annual grasses, such as medusa head and goatgrass, do benefit from late-season rains (Eviner, Rice and Malmstrom in prep.). In addition to shaping community composition, this temperature and moisture variability strongly regulates the amount and timing of net primary production (discussed later in the chapter, under “Ecosystem Functioning”).

FIRE

Fire can have strong impacts on grassland structure and function, with effects depending on the timing, intensity, and fre-

Biota Diverse biota rely on California’s grasslands for habitat and actively shape grassland structure and function through their interactions (see Figure 23.1).

L ARGE HERBIVORE S

Herbivory is a critical controller of most of the world’s grassland ecosystems, many of which have evolved under grazing pressure. The extent of adaptation to grazing in California’s native flora is unclear. California’s native grassland flora was exposed to grazing and browsing by the rich megafauna present during the Rancholabrean (150,000 years before present [YBP] to 11,700 YBP), including bison, elk, deer, mammoth, pronghorns, horses, and camels. (Edwards 2007). These mega-

4000 B

3000 2000

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0

Dry matter (kg ha-1)

quency of burning (D’Antonio et al. 2006). The effects of any individual fire are generally limited to less than three years (Bartolome et al. 2007) and include decreased soil moisture (Henry et al. 2006), increased soil available nitrogen and phosphorus, and increased rates of nitrogen mineralization and nitrogen fixation (D’Antonio et al. 2006, Reiner 2007). Fire also has short-term effects on soil microbial community composition, with decreased gram negative and positive bacteria (Docherty et al. 2012) and a slight decrease in extracellular enzyme activity (Gutknecht et al. 2010). Over the long term, frequent fires can decrease soil nitrogen and sulfur due to repeated volatilization losses (D’Antonio et al. 2006). Impacts of fire on plant communities are varied (D’Antonio et al. 2006) and depend on the dominant vegetation prior to burns. Fires increase species richness of non-natives in areas dominated by non-natives before the burns and increase natives in native-dominated areas (Harrison et al. 2003). In general, fires increase the prevalence of forbs and legumes by removing thatch, thus increasing light and soil temperature. Sustained increases in forbs require annual burns, but particularly for native forbs, this is only true in ungrazed areas (D’Antonio et al. 2006). This is likely because grazing, like burning, removes thatch, thus increasing legumes and forbs. Spring burns favor native over non-native forbs, although the effects are weak and depend on burn frequency and grazing regimes (D’Antonio et al. 2006). Fires are often timed to control non-native species. For example, to control late-season noxious weeds such as medusa head and goatgrass, prescribed burns are targeted in the late spring, after most other annuals have senesced but before weed seeds have matured and dropped. The senesced annuals are dry enough to support a moderately intense fire, which can kill the seeds of late-­ season weeds. This can decrease weeds over the short-term but must be repeated to maintain weed control (Reiner 2007). Fire regimes have been greatly altered by human activity. Native Americans frequently burned to enhance grassland production, alter grassland communities, and convert shrublands to grasslands (Bartolome et al. 2007). In the nineteenth century, fire frequency in the Central Coast was one to five years, but it now has decreased to twenty to thirty years (Greenlee and Langenheim 1980). Near urban areas, however, fire frequency has increased (Bartolome et al. 2007). These changes in fire regime have strong potential effects on ecosystem and community dynamics (D’Antonio et al. 2006, Bartolome et al. 2007).

4000 F

3000

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1000 0 4000 G 3000 A

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1000 0

Sep Oct Nov Dec Jan Feb Mar Apr May Jun

Month FIGURE 23.5 Seasonal forage production, as affected by seasonal weather patterns, at the San Joaquin Experimental Station (data from 1935 through 1984). Curves are associated with the following weather conditions: (A) average fall, winter and spring; (B) warm, wet fall, average winter and spring; (C) cold, wet fall, average winter and spring; (D) dry fall, average winter and spring; (E) average fall, cold winter, average spring; (F) average fall, mild winter, average spring; (G) average fall, short winter, early onset of warm spring temperatures; (H) average fall, long winter, late onset of warm spring temperatures. Source: George et al. 2001.

fauna were largely absent during the Holocene (11,700 YBP to the present), leaving a significant time period when plants could have adapted to the absence of megafauna. Even in the absence of megafauna, California grasslands continued to experience high rates of herbivory by rodents, rabbits, hares, birds, and elk (Edwards 2007). While deer and pronghorn have been prevalent for part of the Holocene, they are mostly browsers, so their main effect is likely exclusion of woody species with more modest impacts on herbaceous community composition (Edwards 2007). Herbivore identity has large impacts on vegetation composition. Cows and horses preferentially consume grass, while sheep and deer preferentially consume forbs. Antelopes consume grasses, forbs, and shrubs, changing preference with season (Edwards 2007). Similarly, tule elk consume forbs in the spring and summer and consume grasses in the fall and winter (Johnson and Cushman 2007). Removing elk from coastal grasslands decreased annual plant cover and increased some non-native perennial grasses but did not affect other perennials (Johnson and Cushman 2007). Gr asslands  455

8000

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6000

5000

4000

3000

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Year FIGURE 23.6 Effects of residual dry matter (RDM) on aboveground net primary production (ANPP) at Sierra Foothills Research and Extension Center from 2001 through 2003. From left to right within each year, gray bars are 225 kg/ha RDM, dotted bars are 560 kg/ha RDM, white bars are 900 kg/ha RDM, and black bars are 5,000 kg/ha RDM. Within each year, bars labeled with different lowercase letters differ significantly from each other. Source: Bartolome et al. 2007.

Absolute cover (%)

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Residual dry matter (kg ha -1 ) FIGURE 23.7 Impacts of residual dry matter (RDM) treatments on cover of forbs (left) and clovers (right) across five years at Sierra Foothills Research and Extension Center. Source: Bartolome et al. 2007.

Domesticated livestock have been significant controllers of California’s grassland dynamics and economics since the 1770s on the coast and the 1820s inland (Jackson and Bartolome 2007; see Chapter 37, “Range Ecosystems”). While most livestock in California are cattle (approximately 5 million), they also include sheep (0.5 million), goats, pigs, and horses (Jackson and Bartolome 2007). Livestock impact grasslands in a number of ways. The first is consumption of live plant material. Precise timing of grazing has been used to control weeds, with livestock consuming the weedy species before it is able to produce viable seeds (Huntsinger et al. 2007). 456  Ecosystems

A ­second mechanism driving grazing impacts is the accumulation of thatch or residual dry matter (RDM), the amount of senesced material remaining before the start of a new growing season (Bartolome et al. 2007). High RDM causes shading and lower temperatures, which can suppress new plants by decreasing seed germination and seedling growth (Figure 23.6, Autumn). Increasing RDM can decrease species richness, forbs, and legumes (Figure 23.7); and increase tall grasses such as wild oats and ripgut brome (Bromus diandrus) and other large-seeded species (Bartolome et al. 2007, Corbin et al. 2007a, Amatangelo et al. 2008). High RDM also decreases root to shoot allocation (Betts 2003), potentially affecting ecosystem processes such as erosion control, water dynamics, and carbon and nitrogen cycling. However, some amount of RDM benefits grasslands by increasing germination and production (see Figure 23.6) and controlling erosion (Bartolome et al. 2002, Corbin et al. 2007a). In sites with more than 38 centimeters of rainfall per year, RDM affects biomass production (Bartolome et al. 2002), with aboveground production generally highest at intermediate amounts of RDM (Amatangelo et al. 2008). The ideal amount of RDM varies by climate and topography, with higher RDM levels recommended at wetter sites and on steeper slopes (Bartolome et al. 2002). RDM levels are achieved through consumption of live plant tissue but also through consumption and trampling of senesced tissues, so early fall grazing can mitigate initially high RDM. Grazing effects depend on livestock species as well as grazing timing, intensity, duration, and frequency (see Figure 23.1) (see Chapter 37, “Range Ecosystems”). Grazer impacts also vary through interactions with environmental conditions (climate, soil, elevation, slope/aspect, land use history) and initial plant community composition (Huntsinger et al. 2007). A meta-analysis of grazing impacts across diverse soils and precipitation conditions in California emphasized the context-dependent effects of grazing on plant communities (Stahlheber and D’Antonio 2013). For example, grazing effects on non-native forbs vary across a precipitation gradient (but could not be separated by interior versus coastal grasslands), strongly increasing non-native forb cover in dry sites and decreasing it at wetter sites. Grazing increased native forb cover in interior grasslands but reduced it in coastal grasslands (Stahlheber and D’Antonio 2013). Relative cover of non-native and native grasses more strongly reflects season of grazing and more weakly responds to site conditions. Wet-season grazing enhances native grasses (particularly at dry sites) while decreasing non-native grasses. On average, grazing in California’s grasslands increases non-native forb cover (but not richness), increases native forb richness (with little change in cover), increases non-native grass richness (with little change in cover), and increases native grass cover (Figure 23.8) (Stahlheber and D’Antonio 2013). The prevalence of case studies that contradict these trends, however, highlights the need for site-specific management guidance (see Chapter 37, “Range Ecosystems”). For example, in one coastal grassland, grazing increased native forb prevalence (Hayes and Holl 2003). In another case study, grazing decreased grass cover, increased forb cover, and had no effect on species richness and little effect on natives (Skaer et al. 2013). While grazing exclosures have been suggested as a tool to increase native vegetation, decades of livestock exclosure have inconsistent effects across sites (D’Antonio et al. 2006). Grazing also alters soil properties. Moderate to high grazing (especially in the wet season) can increase soil bulk

Native forbs Native grasses –0.22

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Effect size FIGURE 23.8 Effects of grazing on percentage of cover of exotic forbs, exotic grasses, native forbs, and native grasses, as determined by meta-analysis. Positive effect sizes indicate positive effects of grazing, and negative effect sizes indicate negative effects of grazing. Error bars indicate 95% confidence intervals. Source: Stahlheber and D’Antonio 2013.

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Small mammals are generally abundant in California’s grasslands (Lidicker 1989), with varying population numbers and community dominance across sites and years (Pearson 1963, Lidicker 1989, Hobbs and Mooney 1985). Key small mammals include pocket gophers (Thomomys bottae), ground squirrels (Spermophilus beecheyi), mice (Reithrodontomys megalotis, Peromyscus maniculatus, Mus musculus), voles (Microtus californicus), moles (Scapanus spp.), rabbits (Sylvilagus spp., Lepus californicus), and in some regions kangaroo rats (Dipodomys heermannii) (Lidicker 1989, Schiffman 2007). Small mammals act as herbivores, granivores, and seed dispersers. Some species also cause significant soil disturbance (Schiffman 2007). Small mammals can reduce plant biomass through substantial herbivory and granivory (Bartolome et al. 2007). For example, in the San Joaquin Experimental Range, gophers, squirrels, and kangaroo rats consumed at least 33% of annual aboveground production (Fitch and Bentley 1949). In a grassland in the coastal hills, removal of small mammals increased aboveground biomass by 40% to 87%, partly by increasing grass abundance (Figure 23.9) (Peters 2007). During population peaks small mammals can consume up to 93% of the annual seed crop (Pearson 1964), and herbivory of live plants can decrease seed production by up to 70% (Batzli and Pitelka 1970). Small mammals can strongly alter plant community composition (Hobbs and Mooney 1991, Bartolome et al. 2007, Cushman 2007, Keeler-Wolf 2007). Seed predation can range from 0% to 75% of seed production of preferred species (e.g., wild oats) (Marshall and Jain 1970, Borchert and Jain 1978), substantially shifting plant dominance. Density of preferred seed species can decline by 30% to 62%; the resulting competitive release can increase growth and fecundity of nonpreferred plant species (Borchert and Jain 1978). Voles and mice decrease purple needle grass density, likely through granivory (Orrock et al. 2008). Similarly, squirrels and rabbits decrease purple needle grass establishment by 52%, recruitment by 30%, and reproduction by 43%. These effects are

Exotic grasses

stropods ga

SMALL MAMMALS

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With ma

­density, which can in turn reduce water infiltration (Jackson and Bartolome 2007). Consumption of plant material can short-circuit the decomposition cycle, increasing the speed of nutrient release from plants and often concentrating nutrients in areas where animals congregate (e.g., under shade trees) (Jackson and Bartolome 2007). While the grazing effects on soil nutrients vary, grazing in California’s grasslands generally increases soil nitrogen availability but decreases phosphorus and sulfur (Vaughn et al. 1986, Stromberg and Griffin 1996). Feral pigs, formerly domesticated livestock, can strongly influence grassland dynamics, especially as their populations increase rapidly. They disturb large soil areas at 5 to 15 centimeter depths in search of bulbs, roots, fungi, and invertebrates. This disturbance leads to shortterm decreases in plant diversity and long-term increases in non-native plants and decreases in oak seedlings (Cushman 2007). Coastal studies have shown that native perennials can reestablish in pig exclosures, but also that pigs tend to avoid disturbing established native bunchgrasses (Cushman 2007). While feral pig disturbance can alter soil processes in other systems, effects on soil nutrient availability and cycling rates have not been shown in California’s coastal grasslands (Cushman 2007).

396 g m-2

FIGURE 23.9 Effects of mammals and gastropods on aboveground biomass and composition during the 2001–​2 growing season. Source: Peters 2007.

greatest when purple needle grass is located near patches of black mustard (Brassica nigra) (Orrock et al. 2008, Orrock and Witter 2010). Pocket gopher populations can range from 26.6 to 100.8 ha-1 (Lidicker 1989), with larger populations in ungrazed than grazed sites (Stromberg and Griffin 1996). Gophers consume approximately 8% of aboveground biomass (Lidicker 1989) and likely more belowground, since they specialize on roots and bulbs (Lidicker 1989). They preferentially feed on patches of legumes, forbs (Hunt 1992, Eviner and Chapin 2005), and geophytes (plants with storage organs that are underground) (Seabloom and Richards 2003, Schiffman 2007) when these are available. Despite these preferences, their diets often mirror dominant plant species composition such as annual nonnative grasses and forbs (Hobbs and Mooney 1985, Schiffman 2007). Gophers disturb 1% to 30% of the soil surface each year, and on average any given soil surface turns over every three to fifteen years (Hobbs and Mooney 1985, Bartolome et al. 2007). Mounds are preferentially built in patches Gr asslands  457

with high soil shear strength, usually in grass patches with high root surface area (Eviner and Chapin 2005). This soil disturbance can strongly alter the plant community, but which plant species colonize mounds can vary year to year (Hobbs and Mooney 1985). In general, gophers increase the prevalence of forbs (Tyler et al. 2007) and annual grasses while reducing perennial grasses (Bartolome et al. 2007). Burrowing is also a significant activity of ground squirrels, kangaroo rats, mice, voles, and moles. This burrowing can increase plant diversity, especially of native forbs, and can increase prevalence of non-native species adapted to disturbance. Native perennial grasses often decrease (Schiffman 2007). Burrowing can reduce soil bulk density and increase soil temperatures and short-term pools of soil nitrogen (Eviner and Chapin 2005, Canals et al. 2003, Bartolome et al. 2007). Ground squirrel densities can range from 4.2 to 45.2 ha-1 (Lidicker 1989) and tend to increase with livestock grazing (see Chapter 37, “Range Ecosystems”). Ground squirrels form extensive burrows (Bartolome et al. 2007). They directly consume 3–​7% of net primary production and harvest (but do not consume) up to an additional 16.8% of standing biomass (Lidicker 1989). Their most important food items include several forbs (filaree, tarweed, buckwheat (Eriogonum spp.), popcorn flower (Plagiobothrys spp.), ripgut brome seeds, and acorns (Fitch 1948). Their extensive burrows provide habitat for burrowing owls and many other animals (Lidicker 1989). Like ground squirrels, kangaroo rats also tend to increase in population size under grazed conditions (see Chapter 37, “Range Ecosystems”). Their burrows increase non-native annuals and decrease natives (Schiffman 1994), and they can cache high quantities of seeds (Lidicker 1989). Kangaroo rats remove up to 16% of standing biomass (Fitch and Bentley 1949), although much of this plant harvest is associated with building runways and nests and not direct consumption. They are primarily granivores (Schiffman 2007) and can consume up to 95% of their preferred seed species, red stemmed filaree (Soholt 1973). Voles have dramatic population variation, with densities ranging from 0.25 to 1,110 ha-1 (Lidicker 1989). Their densities tend to decline with increased grazing (Bartolome et al. 2007), and they can be absent from heavily grazed sites (Lidicker 1989). At high densities they can harvest 61.4% of grassland productivity (Lidicker 1989), resulting in 50% to 85% decreased cover of their preferred food species (wild oats, ripgut brome, Italian ryegrass [Festuca perennis]) and 70% decreased seed fall (Batzli and Pitelka 1970, Batzli and Pitelka 1971). Their activity can both increase (Fehmi and Bartolome 2002, Bartolome et al. 2007) and decrease (Lidicker 1989) plant species richness. Voles selectively feed on legumes, some grasses, and some forbs (e.g., filaree [Rice 1987], yarrow [Achillea], figwort [Scrophularia], wild lettuce, clover [Medicago], dock [Rumex]), and some of these species are absent from areas with vole activity (Lidicker 1989). Their selective granivory can alter plant community composition (Cockburn and Lidicker 1983), decreasing wild oats while increasing other annuals such as ripgut brome, Italian ryegrass, and foxtail barley (Hordeum murinum) (Borchert and Jain 1978). Rabbits are abundant in California’s grasslands and intensively use this system (Zedler and Black 1992). Like other small mammals, during population outbreaks they can consume great quantities of biomass and alter community composition (Vivrette and Muller 1977).

458  Ecosystems

INSECTS, ANNELIDS, AND GASTROPODS

California’s grasslands host a diverse and abundant insect fauna. Arthropod biomass has been measured at 126 g m-2 belowground and 1.2g/m-2 aboveground (Burdic et al. 1979, Heady et al. 1992). The roles of many insects have not been studied in this system, but some are known to impact structure and function. Ants are seed consumers and dispersers. Most studies on their effects have taken place in serpentine grasslands, where they have a density of one ant mound per 100 m 2 with feeding paths 10–​12 meters long (Hobbs and Mooney 1985). Some studies have shown that ant foraging alters the spatial distribution (Peters et al. 2005) and composition (Hobbs 1985) of plants (Hobbs 1985), but others studies have seen no impact of ant seed dispersal on plant communities (Brown and Human 1997). The selectivity of ants likely varies by year, with low selectivity in dry years with low food availability (Peters et al. 2005). Ant preference for certain plants also changes throughout the season as initially preferred plant seeds are consumed (Hobbs 1985). Ant mounds, though they make up a small area of grassland (approximately 0.6% [Hobbs 1985]), consistently have been found to increase legumes and non-native annual grasses (Peters et al. 2005), while decreasing forbs and enhancing seed production of other species (e.g., peppergrass [Lepidium nitidum]) (Brown and Human 1997). Ant mounds also increase soil bacteria, fungi, microarthropods, and nematodes (Boulton et al. 2003) as well as soil organic matter and nutrients such as phosphorus, potassium, and nitrogen (Beattie 1989, Boulton et al. 2003). Grasshoppers in California’s grasslands, though little studied, strongly affect vegetation composition and standing biomass in other grasslands, consuming as much as 25–​92% of standing vegetation (Joern 1989). California hosts almost two hundred species of grasshoppers that feed on grasses and forbs, making it highly likely that they are key players in this system. In a native California perennial bunchgrass stand, grasshopper density averaged 2.3 m-2 (June through August) with an annual consumption rate of 140 kg ha-1, large enough to cause economic forage losses (Porter et al. 1996). Most grasshoppers in California reach maturity in late spring and summer, so they have little effect on annual grasses, which have largely senesced by this time (Porter et al. 1996, Joern 1989). Thus grasshoppers in California consume more native than non-native grasses, although their most abundant food item is forbs (Porter and Redak 1997). Care must be taken in generalizing the impacts of these few studies, since they focus on one grasshopper species, and different grasshopper species are frequently associated with different plant species (Stroehecker et al. 1968). Gastropods (e.g., slugs, snails) can consume high amounts of aboveground biomass, and their exclusion can increase aboveground biomass 28–​71% (see Figure 23.9) (Peters 2007). They strongly prefer certain plant species and can affect seedling survival (Peters et al. 2006, Strauss et al. 2009, Motheral and Orrock 2010). Their selectivity varies by season, with higher consumption of grasses in fall (leading to higher legume and forb cover) but higher forb consumption in winter (leading to higher grass cover). By spring, gastropod presence increases grass cover at the expense of forbs (Peters 2007). The impacts of gastropods on the plant community can be so great that their feeding behavior mediates about half of the changes in plant community composition seen in response to experimental global changes (Peters et al. 2006).

Earthworms are also important players in California’s grasslands. Earthworms stimulate litter mass loss rates by breaking up litter and incorporating it into the soil. Earthworm burrowing and casting also increase water infiltration and aeration by increasing macropores (Standiford et al. 2013). Both native and non-native earthworms occur in California grasslands, with non-native earthworms dominating disturbed and fertile environments and natives dominating relatively undisturbed grasslands (Winsome et al. 2006). The non-native earthworms are more active than natives, leading to greater physical disturbance of the soil. Through this increased activity, non-native earthworms increase plant growth and uptake, enhance N turnover through litter decomposition, and decrease microbial biomass (Winsome 2003).

BIRDS

California’s grasslands are primary habitat for some bird species and provide feeding and/or nesting grounds for other species. Their use can be seasonal or year-long (reviewed in Lidicker 1989, CPIF 2000, Shuford and Gardali 2008). While few studies document the ecological impacts of these birds on California’s grasslands, there are some critical roles played by grassland birds in general. Birds can have substantial impacts on plant populations and species composition through seed dispersal and granivory, with effects that are distinct from those of granivorous small mammals (reviewed in Espeland et al. 2005). Many of the same bird species are also important insectivores, controlling populations of grasshoppers and other potential pest insects, and sometimes also disturbing soil to feed on insects, grubs, and worms (Fix and Bezener 2000, Sekercioglu 2006). Examples of birds in California’s grasslands that are both granivores and insectivores include savannah sparrow (Passerculus sandwhichensis), grasshopper sparrow (Ammondramus savannarum), horned lark (Eremophila alpestris), western meadowlark (Sturnella neglecta), vesper sparrow (Pooecetes gramineus), and lark sparrow (Chondestes grammacus) (Lidicker 1989, Fix and Bezener 2000, see both references for a more extensive list). The relative importance of seeds versus insects in bird diets often vary by season and by species (Shuford and Gardali 2008). Many of these birds nest on the ground from early spring through July, so their breeding can be disrupted by mowing, grazing, disking, or burning during the spring (CPIF 2000). Their populations have been steadily declining, at least partly due to loss of continuous grassland habitat (CPIF 2000, Rao et al. 2008). Predaceous birds can have significant effects on grassland structure and function through their controls over the populations of small mammals. In order to avoid predation, in the presence of birds, small mammals alter their behavior and habitat use, leading to less use of areas with short or sparse vegetation (Sekercioglu 2006). Most predatory birds also feed on smaller birds, amphibians, reptiles, large insects, and sometimes carrion (Fix and Bezener 2000). Key avian predators in California’s grassland include hawks (red-tailed [Buteo jamaicensis], ferrunginous [Buteo regalis], Swainson’s [Buteo swainsoni], northern harrier [Circus cyaneus]), owls (burrowing [Athene cunicularia], short-eared [Asio flammeus]), and the white-tailed kite (Elanus leucurus) (Lidicker 1989, CPIF 2000, Shuford and Gardali 2008).

NONAVIAN PREDATORS

Since small mammals and insects can have such large effects on California grassland structure and function, regulation of these groups by predation has significant impacts on these grasslands (Schiffman 2007). Predatory animals are diverse, including birds, snakes, coyote (Canis latrans), fox (Vulpes fulva, Urocyon cinereoargenteus), badger (Taxidea taxus), alligator lizards (Elgaria spp.), and the domesticated/feral cats (Felis domesticus) (Lidicker 1989). These species can have diverse diets, including insects, birds, bird eggs, small mammals, and in some cases, one another (Fix and Bezener 2000, Jameson and Peeters 2004, Stebbins and McGinnis 2012). Important insectivores include the Pacific tree frog (Pseudacris regilla), tiger salamander (Ambystoma spp.), skinks (Eumeces spp.), and a variety of lizards (western fence lizard [Sceloporus occidentalis], coast horned lizard [Phrynosoma blainvillii]) (Stebbins and McGinnis 2012). Omnivores are also common, eating a wide variety of plant species and tissues, as well as insects, earthworms, amphibians, reptiles, and small mammals (e.g., skunk [Mephitis mephitis], raccoon [Procyon lotor]) (Lidicker 1989, Jameson and Peeters 2004). As described earlier, the small mammals with the largest impacts on grasslands include ground squirrels, gophers, and voles. The primary predators of ground squirrels include raptors (e.g., red-tailed hawk) and the western rattlesnake (Crotalus viridis). White-tailed kites, gopher snakes (Pituophis melanoleucus), and garter snakes (Thamnophis sp.) are considered the most important predators of voles. Key predators of gophers include gopher snakes, western rattlesnake, red-tailed hawk, barn owl (Tyto alba), great-horned owl (Bubo virginianus), and coyote (Lidicker 1989).

MICROBE S

California’s grasslands have high soil microbial biomass and richness (Sanchez-Moreno et al. 2011), with a dynamic microbial community that changes in response to plant communities (Hawkes et al. 2005, Batten et al. 2006), temperature, and moisture (Waldrop and Firestone 2006). Microbial community shifts can have important effects on plant communities and ecosystem processes. Seasonal shifts in microbial communities due to temperature and moisture lead to shifts in the soil enzymes that mediate decomposition and nutrient cycling, with many enzymes peaking in the early spring and/or winter and least active in the summer (Waldrop and Firestone 2006). Dry conditions can decrease bacterial biomass and can decrease (Alster et al. 2013) or increase enzyme activity (Henry et al. 2005). Higher spring precipitation reduces the abundance and diversity of fungi and increases decomposition rates (Hawkes et al. 2011). Microbial communities can directly affect plant performance, altering plant growth rate and root-toshoot allocation, with effects varying by plant species (Brandt et al. 2009). Plant communities also can shape microbial communities. For example, non-native grasses have increased the population size and altered the composition of the ammonium oxidizer community, leading to more than doubled rates of nitrification over native grass soils (Hawkes et al. 2005). Increased nitrification can have strong effects on plant nitrogen availability, nitrogen retention, and water quality. Arbuscular mycorrhizae (AM) are fungal symbionts with plants, exchanging plant carbon for various resources includ-

Gr asslands  459

ing nitrogen, phosphorus, and/or water. Most grassland plants are mutualistic with these fungi (Hopkins 1987, Harrison and Viers 2007), and the composition of the AM community can alter plant growth and seed production, nutrient uptake, root-to-shoot allocation, and drought stress tolerance (Allen and Allen 1990, Nelson and Allen 1993, Harrison and Viers 2007). AM also strongly enhance soil aggregate formation, which can affect carbon and nutrient dynamics, soil water infiltration and storage, and erosion control (Rillig et al. 2002). AM in California’s grasslands play a particularly important role in plant phosphorus uptake. In the presence of AM, plant production is nitrogen-limited, but without the AM symbiosis, plants are limited by phosphorus (Grogan and Chapin 2000). AM hyphal networks can associate with many individual plants simultaneously, leading to transfers of phosphorus (and possibly other resources) among diverse plant species. For example, when radioactive phosphorus was added to a given plant, that phosphorus was transferred to 20% of the plant’s close neighbors through the AM network (Chiariello et al. 1982). This AM network among plant species can influence dynamics between native and non-native plants. For example, in the presence of the AM community the non-native Napa star thistle (Centaurea melitensis) dominated over native purple needle grass. However, when AM biomass was reduced, the non-native plant was much less competitive (Callaway et al. 2003), suggesting that the AM network provided the non-native plant with resources from the native plants. AM communities change in response to environmental conditions, with much change not in direct response to environmental changes but mediated through vegetation changes (Rillig et al. 1998). Because AM species have plant speciesspecific effects, vegetation-induced changes in AM community composition can alter plant competitive outcomes (Allen and Allen 1990). A number of studies have shown that the AM community differs under native and non-native plants (Hawkes et al. 2006, Nelson and Allen 1993, Vogelsang and Bever 2009), with non-native plants exerting stronger effects on AM than native plants do (Vogelsang and Bever 2009). Non-native plants alter not only the soil AM community but also that associated with native plants. For example, when wild oats and purple needle grass grow as neighbors, the AM associated with wild oats dominated purple needle grass roots but purple needle grass did not affect the AM on wild oat roots (Hausmann and Hawkes 2009). This effect was particularly strong when wild oats established before purple needle grass (Hausmann and Hawkes 2010). In another study nonnatives and natives both grew best associated with their own AM communities (Vogelsang and Bever 2009). Finally, nonnative plant effects on the AM community increased the seed production of non-native plants but not native plants (Nelson and Allen 1993). Dynamics between non-native and native plant species can also be mediated by microbial and viral pathogens. Crown rust can decrease wild oats while increasing purple needle grass (Carsten et al. 2001). In contrast, barley yellow dwarf virus and cereal dwarf virus can negatively affect both native and non-native grasses, but they have a stronger negative effect on natives, particularly because the non-native annual grasses enhance transmission of the viruses to natives (Malmstrom 1998). When exposed to these viruses and to competition with non-native annuals, first-year survivorship of natives can be halved (Malmstrom et al. 2006), with other studies showing the viruses can decrease native sur460  Ecosystems

vival 0–​80% and fecundity 30–​70% (Borer et al. 2007). Grazing can interact with these viruses, but overall impacts are not unclear, with studies showing that vertebrate herbivores can increase plant infection by viruses (Borer et al. 2009) but that survivorship of the infected plants can increase (Malmstrom et al. 2006). The soil food web is an important mediator of biogeochemical processes. While only a limited number of studies have addressed it in California’s grasslands, we know that abundance and richness of groups such as protozoa and nematodes are high but vary greatly across sites, seasons, and years (Freckman et al. 1979, Heady et al. 1992, SanchezMoreno et al. 2011, Baty 2012). Processes such as litter decomposition are strongly controlled by the size and composition of the food web, which is in turn controlled by both resource availability and predation (Barstow 2011).

Interacting Factors: Transition of California’s Grasslands to a Non-Native-Dominated State Frequent interactions among biotic and abiotic factors determine the structure and function of California’s grasslands (see Figure 23.1). For example, the effects of gopher mounds on plant composition differ with precipitation (Hobbs and Mooney 1991), as the effects of burning on plant communities vary with grazing regime (D’Antonio et al. 2006). The interplay of multiple factors is perhaps best demonstrated by a suite of hypotheses about the causes of non-native plant domination in California’s grasslands. The composition of California’s grasslands at the time of European settlement is not well documented (Wigand et al. 2007); it is unclear whether the currently common native species were previous dominants or were historically unusual species able to survive changing conditions. There has been substantial debate about the pre-European composition of this system, with theories ranging from: (1) it was dominated by native perennial bunchgrasses interspersed with native forbs, and replacement by exotic annuals was due to overgrazing and drought; versus (2) it was dominated by wildflowers (both annual and perennial forbs), which declined due to the competitive nature of the newly introduced exotic grasses and forbs (reviewed in Minnich 2008). While there are strong advocates for both of these alternatives, it is generally accepted that native systems likely contained perennial bunchgrasses and forbs, rhizomatous grasses, and annual forbs and grasses, with different plant groups dominating different regions (Bartolome et al. 2007). Perennial grasses likely dominated wetter areas, such as those adjacent to the coast, the windward aspect of the coast range, and wetter areas of the Central Valley. Annual forbs likely were present in all of California’s grasslands but dominated in drier areas, including the foothills, the interior coast ranges, and in the drier areas of the Central Valley (D’Antonio et al. 2007). The abundance of forbs on most sites were likely to annually vary from rare to abundant, depending on weather and disturbance regimes (Schiffman 2007). Using the term “grasslands” to describe this diverse group of communities can underplay the current and historical importance and prevalence of forbs, and many advocate returning to the term “prairie,” which was historically used to describe these diverse systems in California (Holstein 2011). While we can only speculate about the composition of historical plant communities, we know that non-native species replaced the dominant native vegetation during the 1700s

and 1800s (Bartolome et al. 2007, D’Antonio et al. 2007). Non-native species invasions occurred in a number of waves (Heady et al. 1992, D’Antonio et al. 2007, Bossard and Randall 2007, Minnich 2008). Species such as wild oats, filaree, and mustard (Brassica) were found in adobe bricks of early Spanish missions, indicating their prevalence even before the mid1800s, when European settlements and livestock expanded. Bromes (Bromus spp.) and barleys/foxtails (Hordeum spp.) spread in the 1860s and 1870s. In the late 1800s hairgrass (Aira), foxtail brome, and Napa starthistle invaded. Species currently invading California’s grasslands include barbed goatgrass, medusa head, and yellow starthistle (Centaurea solstitialis) (Heady et al. 1992, D’Antonio et al. 2006, D’Antonio et al. 2007, Keeler-Wolf et al. 2007) (see Figure 23.4c); coastal grasslands are also currently being invaded by non-native perennial grasses, including velvet grass (Holcus lanatus), tall fescue (Festuca arundinacea), Harding grass (Phalaris aquatica), and orchard grass (Dactylis glomerata) (Corbin and D’Antonio 2010). California’s grasslands now contain four hundred nonnative plant species (Bartolome et al. 2007), amounting to 37% of California’s invasive flora—​t he largest of any ecosystem in the state (Bossard and Randall 2007). A number of hypotheses, all with strong experimental support, address what caused the widespread invasion of nonnative species. Most of these hypotheses focus on transition from native perennial bunchgrasses (Bartolome et al. 2007, D’Antonio et al. 2007), which will be the focus of the next discussion. In areas that were not dominated by native perennial grasses, other mechanisms (e.g., superior competitor ability of invaders) may have driven the transition from native to exotic domination (Minnich 2008). Many have argued that the domination of non-native species resulted primarily from the competitive superiority of non-natives due to their rapid early-season growth, drought tolerance, high seed production, and earlier seed establishment (Bartolome and Gemmill 1981, D’Antonio et al. 2007). However, competition on its own was likely not enough to drive such dramatic shifts from native perennial bunchgrasses. Well-established stands of native perennial bunchgrasses resist invasion. Even though newly established native grasses can be initially invaded by non-natives, in a number of cases natives persisted in these invaded patches and eventually suppressed non-native grasses (Corbin and D’Antonio 2004, D’Antonio et al. 2007, Eviner et al. 2013). Thus the extensive transition from native to non-native domination that took place likely also required a stressor that decreased the performance or cover of natives if the previous dominants were perennial grasses. Drought and overgrazing are the most commonly hypothesized such stressors driving the native to non-native transition (D’Antonio et al. 2007). While native grasses might have evolved under seasonal grazing and browsing, in the 1700s they were exposed to heavy, year-round livestock grazing. This could have exceeded the grazing tolerance of natives, while Mediterranean non-natives could tolerate it (Bartolome et al. 2007, Hille Ris Lambers et al. 2010). Overgrazing also decreased productivity through erosion and loss of soil fertility (Allen-Diaz et al. 2007). Moreover, the native grassland plants had evolved under wetter, longer growing seasons (Dyer 2007), and were likely particularly hard hit by severe, multiyear droughts in 1850–​1851 and 1862–​1864. Annuals cope with the prolonged dry season by producing seeds at the onset of summer and dying. This strategy might have allowed them to establish under low-rainfall conditions and to be poised to spread when the droughts ended (Reever Morghan et al. 2007).

In addition to drought and overgrazing, a number of other well-supported mechanisms might have contributed to invasion by non-native species. Increased settlement by Europeans across California wrought major changes to hydrology and fire regimes. River damming and levees destroyed many fertile, moist floodplains, decreasing the water and silt deposition that supported rich vegetation (Dyer 2007). Native Americans also managed grasslands through high-frequency burns, and the cessation of these burns may have increased non-natives (Dyer 2007, Bartolome et al. 2007). Another key land management change was the rise of crop agriculture, with extensive tilling that the native perennials could not survive (D’Antonio et al. 2007). In addition to land use changes, extensive biotic interactions could have contributed to the vegetation transition. In the late 1800s to early 1900s, increased hunting pressure on predators led to extremely high abundance of small mammals. Their extensive soil disturbance might have favored annuals, with their high seed production and ability to establish quickly on disturbed areas (Schiffman 2007). An already stressed native community could have been further decimated by grasshopper outbreaks. Because these outbreaks occur during summer, they would negatively affect perennials but have no effect on non-native annuals, which are already dead at this time of year (Joern 1989). Once the transition to annual non-natives occurred, a number of mechanisms could have maintained the invaded state. Barley and cereal yellow dwarf viruses decrease native grass growth, survivorship, and fecundity (Malmstrom et al. 2005a, 2005b). The presence of non-native grasses more than doubles infection of native grasses by these viruses (Malmstrom et al. 2005a, 2005b), partly by increasing abundance of aphid—​the vector of these pathogens (Borer et al. 2009). Nonnative grasses also alter soil chemistry and microbial communities, which can feed back to favor non-natives over native plants (Grmn and Suding 2010, Hausmann and Hawkes 2010). Finally, the decline of native grasses might have caused widespread seed limitation, preventing natives from reestablishing on their own (Hamilton et al. 1999, Seabloom et al. 2003).

Ecosystem Functioning This review of ecosystem function focuses mostly on the annual grassland, since it is the dominant type in California. As discussed earlier, in most of the world’s grasslands dominance of annual species is limited to early successional stages. Most paradigms for understanding and managing grasslands thus focus on perennial grasslands. These frameworks are not adequate for understanding annual-dominated systems, where the annual growth habit, coupled with high interannual variability in precipitation, strongly influence functioning and management needs (Heady et al. 1992, Bartolome et al. 2007).

Net Primary Production (NPP) Timing of plant production is driven by seasonality of temperature and moisture (see Figure 23.3), with an initial pulse of production early in the season when temperature and moisture are both ideal. When the first rains occur during colder periods, lower temperatures do not inhibit stand establishment but can limit growth (Evans and Young 1989). In Gr asslands  461

the winter, low temperatures inhibit aboveground biomass growth but root growth continues, reaching its peak before mid-March (Evans and Young 1989, Heady et al. 1992). When midwinter droughts occur, water can limit growth during the rainy season (Corbin et al. 2007a). As temperatures rise in mid- to late February, aboveground NPP increases, with peak biomass often occurring in mid-April to late May, just before the soil dries and plants begin to senesce (Evans and Young 1989, Heady et al. 1992). The high variability in annual weather coupled with the annual growth form of the dominant plants renders NPP in California’s grasslands extremely variable, with typical annual variations of at least 50% of mean NPP (Bartolome et al. 2007). For example, from 1935 through 1999 at the San Joaquin Experimental Range, aboveground NPP ranged from 1,008 to 5,040 kg ha-1, while at Hopland Research and Extension Center, aboveground production ranged from 1,008 to 3,920 kg ha-1 from 1953 through 1999 (George et al. 2001). Although California grasslands have been studied intensively for decades, limited ability persists to explain this variability in production using the conventional predictors of climate, soil type and residual dry matter (RDM) (George et al. 2001). NPP only weakly relates to total annual precipitation and is much more strongly affected by timing of rainfall during adequate temperatures for growth (see Figure 23.5) (Pitt and Heady 1978, George et al. 2001, Reever Morghan et al. 2007, Suttle et al. 2007). In general, NPP tends to be highest in years with high and steady rainfall in November through February (Murphy 1970, Pitt and Heady 1978, Reever Morghan et al. 2007, Chou et al. 2008), particularly when temperatures are higher in this period (Pitt and Heady 1978, George et al. 2001). However, this generalization does not always hold—​ even in longterm datasets, the timing and total amount of precipitation do not always correlate with production (Pitt 1975, Duncan and Woodmansee 1975). Moreover, different sites respond uniquely to timing of rainfall. Sites in northern California’s coastal range and foothills have their highest NPP when fall and winter are warm and wet. In contrast, a drier southern California site has its highest NPP in years with high spring precipitation (George et al. 2001). Plant community composition can strongly shape the impacts of spring rains on NPP. When vegetation at a site is dominated by species that senesce early to midspring, spring rains (March, April) either have no effect or decrease NPP (Pitt and Heady 1978, Reever Morghan et al. 2007, Chou et al. 2008). However, late-season rains increase NPP at sites with late-season species, particularly in sites with summer annuals, which can produce up to 10% of a site’s NPP. In these cases, the duration of the rainy season can determine the duration of the growing season (Hooper and Heady 1970, Chiariello 1989). In perennial-dominated grasslands along the coast, the duration of the growing season is also extended by moisture supplied through fog inputs (Corbin et al. 2005). While weather conditions are likely the strongest controls over production (Corbin et al. 2007a, Bartolome et al. 2007), there are a number of other factors that also play an important role, within the constraints of weather patterns. When moisture is not limiting to growth, soil nutrients are the next-most limiting factor (Pitt and Heady 1979, Harpole et al. 2007). Nitrogen is the most commonly limiting nutrient to plant growth in this system, but NPP can also be limited by phosphorus or sulfur, depending on the site and the vegetation community. Some sites can respond equally to 462  Ecosystems

nitrogen additions versus sulfur and phosphorus additions. Nitrogen additions will enhance grass production, while sulfur and phosphorus will stimulate legume production, if they are present (Jones et al. 1970; Jones and Martin 1964; Jones et al. 1983). Fertilization in the fall is particularly effective in increasing NPP (Jones 1974). Plant production from the previous year can also impact NPP. In sites with greater than 400 millimeters of rainfall per year, maximum production is associated with intermediate residual dry matter (RDM) levels of 840 kilograms per hectare. Too much RDM can suppress production the following year through shading, and too little can decrease production, presumably due to the loss of RDM’s roles in microclimate mitigation, nutrient provision, and water infiltration (Bartolome et al. 2007). Seed production from the previous year, and seedling dynamics also have strong impacts on NPP. Annual plants in California translocate 63–​77% of their aboveground nitrogen to seeds (Woodmansee and Duncan 1980), over 90% of these seeds germinate at the start of the growing season, and up to 50% of germinated seedlings can die within the first six to eight weeks of the growing season. After this initial pulse of thinning, seedling death proceeds steadily throughout the growing season, so that 75–​9 0% of seedlings die throughout the growing season (Bartolome 1979, Young et al. 1981). Seedling thinning results in inputs of very labile litter with low structural material, leading to rapid availability of seedling nitrogen to other plants. Self-thinning acts as a perfectly timed slow-release fertilizer, with release of highly labile nutrients at the time of peak plant competition (Eviner and Firestone 2007). Manipulations of seed density show that seedling thinning can double the NPP compared to planting at seed densities that are too low for thinning to occur (Eviner et al. in prep.). Similarly, increased seed density enhances aboveground productivity in other grasslands (Turnbull et al. 2000, Moles and Westoby 2006). In fact, productivity is often so enhanced in high-density stands, that fertilizer additions cause little if any increase in productivity (while fertilizer does increase growth at low density) (Bolland 1995, Thompson and Stout 1996, Eviner et al. in prep.). Seedling thinning likely plays a role in regulating the annual variability in NPP, since these grasslands experience dramatic variations in seed production (four- to onehundred-fold variation), seedling numbers (two- to sixfold variation), and self-thinning (one- to fivefold variation) from year to year at a given site and across sites within a given year (Heady 1958, Bartolome 1979, Young et al. 1981). Through effects on seed production, weather patterns in a given growing season may impact productivity of the following growing season. For example, winter droughts and low spring precipitation can greatly decrease seed production (Heady et al. 1992), which may lead to lower production the following growing season.

Decomposition Breakdown of litter is critical for nutrient recycling, and for regulating excess thatch accumulation. In California grasslands, root litter typically decomposes within a year, while aboveground litter takes two to two and a half years to turn over (Savelle 1977). There are a number of key controllers of decomposition rates in grasslands. Plant senescence creates litter in the late spring, when moisture conditions are not conducive to microbial activity. Despite this, grassland litter

tends to lose 8–​10% of its mass and 20% of its lignin during the first summer, due to photodegradation. After one year, litter exposed to sunlight in the summer has double the mass loss compared with litter that is shaded (Henry et al. 2008). As litter layers thicken, due to increased production, photodegradation effects on mass loss remain constant, but lignin breakdown decreases substantially (Henry et al. 2008). This photodegradation can have significant impacts on ecosystem carbon dynamics in the early fall. The first significant rains usually induce a 5% litter mass loss through leaching (Savelle 1977) and a respiration pulse that can be responsible for up to 10% of ecosystem carbon loss a year. These losses likely depend on photodegradation (Ma et al. 2012). Once the rainy season begins, soil temperature and moisture determine microbial activity and decomposition rates. The peak timing of microbial decomposition roughly coincides with that of plant growth but is a bit more buffered from low temperatures (Savelle 1977, Heady et al. 1992, Eviner and Firestone 2007). Microbial activity can also persist under lower moisture conditions than plants, because they can exploit moist microsites within the soil. Because of their short generation times, microbes can respond more readily to lateseason rains than plants, leading to pulses of decomposition after these rains (Chou et al. 2008). Drought conditions can decrease decomposition through moisture limitation but also through shifts in the microbial community that persist even when moisture conditions become ideal (Allison et al. 2013). Litter chemistry is an important determinant of the rate of mass loss and nutrient release. As in other systems, higher nitrogen content of litter leads to faster decomposition rates. An important exception to this is some of the litter from forbs, which can contain defensive compounds, such as alkaloids that can inhibit decomposition (Eviner 2001). Litter structure is another key controller of decomposition rates. Standing litter that is not in contact with the soil surface decomposes more slowly than litter in contact with the surface, while buried litter is the most quickly decomposed (Dukes and Field 2000). The more litter buildup there is, the less contact there is with the soil, leading to greater inhibition of decomposition and faster buildup of litter. However, wind and rain, as well as trampling by herbivores can drive standing litter down, enhancing decomposition. Decomposition can be accelerated when litter is incorporated into the soil through gopher activity, which can decrease standing litter by two- to eightfold (Stromberg and Griffin 1996). Macrofauna, which break up litter and incorporate it into the soil, can cause 20% mass loss (Savelle 1977, Heady et al. 1992).

Nitrogen Cycling Nitrogen is the most commonly limiting nutrient to plant growth in California’s grasslands, so its cycling can be a critical controller of NPP, as well as vegetation composition (Corbin et al. 2007a, Harpole et al. 2007). Soil organic nitrogen represents 94% of the system’s nitrogen pool (Eviner and Firestone 2007), but much of this is not readily available to plants and microbes, due to physical and chemical protection. Nitrogen becomes available to plants through litter decomposition, soil organic matter mineralization, atmospheric deposition, and nitrogen fixation by legumes (Woodmansee and Duncan 1980, Pendelton et al. 1983, Vaughn et al. 1986, Center et al. 1989, Heady et al. 1992). Additionally, 37% to 63% of annual internal nitrogen cycling is mediated through seed-

ling thinning, essentially acting as a slow-release fertilizer, providing nitrogen at peak times of plant nitrogen demand (Eviner and Firestone 2007). This is a key example of a driver of ecosystem processes that is unique to annual grasslands. Like other processes, nitrogen cycling, uptake, and loss have strong seasonal trends. At plant senescence, approximately 70% of aboveground nitrogen is stored as seeds, with the remaining as litter. A range of 1% to 75% of seeds may be consumed by granivores during the summer (Heady et al. 1992), leading to potentially high nitrogen release through granivory. Summer dynamics of aboveground litter nitrogen are variable, with some studies showing loss of 25% of aboveground litter nitrogen (and 35% of root litter nitrogen) (Jackson et al. 1988), while others show nitrogen accumulation during the summer, even as mass loss occurs through photodegradation (Henry et al. 2008). This accumulation is due to microbial immobilization of nitrogen, leading to a buildup in soil microbial biomass through the summer, with its annual peak at the end of the summer (Jackson et al. 1988). Rates of nitrogen cycling can be low after spring dry-down and before the first fall rains (Herman et al. 2003, Eviner et al. 2006), but surprisingly, microbial populations and enzyme activity can be maintained through the summer (Treseder et al. 2010, Parker and Schimel 2011), leading to sustained cycling of nitrogen and accumulation of inorganic nitrogen in the soil (Parker and Schimel 2011). Part of the reason for high inorganic nitrogen accumulation is lack of moisture to facilitate gaseous and leaching losses (Eviner and Firestone 2007) as well as generally low plant uptake during the summer, because most annual plants are senesced. However, in grasslands with high biomass of summer annuals, summer uptake can be up to 10 kg N/ha, approximately 8% of the total taken up between October and June, the typical growing season (Chiariello 1989). Fall rains stimulate nitrogen mineralization rates but have an even greater stimulatory effect on microbial immobilization and microbial biomass (Herman et al. 2003). Repeated wet-dry cycles, which are typical between fall rains, further stimulate nitrogen mineralization, microbial biomass, and microbial activity (Xiang et al. 2008). High immobilization does not prevent some nitrogen leaching loss, which is often at its peak within a few weeks after wet up (Figure 23.10) (Jones et al. 1977, Vaughn et al. 1986, Jackson et al. 1988, Davidson et al. 1990, Maron and Jeffries 2001, Lewis et al. 2006). In the winter, low temperatures cause nitrogen cycling rates to decrease, with immobilization decreasing to a greater extent than mineralization, leading to net mineralization occurring in the winter (as opposed to net immobilization in the fall) (Jones and Woodmansee 1979, Schimel et al. 1989, Davidson et al. 1990, Maron and Jeffries 2001, Herman et al. 2003, Eviner et al. 2006). Low temperatures limit plant uptake as well, leading to an increase in soil inorganic nitrogen levels (Vaughn et al. 1986, Jackson et al. 1988). Warming in early spring increases nitrogen cycling rates as well as plant and microbial uptake of nitrogen, and 82% of plant nitrogen uptake is completed by this time, even though only 45% of plant production has occurred. As the soil dries out in the spring, nitrogen cycling rates decrease (Eviner and Firestone 2007), and plant nitrogen availability may be restricted by lack of soil moisture (Everard et al. 2010). While the general seasonal trends are presented above, these seasonal patterns can vary year to year (Herman et al. 2003). Nitrogen cycling rates also vary depending on which plant species are dominant (Eviner et al. 2006, Eviner and FiresGr asslands  463

Streamflow (mm d-1)

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Date FIGURE 23.10 Seasonal fluxes of streamflow (upper) and stream nitrogen load (lower) at Sierra Foothills Research and Extension Center, by water year (October–​September) from 1981 to 2000. Source: Lewis et al. 2006.

tone 2007, Corbin and D’Antonio 2011). Soil disturbance by gophers and feral pigs have the potential to increase nitrogen availability over the short term (Canals et al. 2003, Eviner and Chapin 2005), but this is not always the case (Cushman et al. 2004, Tierney and Cushman 2006, Eviner and Firestone 2007). Grazing can vary in its effects on nutrient availability, but a number of studies have indicated that grazing California’s grasslands can increase soil nitrogen and sulfur but lower phosphorus (Jackson et al. 2007).

Water Balance Water availability is the key driver of this system’s structure and function, so that water inputs, infiltration, storage, and losses determine water supply for plants, animals, and humans (Parton and Jackson 1989, Reever Morghan et al. 2007, Salve and Torn 2011, Salve et al. 2011). While water availability partly depends on precipitation inputs, the ability of the system to capture and store water is critical to support annual peak evapotranspiration in the spring, when precipitation is low and infrequent (Ryu et al. 2008). Water capture is determined by infiltration into the soil versus runoff over the surface. Early in the season, when soils are dry, water infiltration is largely determined by soil type and porosity 464  Ecosystems

(determined by soil texture, soil channels from roots and soil fauna, and soil disturbance from organisms such as gophers). Standing vegetation and litter can decrease physical compaction of the soil by decreasing the impact of raindrops, and vegetation and litter also slow runoff, allowing more time for water to infiltrate into the soil. Early season rains wet up the soil surface, and with increasing rain, the soil gradually wets up deeper, creating a “wetting front.” The depth and speed of this wetting front depend on the magnitude and frequency of early season rains, as well as soil channels from roots and macrofauna (Salve and Torn 2011). When small precipitation events occur, with a week or more between them, the initial soil moisture at the surface is lost due to evapotranspiration, and the next rains recharge surface soil, rather than increase the depth of the wetting front. In order to recharge soil moisture below a 0.2 meter depth, substantial rainfall over successive days is needed (e.g., 66 millimeters or more), and deep soil (1.5 meter depth) is not saturated until later in the growing season, when successive storms have occurred (Salve and Tokunaga 2000). In low rainfall years, deep soil may not be recharged (Reever Morghan et al. 2007). Once the soil is recharged, water storage is determined by water-holding capacity of the soil, which is determined by soil texture, organic matter, and depth. Precipitation beyond this water-holding capacity leads to a mix of surface runoff

(due to lack of infiltration) and leaching through the soil column and into the groundwater and/or streams. Because much of the rain falls at a time of low plant growth and evapotranspiration, 19–​76% of precipitation can be lost through streamflow (Nixon and Lawless 1960, Parton and Jackson 1989, Lewis et al. 2006, Reever Morghan et al. 2007). Initial rains tend to have lower loss to streamflow, since they are still wetting up the soil column. Once the soil column has wet up (estimated at 15 to 25 centimeters of accumulated rainfall), 70% of additional rainfall is lost to the system (Dahlgren et al. 2001, Lewis et al. 2006). Most of this is lost in limited pulses throughout the growing season (see Figure 23.10). For example, in one study, moisture moved below the top 0.5 meter of soil only during five significant rain periods, amounting to the twenty-one days of the growing season that received 71% of that year’s precipitation (Salve and Tokunaga 2000). This can lead to highly variable streamflow from grasslands. Annual evapotranspiration rates are less variable than streamflow, because most evapotranspiration occurs in the spring, when precipitation is low or has stopped for the season, and thus much of the annual evapotranspiration is derived from water stored in the soil (Ryu et al. 2008, Salve et al. 2011). Plant traits can influence the timing and amount of evapotranspiration, due to their different phenologies and rooting depths. Those with deeper roots can access soil water that is unavailable to many other grassland species and tend to be active later into the dry season (Enloe et al. 2004, Reever Morghan et al. 2007). Management practices can impact water dynamics. For example, in a study of long-term (thirty to forty years) grazing treatments, compared to ungrazed sites, heavy grazing increased soil compaction and bulk density, decreasing infiltration. This led to heavily grazed plots having two- to fivefold higher runoff and a twofold reduction in water storage. The effects of light grazing were more similar to ungrazed than heavily grazed treatments (Liacos 1962). Conversion of woody systems (oak woodlands, chaparral, coastal sage scrub) to grasslands also has large impacts on water dynamics, increasing streamflow by an average of 60%, due to lower evapotranspiration rates. However, grasslands have higher infiltration rates than woody systems, leading to more gradual release of water into streams. This results in lower maximum volume of peak storm streamflow (e.g., can minimize flooding) but longer periods of stormflow (Lewis 1968, Dahlgren et al. 2001). However, there are exceptions to these patterns, with removal of woody vegetation on some sites increasing deep water storage but not affecting runoff (Veihmeyer 1953).

Ecology and Management of Ecosystem Services California grasslands have the potential to provide a number of key ecosystem services, including forage and livestock production, weed control, pollination, carbon sequestration, water supply and purification (including erosion control), fire control, recreation, and scenic vistas (FRAP 2010, Cheatum et al. 2011, Ferranto et al. 2011). In addition, maintenance and enhancement of plant and wildlife diversity is a key management goal in many restoration and conservation efforts and is important in enhancing the delivery and resilience of most ecosystem services. This section highlights successful management strategies for ecosystem services, although it is

important to keep in mind that the effects of management are constrained by environmental factors, which can have a strong impact on any of these ecosystem services (see Figure 23.1) (Jackson and Bartolome 2007). This results in strong site-specific and year-specific effects of management, requiring adaptive management approaches (see Chapter 37, “Range Ecosystems”). Another considerable challenge is balancing management for multiple goals, since all ecosystem services are desirable but rarely achievable simultaneously. Focal ecosystem service goals depend on who owns the land. Eighty-eight percent of California’s grasslands are under private ownership (FRAP 2003, Jantz et al. 2007), with 53% designated as agricultural (for grazing), 20% as open space, 22% as residential (at very low to low density), and 5% as “other” (Jantz et al. 2007). Fortunately, many private landowners manage for services other than forage production, with 50% or more managing for each of the following services: fire control, wildlife habitat, water quality, and erosion control. In addition, approximately 40% of private landowners actively remove non-native species and 40% plant native species (Ferranto et al. 2011).

Biodiversity Biodiversity is a key controller of ecosystem services and is a focal goal of many conservation and restoration efforts. Even in their invaded state, California grasslands are species rich, averaging greater than fifty plant species per 30 x 30 meter area (Heady et al. 1992). California grasslands also contain a number of rare and unique habitats, including vernal pools, serpentine grasslands, and riparian systems, which are hotspots of native diversity. California’s grasslands are critical habitat for diverse plants and animals, including many endemic, threatened, and endangered species. These grasslands provide habitat for nearly 90% of species in the Inventory of Rare and Endangered Species in California (Skinner and Pavlik 1994), and seventy-five federally listed threatened or endangered species, including: fifty-one plants, fourteen invertebrates, and ten vertebrates (Jantz et al. 2007). Examples of key species of concern include the San Joaquin kit fox (Vulpes macrotis mutica), burrowing owl (Athene cunicularia), bay checkerspot butterfly (Euphydruas editha bayensis), Swainson’s hawk (Buteo swainsoni), California tiger salamander (Ambystoma californiense), and California quail (Callipepla californica) (Barry et al. 2006, Cheatum et al. 2011). Because 88% of California grasslands are privately owned (Jantz et al. 2007), conservation and restoration of diversity largely depends on private land owners. Large ranches are critical for wildlife conservation, providing habitat and connectivity between habitats (FRAP 2010). Wildlife are a priority for many landowners, with more than 50% managing for wildlife habitat (Ferranto et al. 2011), and many employing management strategies for specific species (SRDC 2006, Barry et al. 2006). For example, moderate grazing can benefit kit foxes, which prefer grasslands with aboveground biomass less than 560 kilogram per hectare. Mixed grazing patterns benefit burrowing owls, which prefer heavily grazed areas for nesting but require areas with tall grass cover to provide habitat for voles, their preferred prey (Barry et al. 2006). The checkerspot butterfly requires the native forb, California plantain (Plantago erecta), which increases in prevalence under moderate to high grazing by cattle (which prefer grasses over forbs) (Weiss 1999). Gr asslands  465

The plant community is another key focus of conservation and restoration efforts, with 40% of landowners employing management practices to decrease non-native plants and increase natives (Ferranto et al. 2011). At present, large-scale, complete eradication of non-natives is not feasible. Many restoration efforts have failed to achieve long-term self-sustaining native communities due to high rates of reinvasion of nonnatives (Malmstrom et al. 2009). However, some restoration sites have been successful, particularly with repeated, longterm management of non-natives through burning, mowing, and grazing (Bossard and Randall 2007). Because of this, current restoration goals focus on decreasing weeds, while maintaining or enhancing native grasses and forbs (Stromberg et al. 2007). These restoration efforts focus on three stages: reducing non-natives, restoring natives, and controlling reinvasion (Bartolome et al. 2007). Reducing non-natives is achieved as just described, with a combination of carefully timed grazing or burning, herbicide applications, thatch removal, and sometimes tillage (Bossard and Randall 2007, D’Antonio et al. 2007). Natives are then planted, either as drilled seeds or plugs (Stromberg et al. 2007). The use of local genotypes in restoration can be important, because they are best-suited for local environmental conditions (e.g., coastal versus inland, soil type), plant competitors, and management regimes (Knapp and Rice 1998, McKay et al. 2005, Bartolome et al. 2007). Once the natives are planted, aggressive management of weeds is typical for the first two to three years (Bartolome et al. 2007). However, a number of studies have shown that the natives can be competitive with the weeds (Seabloom et al. 2003), and that annual non-natives may dominate native restoration plots for the first few years, but then natives become dominant over the long-term, even without weed management (Corbin and D’Antonio 2004, Eviner et al. 2013). While native suppression of non-natives has been documented in a few cases, this level of successful restoration is still rare and is likely limited to key environmental conditions (e.g., moist coastal sites, valley bottoms with deep soils, and access to groundwater). Most successful restoration projects require long-term, aggressive management of annual non-natives (Bartolome et al. 2007, Malmstrom et al. 2009).

Forage Production The largest direct economic benefit of California’s grasslands comes from providing forage for livestock. Grasslands annually provide 75% of California’s livestock forage (Corbin et al. 2007a, CCCC 2009, Cheatum et al. 2011). Forage availability depends on aboveground plant production and the palatability of plant biomass. Palatability is strongly influenced by plant species composition, with species differing in tissue quality and in how long they remain green into the spring and summer (green forage is much more nutritious than senesced litter). High-quality forage species include legumes such as clovers and lupines, forbs such as filaree, and some grasses with longer green forage periods (e.g., Italian ryegrass) (George et al. 2001). Low-quality species include recent invaders such as yellow starthistle, medusa head, and barbed goatgrass, which have lower production and lower forage quality than the naturalized invaders and can decrease livestock productivity 50–​75% (Jacobsen 1929, Pitcairn et al. 1998, Gerlach and Rice 2003, Malmstrom et al. 2009). While environmental factors have the strongest impacts on both production and plant composition (see Figure 23.1), 466  Ecosystems

short-term improvements in forage production and composition can be achieved through planting legumes and fertilizing with nitrogen, and in some areas, fertilizing with phosphorus and sulfur (Heady et al. 1992). Forage quality and production can also be increased by controlling low-quality plants (particularly invasive noxious weeds) through the use of herbicides and carefully timed grazing or burning (Heady et al. 1992, Jackson and Bartolome 2007). Grazing management is one of the most effective and flexible tools for managing vegetation composition and production (Huntsinger et al. 2007). For example, grazing to maintain threshold levels of residual dry matter (RDM) can have a positive effect on production (Jackson and Bartolome 2002). Conversely, overuse of forage in one year can reduce production in the following year. To restore forage production in degraded grasslands, ranchers have moved away from continuous, season-long grazing, and are resting pastures and employing grazing rotations during key seasons, depending on the management goal (FRAP 2003). Because of these changes in grazing management, grassland conditions have been static or improving over the past few decades (FRAP 2003).

Pollination Many plant species in California grasslands are wind-pollinated or can self-polinate and thus do not require pollinators (Moldenke 1976). However, this is not the case for all species, and interactions between pollinators and many forb species can be critically important for gene flow (Chiariello 1989). These grasslands support high pollinator diversity and abundance (Wood et al. 2005, Colteaux et al. 2013) and are critical for providing pollen sources for both native bees and the honey bee during seasons when surrounding crops are not flowering (Moldenke 1976). Pollination of agricultural crops relies on the proximity of wildlands (Kremen et al. 2004), and since grasslands are often adjacent to agricultural crops, grasslands support a large portion of the pollinators for California’s agriculture (Chaplin-Kramer et al. 2011). The main threats to pollinators in California’s grasslands are habitat loss and invasion of non-native grasses, which decrease the abundance of forbs (Black et al. 2009). Pollinators require a diverse community of forbs, containing species that differ in phenology and morphological traits, so that collectively they bloom throughout the season and support diverse pollinator morphologies (Black et al. 2009). Diversity of forbs is particularly critical because many grassland forbs have short flowering times and vary in timing of flowering, depending on rainfall (Moldenke 1976). Restoration efforts to increase the prevalence of native forbs in California’s grasslands have successfully enhanced native pollinator populations and diversity (Black et al. 2009). Forb patches should be at least 0.2 hectares but are more effective when containing a core habitat of at least 0.8 hectares, surrounded by multiple smaller patches. These forb patches should be within 150 meters to 600 meters of nesting sites and crops that need to be pollinated, given the typical flight range of bees (Black et al. 2009). Grassland management practices that are typically used to enhance forbs (grazing, burning, and mowing) can have mixed effects on pollinators. While these management practices maintain forb cover and diversity, they can also disrupt pollinators by ruining nesting sites and interfering with immediate food supply (Black et al. 2009). To the extent possible, mowing and burning should

be timed to avoid flowering, and should avoid any summer blooms, when flowers are rare (and thus more crucial to pollinators) (Black et al. 2009). For both mowing and fire, these treatments should occur on no more than 33% of the habitat per year. This is particularly critical for fire, which can cause longer-term decreases in bee populations than mowing (Black et al. 2009). Livestock can destroy nests and trample bees and consume pollinator food (particularly livestock such as sheep, which prefer forbs) (Sugden 1985). Like mowing or fire, grazing management should be timed to minimize impacts on forbs during flowering times. When this is not possible, due to the need to control for noxious grasses, grazing should occur in small areas on any given year (Black et al. 2009).

Water Quality and Supply Almost all of California’s surface water passes through grasslands and oak woodlands (Tate et al. 1999). Thus grasslands can have strong impacts on water flow and quality. As discussed with water balance, grasslands have lower evapotranspiration than woody systems, so a higher proportion of rainfall flows into streams. In addition, because grassland soils have high infiltration, they attenuate any given storm event, leading to gradual release of the storm water to the streams (Lewis 1968, Dahlgren et al. 2001). This both reduces flood risk but also allows for continued streamflow into the dry season. Since these grasslands are naturally effective in water provision and flood control, management practices should focus on not compromising water infiltration and storage. For example, minimizing high densities of livestock during the wet season can prevent soil compaction, allowing for water infiltration. Water quality can be a key concern, since grasslands are susceptible to erosion due to typically thin soils and prevalence of steep topography (FRAP 2003). Grasslands can also be a source of nitrogen early in the growing season, when leaching rates are high (Jackson et al. 2007). However, grasslands can also serve as important filters of pathogens, nutrients, and sediments, and are effective buffer strips between agricultural and urban uplands and streams (Tate et al. 2006, Atwill et al. 2006). However, the ability of these grasslands to filter pollutants can be overwhelmed during large storms, so not surprisingly, nitrogen and sediment inputs into streams tend to be associated with high precipitation periods (Lewis et al. 2006). Grazing can be associated with impaired water quality, particularly on the North coast (FRAP 2003), but light grazing can also enhance water quality (Barry et al. 2006).

Carbon Sequestration California’s grasslands contribute significantly to regional carbon storage due to their large spatial extent, as well as high quantity of carbon storage per unit area (similar in quantity to temperate perennial grasslands, which are well known for their high carbon storage) (Silver et al. 2010). High root allocation contributes to soil organic matter storage, and rooting depth can impact the depth distribution of soil carbon. Deeper soil carbon tends to be more stable than surface carbon, as it is less likely to undergo disturbances such as gopher or earthworm activity, and decomposer activity is lower due to fewer resources (Silver et al. 2010). Across sites, soil carbon tends to increase with increasing soil clay content and is

highest in grasslands with intermediate aboveground net primary production (Silver et al. 2010). On average, California’s grasslands are carbon neutral, varying between being a weak source and a weak sink, depending on annual weather patterns (Xu and Baldocchi 2004, Ma et al. 2007, Kroeger et al. 2009). As with other ecosystem processes, carbon dynamics are more strongly affected by the seasonality of precipitation than the total annual precipitation (Ma et al. 2007, Chou et al. 2008). When late-phenology plants are present, longer growing seasons with wetter springs increase net primary production to a greater extent than decomposition (Berhe et al. 2012), resulting in net storage of soil carbon (Ma et al. 2007). However, when late-phenology plants are absent, late-season rains stimulate soil respiration but do not alter net primary production (Chou et al. 2008), leading to carbon loss. Higher rains during the winter can increase loss of soil organic matter, despite increases in net primary production, possibly due to decreased roles of iron and aluminum oxides in stabilizing soil carbon (Berhe et al. 2012). Despite the annual source-sink fluctuations, there is potential to increase carbon sequestration in some California grassland sites, although these protocols have not been approved for carbon credits (FRAP 2010), and important trade-offs may exist. Woody species increase soil carbon storage in California’s grasslands (Silver et al. 2010) but may also decrease water supply, as has happened in other semiarid regions (Mark and Dickinson 2008). It is assumed that native perennial grasses increase soil carbon storage, and observational studies have shown that soil carbon is higher under native perennials than under non-natives annuals (Koteen et al. 2011). However, in the Koteen study it is not clear whether natives preferentially establish on soils with higher soil carbon, or if they promote higher carbon in soils where they are present. Whether perennial grasses can enhance soil carbon can be more reliably determined through experimental plantings of native versus non-native plants on the same soil types, or by comparing restored versus adjacent unrestored areas that are on the same soil. Such studies have not detected a difference in total soil carbon between natives versus non-natives (Potthoff et al. 2005) but have found that the distribution of soil carbon changes. Soils associated with perennial grasses have deeper soil carbon than soils associated with annuals (Eviner et al. in prep.) and thus could lead to longer-term sequestration. Legumes can increase soil organic matter and microbial biomass carbon (Eviner et al. 2006, Potthoff et al. 2009) but may also enhance nitrous oxide emissions, a more potent greenhouse gas than carbon dioxide. Addition of inorganic nitrogen fertilizer has mixed impacts on soil carbon, sometimes increasing soil carbon storage through increased net primary production and litter quality, but other times decreasing it through decreasing root allocation and stimulating microbial breakdown of organic matter (Conant et al. 2001). Similar to legumes, fertilizer additions have the likely trade-off of increasing nitrous oxide production. In general, grazing has mixed effects on soil carbon storage (Conant et al. 2001, Derner and Schuman 2007), and broad comparisons of grazed versus ungrazed sites in California show no consistent effects of grazing on soil carbon (Silver et al. 2010). Light grazing does not tend to impact soil organic matter in California grasslands (Jackson et al. 2007), although overgrazing that results in high erosion has the potential to greatly decrease soil carbon. Carbon sequestration will be particularly vulnerable to wildfires and droughts, so is likely to decrease in response to climate change (FRAP 2010). Gr asslands  467

Fire Control While fires can be harmful to human infrastructure and air quality, they tend to be less of a threat in grasslands than in woodlands and shrublands (FRAP 2010). Fire control in grasslands is primarily managed through decreasing fuel load through grazing, prescribed fire, and/or mowing (FRAP 2010). The level to which thatch is removed has great impacts on fire severity. For example, a fuel load of 2,242 kilograms per hectare can lead to fires with 15-meter-long flames, while grazing to half that fuel load can limit flames to 1–​3 meters long. Grazing down to 560 kilograms per hectare leaves a fuel load that cannot support a continuous fire, so only isolated patches will burn (Barry et al. 2006).

Impacts of Humans on Grasslands As reviewed above, California’s grasslands greatly changed with European settlement, largely through the introductions of non-native plants (Bossard and Randall 2007) and domesticated livestock (Allen-Diaz et al. 2007) as well as conversion of grasslands to cropping systems. Several million hectares of California’s grasslands have been cultivated, with a peak of grassland conversion occurring in the late 1800s (Heady et al. 1992). More recent land use changes also strongly affect grassland structure and function. Extensive areas of grasslands were created from woody-dominated systems, particularly in the 1950s to 1960s, in an attempt to increase forage production (Standiford and Tinnin 1996). Currently in California, grasslands are the ecosystem most at risk from development (FRAP 2010). On average, over the past few decades, more than 190 square kilometers of grassland per year have been lost to vineyards, orchards, dispersed housing, and urban development, and this loss of grassland will continue in the future, particularly with losses to vineyards and urban areas (Jackson et al. 2007, FRAP 2010). Many large ranches are being subdivided, and these smaller parcels receive less management for species conservation and ecosystem services (Ferranto et al. 2011). In fact, many grassland areas are now experiencing undergrazing, where lack of fire or grazing leads to thatch buildup, domination by species such as ripgut brome, and declines in key services such as productivity, wildlife habitat, pollination, and plant diversity (Biswell 1956, Bartolome et al. 2007). Many shifts in disturbance regimes have occurred in California’s grasslands. Over the past few centuries, the hydrology of the Central and San Joaquin Valleys has been drastically altered by dams and levees, altering the types of grassland habitats supported, and preventing the flooding regimes that regularly maintained soil fertility (Corbin et al. 2007a). There have also been substantial changes in the fire regime. On the Central Coast, fires occurred every three to five years before 1880 and now occur every twenty to thirty years (Reiner 2007). In the Sierra foothills the fire return internal was twenty-five years before European settlement, then changed to seven years after settlement, and since the 1950s, fire suppression has led to rare fires (McClaran and Bartolome 1989). Particularly in areas that have reductions in both grazing and fire, these grasslands are susceptible to increased thatch buildup, higher fuel loads, and lower diversity (particularly of forbs and legumes). Nitrogen deposition is increasingly affecting California’s grasslands, but its effects are patchily distributed. Approximately 30% of California grasslands have at least 5 kg N/ha/ yr deposition, with levels up to 45 kg/ha/yr in southern Cali468  Ecosystems

fornia and 16 kg/ha/yr in northern California (Weiss 2006, Dukes and Shaw 2007). This nitrogen deposition can increase production (Dukes et al. 2005), decrease diversity (especially of forbs), and stimulate decomposition rates (Allison et al. 2013). Nitrogen additions tend to increase non-native grasses (Dukes and Shaw 2007), and nitrogen deposition rates are high enough to enhance non-native grasses on 44% of California’s grassland area (Fenn et al. 2010). Climate change is likely to have significant impacts on the structure and function of California’s grasslands. In this century, temperature rises are expected of 1.7oC to 3oC under low emissions, and 3.8oC to 5.8oC under high emission scenarios (Dukes and Shaw 2007, Cayan et al. 2008), with more warming inland than on the coast (Pierce et al. 2013). Summer temperatures will become markedly hotter. A modestly cool July in 2060 will be the same temperature as our hottest July temperatures to date. Mean temperatures in the winter will also increase, but the coolest days will be as cool or cooler than they are now (Pierce et al. 2013). Warming in the winter is expected to increase production and accelerate flowering and senescence of many species (Dukes and Shaw 2007), but cooler days may make plants more susceptible to frost kill. Annual changes in precipitation are likely to be modest, but there will be marked trends in seasonal patterns (Figure 23.11). For example, in northern California, winters will be 1–​10% wetter, but times of peak plant growth will be drier, with spring precipitation decreasing by 11–​18% and fall precipitation decreasing 3–​8% (Pierce et al. 2013). Southern California is also likely to have drier springs and falls, but unlike northern California, its winters will also be drier (1–​5%) and its summers will be wetter (46–​59%) due to monsoons (Pierce et al. 2013). While projections of precipitation changes are mixed (Dukes and Shaw 2007), all precipitation projections agree that there will be increased variability in precipitation across years, with increased frequency of El Niño events and a projected 1.5–​2.5-fold increase in drought frequency (Reever Morghan et al. 2007, Dukes and Shaw 2007). In addition, extreme rain events are likely to increase in frequency and magnitude, with a 10–​50% increase in large three-day rain events by 2060 (Pierce et al. 2013). The effects of these changes on precipitation will depend on when the precipitation falls. Increased precipitation during the rainy season will have little impact on overall production and species composition but can increase shoot production and decrease root production (Zavaleta et al. 2003, Dukes et al. 2005). Late-season rains have variable effects, depending on the study, but responses include increased perennials (Suttle et al. 2007), increased non-natives (Suttle and Thomsen 2007), increased abundance and diversity of forbs, and increased diversity of grasses (Zavaleta et al. 2003). Warmer and drier conditions are expected to increase shrubland area at the expense of grasslands, resulting in a 14–​58% decrease in forage production by the late 2000s (CCCC 2009). However, other climate scenarios predict an increase in the extent of grasslands at the expense of woody vegetation, as increased temperatures and increased frequency of droughts significantly enhance the frequency, intensity, and extent of fires, which woody species cannot tolerate (Dukes and Shaw 2007). Elevated carbon dioxide is another change that California’s grasslands are experiencing, which can lead to shifts in plant and microbial communities, independent of the changes in temperature that they can induce. The impacts of elevated CO2 will partially offset decreases in precipitation, since elevated CO2 increases water use efficiency of most plants, which

ΔP yearly (%) –0

–0 –1 –3 –4

7

–5

–5

–5

59

NorCal Northeast NorCal central Cal coast Sac/Cent Valley Sierra Central Nevada coast San Joaq valley Inland empire SoCal AnzaSoCal mtns Borrego coast

–5 –8

46

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ΔP Sep–Nov (%) –3

46

–1 –5

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–8 2

–29 –23

50

–10

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–1

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–30

10

2

–9

–20 –14 –13 7

9

5

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ΔP Mar–May (%)

ΔP Dec–Feb (%)

–4

2

0

10

20

30

% Change in precipitation FIGURE 23.11 Predicted percentage change in precipitation (ΔP) during the period 2060–​2 069, compared with the period 1985–​1994. Precipitation changes vary by region and by month. Top left: annual percentage change in precipitation. Top center: percentage change in precipitation over December, January, and February. Top right: percentage change in precipitation over March, April, and May. Bottom left: percentage change in precipitation over June, July, and August. Bottom center: percentage change in precipitation over September, October, and November. Bottom right: regional splits, as designated on all maps. Source: Modified from Pierce et al. 2013.

then increases spring soil moisture. This prolongs the growing season (Harpole et al. 2007) and accelerates nitrogen cycling (Dukes and Shaw 2007).

Management under Future Conditions Managing California’s grasslands under multiple environmental changes will be challenging, particularly when considering the need to balance management for multiple organisms and ecosystem services. Despite the complexity of the controls and responses of these grasslands, there are some clear challenges that lay ahead, and some relatively simple principles to consider for managing these challenges. While presented as discrete challenges, grassland managers will need integrated approaches to address all of these, simultaneously.

Management Challenge 1: The Interaction of Changing Precipitation Patterns and Non-Native versus Native Plants While predictions of future precipitation patterns are uncertain, all climate predictions emphasize that the annual varia-

tion of precipitation will be high. There will be an increased frequency in years with shorter and drier growing seasons as well as more frequent years with longer and wetter growing seasons (CCCC 2009). Vegetation composition will vary strongly along with precipitation. Late-season rainfall benefits the most recent grassland invaders, which are noxious weeds (e.g., goatgrass, medusa head, yellow starthistle). These are a management priority for both conservation and rangeland managers because these weeds decrease plant diversity, production, and forage quality (Pitcairn et al. 1998, Gerlach and Rice 2003, Malmstrom et al. 2009). These late-season noxious weeds decline during shorter, drier growing seasons, particularly when competing with other species that can use soil moisture early in the season (Eviner et al. in prep.). Fluctuating precipitation may allow for noxious weed control through restoration of native perennial grasses. Many native grasses overlap in phenology with the late-season noxious weeds, and once established, natives can suppress these weeds by up to 90% (Eviner et al. 2013). These natives are resilient to short-term droughts but also benefit from late-season rains (Reever Morghan et al. 2007), so are likely to establish and persist under these fluctuating conditions, providing control of the late-season noxious weeds during the years that receive late rainfall. Gr asslands  469

Management Challenge 2: Managing Fragmented Grasslands for Diversity Much of our current grassland area is under ranching, but many ranchers are uncertain if they, or future generations, will continue ranching, putting grasslands at risk for subdivision and development (Ferranto et al. 2011, Cheatum et al. 2011). As working ranches convert to dispersed housing with large properties, much less management for ecosystem services occurs, and grazing is often absent (Ferranto et al. 2011). Without grazing, thatch can build up to high levels, which can become a fire hazard, lower abundance and diversity of forbs and grasses, and decrease habitat for animals that are conservation targets, such as ground nesting birds (Barry et al. 2006). In addition, high thatch (5,000 kilograms per hectare) can increase the prevalence of noxious non-native weeds such as goatgrass and medusa head (Bartolome et al. 2007). The consequences of removal of grazers, without substituting controlled burns or mowing, has already been well acknowledged in many grassland reserves and parks, which are increasingly using short-term livestock rotations to remove fire fuel, manage non-native species, enhance plant diversity, and improve wildlife habitat (Weiss 1999, CCWD 2005, SRDC 2006). Conservation professionals will need to work with owners of small grassland parcels to implement some type of thatch removal, through controlled burns, grazing, or mowing.

Management Challenge 3: Managing for Grassland Resilience in the Face of Multiple Environmental Changes Since California grasslands are experiencing many types of environmental changes, it is critical to consider the simultaneous impacts of these multiple changes, which can interact in important and unexpected ways (Dukes and Shaw 2007). Predicting and managing the impacts of these multiple environmental changes is challenging, particularly considering the strong spatial and temporal variation in these environmental conditions (Bartolome et al. 2007, Hobbs et al. 2007). Long-term studies have demonstrated that California’s grasslands are resilient to fluctuating environmental conditions due to high plant diversity. Different plant species respond to unique suites of environmental conditions, so that rare species under some conditions become common in other conditions (Hobbs et al. 2007). Clearly, functional diversity of species is critical, but under changing conditions, there may be a loss of species with certain strategies. For example, nitrogen deposition tends to favor species with higher aboveground biomass allocation, which may lead to loss of species with higher root allocation that can withstand low soil moisture (Tilman and Downing 1994, Suding et al. 2005, Pan et al. 2011). Areas that lose these deep-rooted plants will lose a key strategy for drought resilience. To deal with these types of functional losses, managers should focus on maintaining biodiversity, while policy makers need to prioritize reversing certain environmental changes. For example, control over precipitation is much harder to achieve than reductions in nitrogen deposition, so while continued efforts should be made to mitigate climate change, we particularly need to push to decrease nitrogen emissions.

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Summary California’s grasslands are distributed across a wide precipitation gradient, ranging from 12 to 200 centimeters per year (Bartolome et al. 2007). The drier interior grasslands tend to be dominated by non-native annuals, while the wetter coastal grasslands tend to be dominated by a mix of native and nonnative perennials. Unique soil conditions (e.g., poor drainage, salinity, heavy metal toxicity) also define distinctive grassland types such as vernal pools, alkali sinks, and serpentine grasslands. Even within each of these grassland types, there is considerable variation in ecosystem structure and function, due to spatial and temporal variability in seasonal and annual weather patterns, topography, soil, disturbance regimes, and interactions among large herbivores, small mammals, insects, microbes, and plant communities. The predominance of annual species likely makes California’s grasslands particularly sensitive to intra-annual and interannual fluctuations in abiotic and biotic controllers. The high variability in multiple controlling factors leads to both challenges and opportunities in land management. Successful management and policy will have to shift away from a one-size-fits-all approach and embrace the reality that different techniques and guiding principles are needed from site to site, due to variations in soil, topography, and weather. In addition, at a given site, management recommendations may vary from year to year, due to high weather fluctuations (see Chapter 37, “Range Ecosystems”). Managers and scientists will need to collaborate on adaptive management approaches to understand how multiple environmental conditions interact to impact a given goal, while exploring the synergies and trade-offs associated with suites of species and ecosystem services needed from grasslands. The dominance of annuals over large areas of grasslands will require sustained management for many different goals but also provides a relative flexibility in “resetting” the system through adaptive management approaches. Grasslands are one of the most altered ecosystems in California (Corbin et al 2007a, Janzen et al. 2007), with non-native plant species comprising over 90% of plant cover in most areas (Bartolome et al. 2007). Despite this, California’s grasslands are a diversity hotspot, averaging greater than fifty plant species per 30 x 30 meter area (Heady et al. 1992) and providing habitat for nearly 90% of state-listed rare and endangered species (Skinner and Pavlik 1994), and seventy-five federally listed plants and animals (Jantz et al. 2007). They also provide 75% of the state’s livestock forage, the main direct economic benefit from these systems (Corbin et al. 2007a, CCCC 2009, Cheatum et al. 2011). These grasslands are critical in regulating water flow (e.g., flood prevention, maintaining streamflow into the dry seasons) (Lewis 1968, Dahlgren et al. 2001) and water quality (Tate et al. 2006, Atwill et al. 2006), and contribute significantly to regional soil carbon storage (Silver et al. 2010). Grasslands also support a large portion of the pollinators needed in California’s cropping systems (Chaplin-Kramer et al. 2011). Because 88% of California grasslands are privately owned (Jantz et al. 2007), conservation and restoration of these grasslands largely depends on private land owners and how they balance management for livestock production, biotic diversity, and ecosystem services (SRDC 2006, Barry et al. 2006, FRAP 2010, Ferranto et al. 2011). Currently, many ranchers actively manage to improve wildlife habitat, decrease noxious weeds, and enhance water quality (Ferranto et al. 2011).

However, as working ranches convert to dispersed housing with large properties, management for ecosystem services declines (Ferranto et al. 2011), and the lack of grazing can increase fires and lower diversity of forbs and grasses (Barry et al. 2006). Other threats to grasslands include conversion to agriculture (particularly vineyards and orchards) and urban areas, and high nitrogen deposition. Climate change is likely to increase the variability in precipitation, making it more challenging to reliably manage for suites of ecosystem services. High species diversity is critical for maintaining resilience of these grasslands to changes in the means and variability of biotic and abiotic controlling factors. Rare species under one set of conditions become the dominants under other conditions, so that the species that maintain ecosystem production vary greatly across time and space in this annual grassland (Hobbs et al. 2007).

Acknowledgments I am grateful to Jeff Corbin, Hall Cushman, Erika Zavaleta, and Hal Mooney for feedback on earlier versions, which improved this chapter. Thanks to Melissa Chapin for improving the figures and her attention to detail. This chapter is based on work supported by the University of California’s Division of Agriculture and Natural Resources competitive grants, the Western Sustainable Agriculture, Research and Education Program, and the U.S. Department of Agriculture’s Agriculture and Food Research Initiative's Managed Ecosystems and Weedy and Invasive Species programs.

Recommended Reading Bartolome, J. W., J. Barry, T. Griggs, and P. Hopkinson. 2007. Valley grassland. Pages 367–​393 in M. G. Barbour, T. Keeler-Wolf, and A. A. Schoenherr, editors. Terrestrial vegetation of California. University of California Press, Berkeley, California. Heady, H. F., J. W. Bartolome, M. D. Pitt, G. D. Savelle, and M. C. Stroud. 1992. California Prairie. Pages 313–​335 in R. T. Coupland, editor. Natural grasslands: Introduction and Western Hemisphere. Elsevier, Amsterdam, The Netherlands. Huenneke, L. F., and H. A. Mooney. 1989. Grassland structure and function: California annual grassland. Kluwer Academic Publishers, Dordrecht, The Netherlands. Stromberg, M. R., J. D. Corbin, and C. M. D’Antonio. 2007. California grasslands: Ecology and Management. University of California Press, Berkeley, California

Glossary Bulk density  The dry mass of soil divided by its volume, used as an indicator of soil compaction. Fecundity  Reproductive capacity or output. Forb  A flowering plant that is herbaceous but not a graminoid (grass, sedge, rush). When the term is used in contrast with grasses, this group often includes both legume and nonlegume plants. When the term is used in contrast with grasses and legumes, it is used to denote the herbaceous species that are neither graminoids nor legumes. Infiltration  The entry of water into the soil. Phenology  The timing of periodic events in the life cycles of organisms, often related to climate patterns (such as seasonality). In California’s grasslands the timing of plant

death in the dry spring is often categorized as early-, mid-, or late-season phenology. Photodegradation  The degradation of molecules by the absorption of light. Resilience  The ability of an ecosystem to recover from a disturbance. Senescence  Programmed breakdown and death of plant tissues. In annual species this allows plants to resorb nutrients from leaves that will die and allocate those nutrients to seeds. Soil orders  The broadest classification of soils, on a global level. Soil shear strength  The ability of soil to remain intact despite force applied against it.

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T WENT Y-FOUR

Chaparral V. THOM AS PAR KER , R . BR ANDON PR AT T, and JON E . KEELE Y

Introduction One of the most dynamic California ecosystems is chaparral. Dominated by evergreen, sclerophyllous shrubs and small trees, chaparral is the most extensive vegetation type in the state (Figure 24.1). The nearly impenetrable tangle of stiff branches of this unusual vegetation inhibits exploration, and as a consequence the public know little about its natural history and unique characteristics. This undervalued ecosystem is recognized instead by the threat of its extensive, high-intensity canopy-burning wildfires that characterize the dry summer and fall seasons of the state. Because urban areas frequently share borders or intermix with chaparral, societal interests often conflict with conservation of this ecosystem, and understanding its history and dynamics are key to appreciating its importance. Chaparral contains numerous plants and animals found in no other habitat, and many of them are rare and threatened. A large number of environmental and biotic influences drive this diversity, but three primary ones are the protracted summer rainless period of California’s Mediterranean-type climate, low-nutrient, and often shallow and rocky soils, and a fire regime that completely burns the vegetation one or more times a century. Within the widespread distribution of

chaparral, different combinations of these processes produce locally unique combinations of species, including more than one hundred evergreen shrub species across this range (Wells 1962, Keeley and Keeley 1988, Sawyer and Keeler-Wolf 1995). Many associations are named based on the dominant species, such as manzanita chaparral (Arctostaphylos sp.), chamise-redshank chaparral (Adenostoma sp.), or mixed chaparral (Sawyer and Keeler-Wolf 1995). Others are named based on the soils, such as serpentine and dune chaparral, or based on the climatic location, such as maritime chaparral (Griffin 1978) and montane chaparral (Hanes 1977) (Figure 24.2). In all these variations the overall character and dynamics of the vegetation are directly related to strong summer droughts, lownutrient soils, and wildfire. The extensive spatial distribution of chaparral means that stands reflect different climatic extremes and interact with different plant assemblages. At lower elevations, for example, the vegetation includes not only the dominant shrubs, but also a postfire flora that persists only a few years after fire. These annuals and short-lived perennials arise phoenixlike after wildfires, blanketing the landscape in colorful displays and subsequently remaining dormant as seeds in soil 479

until the next fire. In more northerly distributions in the coast ranges, chaparral intergrades with mixed evergreen and conifer forests, sometimes as a patchwork, often as successional vegetation. Cold temperatures become important at higher elevations and in cold-air drainage basins (Ewers et al. 2003). For example, chamise drops out at high elevations in interior Sierra Nevada drainages (Westman 1991) and in Arizona chaparral (Mooney and Miller 1985). At higher elevations, especially in the Sierra Nevada and Cascades, a variant referred to as montane chaparral dominates in patches in areas above the winter snowline. While sharing dominant genera and other dynamics with chaparral of lower elevation, this mountain chaparral contains unique species and lacks familiar components like the diverse, postfire herbaceous plant response. Besides the characteristic plants, chaparral contains numerous other organisms of which some are restricted or nearly so to chaparral. These include small rodents like woodrats (Neotoma sp.) and mice (e.g., Peromyscus californicus) as well as birds (e.g., wrentit and Bewick’s wren). Animals are involved in a variety of interactions, from dispersal of seed by scatterhoarding rodents and birds to herbivory and seed predation. Animal activity can be critical for the success of some plants (e.g., Parker 2010) or shift the plant composition of areas (Quinn 1994, Moreno and Oechel 1993, Ramirez et al. 2012). Less studied are other vertebrates and insects (Andres and Connor 2003, Miller 2005) and the soil biota, even though soil nutrients are often limiting to primary production and mycorrhizae influence vegetation dynamics (Horton et al. 1999, Dunne and Parker 1999, Treseder and Allen 2000, Egerton-Warburton et al. 2007). In this chapter we emphasize the principal structure and dynamics of this important ecosystem. The long, summer rainless period has strong impacts on all organisms and on the fire regime that characterizes chaparral. These features make significant impacts from climate change very likely. Attempts to suppress fire also affect chaparral dynamics. Because of its dominance at lower elevations, chaparral also frequently occurs at or near the boundaries of urban developments and metropolitan centers. Conflicts between the impacts of chaparral wildfire and human activities and structures have occurred throughout California’s history but have increased as development encroaches ever more into chaparral regions. Consequently, understanding of chaparral ecology is important not only because of its significance in understanding ecological evolution and ecological services provided by chaparral but also because of its direct impacts on human communities.

Geography of Chaparral Chaparral covers much of the Peninsular, Transverse, and Coastal Ranges and portions of the Sierra Nevada and Cascade Ranges. This represents over 9% of the wildland vegetation in California, with most of it below 2,000 meters in elevation (see recent reviews by Keeley 2000, Davis, Stoms, et al. 1998, Keeley and Davis 2007) (see Figure 24.1). Chaparral is

Photo on previous page: Santa Monica Mountains (Los Angeles County); chamise and redshank (Adenostoma fasciculatum and A. sparsifolium) with big-pod ceanothus in bloom (Ceanothus megacarpus). Photo: R. Brandon Pratt. 480  Ecosystems

particularly abundant in the mountains of southern California, which contain over a third of all chaparral found in that region (Davis, Stoms, et al. 1998), especially on the slopes of the Peninsular and Transverse Ranges (Cooper 1922, Epling and Lewis 1942). Chaparral extends south into northwest Baja California, with disjunct populations on mountainslopes as far south as 28°. Coastal mountain ranges are dominated by chaparral at most elevations in southern California but form mosaics with oak woodlands and conifer forest from central California northward. Throughout its range, chaparral is often replaced by localized patches of grassland. This is sometimes due to moisture and edaphic characteristics but more often due to disturbance (Wells 1962) (see Figure 24.2). North of the San Francisco Bay region, chaparral dominance shifts inland and progressively diminishes to widely scattered patches on interior slopes as far north as Washington State. The foothills of the Sierra Nevada and the southern Cascades similarly are covered by chaparral at elevations generally above 300 meters, giving way to forest with increasing precipitation at higher elevations (Keeley, Baer-Keeley et al. 2005). Within the upperelevation forest regions, chaparral appears on serpentine or other low-nutrient or shallow soils or after wildfires (Cooper 1922) (see Figure 24.2) and may persist as a consequence of self-reinforcing, high-intensity wildfires. Eastward, chaparral forms disjunct patches in mountainous areas in Arizona with particularly large chaparral landscapes above the Mogollan Rim in the middle of the state (Knipe et al. 1979). Some chaparral species form stands in areas of Arizona and Mexico with a summer rainfall regime (Vankat 1989; Bhaskar et al. 2007; Keeley, Fotheringham et al. 2012). Stands dominated by chamise (Adenostoma fasciculatum) are the most abundant type of chaparral in California (Hanes 1977). Chamise often forms nearly pure stands on hot and dry slopes that are generally equatorial-facing with shallow soils. In chamise stands, other common subdominant species are manzanitas (Arctostaphylos spp.) and ceanothus (or California lilac; Ceanothus spp.) along with various sage scrub species (e.g. Salvia spp. and Artemisia californica). Coastal and montane areas may also be dominated by nearly pure stands of ceanothus or manzanita species. On more mesic slopes and in deeper soils within a site, chaparral can contain a broader range of dominant evergreen species (Hanes 1971). Chaparral gives way to other plant associations based on water availability, temperature, soil, aspect, and elevation. At drier, low-elevation sites in nondesert areas, chaparral is replaced by sage scrub species (see Chapter 22, “Coastal Sage Scrub”). In valley bottoms with deeper soils, chaparral is replaced by oak savannas. In some cases, this transition can also be related to cold air drainage and freezing temperatures (Ewers et al. 2003, Pratt et al. 2005), animal activity (e.g., pocket gophers), or fire regime. Replacement by oak woodlands also occurs at higher elevations throughout the range of chaparral, which is generally attributed to greater rainfall. On more mesic sites, in the northern part of the state, chaparral gives way to an evergreen sclerophyllous woodland dominated by species such as tan-bark oak (Notholithocarpus densiflorus; formerly Lithocarpus d.), California bay laurel (Umbellularia californica), oaks, and madrone (Arbutus menziesii) (Cooper 1922). On desert-facing slopes, chaparral forms ecotones with Mojave desert scrub communities in the Transverse and Coast Ranges and with Sonoran desert shrub communities in the Peninsular Range and is replaced by these desert communities at more arid, lower elevations.

FIGURE 24.1 Distribution of chaparral vegetation in California. Data from Cal Fire, Fire Resource and Assessment Program (FRAP). Map: P. Welch, Center for Integrated Spatial Research (CISR).

A

B

C

D

FIGURE 24.2 Examples of variation in chaparral vegetation. Photos: V. Thomas Parker. A Chamise-dominated chaparral (Adenostoma fasciculatum) with big-berry manzanita (Arctostaphylos

glauca) as a co-dominant (Santa Lucia Mountains). B Maritime chaparral in the Santa Cruz Mountains dominated by Arctostaphylos crustacea,

A. andersonii, A. sensitiva, Ceanothus thyrsiflorus, and C. papillosus. C Chaparral occurs on unusual soils, in this case A. myrtifolia and A. viscida on Oxisols near Ione,

California. D Montane chaparral patches in the northern Sierra Nevada dominated by Quercus vaccinifolia,

Arctostaphylos patula, A. nevadensis, Chrysolepis sempervirens, Ceanothus cordulatus, C. integerrimus, Prunus emarginata, and Spiraea splendens alternating with Abies-dominated forest.

Origins Most dominant chaparral taxa date to the early Tertiary, with origins in the Eocene (Keeley, Bond et al. 2012). Assemblages with similarities to contemporary chaparral appear to have been present by the early Miocene, although the center of distribution was likely in the interior reaches of the southwestern corner of North America (Ackerly 2009; Keeley, Bond et al. 2012). By the mid-Miocene, chaparral dominants were evident in fossil floras from western Nevada and California, but the extent to which these communities resembled contemporary California chaparral is unknown. For example, contemporary chaparral in Mediterranean-climate California is markedly different from communities in summer-rain Arizona (Keeley, Fotheringham et al. 2012). Although these two different communities share many of the same shrub dominants, these climates have selected for very different postfire herbaceous assemblages. In the Mediterranean-type climate winter annuals dominate the community, whereas in the summer rain region the herbaceous community is dominated by perennials with a substantial contribution of C4 grasses. Thus, while shrub dominants appear to have very early Tertiary origins, the origin of contemporary California chaparral assemblages is likely tied to the origin of the Mediterranean-type climate. The timing of this event is a matter of some debate. Axelrod (1973) has long contended that the Mediterranean-­t ype climate was late in development, dating from the Pliocene or early Pleistocene. Others argue that this summer-drought climate originated much earlier and was in 482  Ecosystems

place in western Nevada and south-central California by midMiocene (Keeley, Bond et al. 2012; see Chapter 2, “Climate”). Evergreen sclerophyllous leaves occurred in lineages before the onset of the Mediterranean-type climate, suggesting that this feature of Mediterranean-type shrubs was not an adaptive response to that climate per se (Axelrod 1989, Verdú et al. 2003, Ackerly 2004a). Some have interpreted these sclerophyll taxa as relicts of Tertiary origins present today merely by chance avoidance of random extinctions rather than because they are adapted to contemporary environments (Herrera 1992, Valiente-Banuet et al. 2006). Others see a clear adaptive role for these taxa in the current landscape (Keeley, Pausas et al. 2012). Adaptive traits, however, are not necessarily adaptations, which are traits that have arisen via natural selection in response to a particular environmental factor such as the Mediterranean-type climate (Gould and Lewontin 1979). One view is that physiological and morphological traits in evergreen sclerophylls are adaptations to both water deficits and nutrient-poor soils, conditions present both before and after the widespread development of a Mediterranean-type climate (Keeley, Bond et al. 2012). In this model, the primary influence of the Mediterranean-type climate has not been through selection on these traits but rather has contributed to a massive expansion of suitable sites for these drought-adapted shrubs. Thus, in this view, many traits in contemporary chaparral shrubs are adaptations to contemporary conditions of periodic soil drought and nutrient-poor soils, but chaparral assemblages reflect sorting processes beginning with the origin of the Mediterranean-type climate (Ackerly 2004a).

Shrublands that resemble chaparral are also widely distributed in four other global regions: the Mediterranean basin, the Cape Region of South Africa, western and southern Australia, and central Chile—​a ll of which share a similar Mediterranean-type climate (however, see Cowling et al. 2005). The geographical and phylogenetic distances of species inhabiting these shrublands, coupled with their similar structure, led early biogeographers to postulate that the common Mediterranean-type climate of these regions spurred convergent evolution of the evergreen sclerophyllous shrub growth form (Schimper 1903). Mooney and Dunn (1970a) extended this conclusion by providing an ecophysiological model to explain the advantage of evergreen sclerophyllous leaves in Mediterranean-type environments. Following this, extensive comparisons between chaparral and the Chilean Mediterranean-type shrublands called matorral generally supported the convergence hypothesis (di Castri and Mooney 1973, Miller 1981). More recent fossil evidence showing that many of the ancestral evergreen sclerophyllous species predated the Mediterranean-type climate has been interpreted to mean that the convergence hypothesis is not supported by the widespread presence of evergreen sclerophyllous species in these regions (Axelrod 1989). Much of this debate is tied to an inordinate emphasis on similarities in the general climatic parameters of winter rain and summer drought evident across the five Mediterraneantype climate regions, rather than on differences among the regions. Evolutionary convergence is predicted when taxa evolve in similar “environments,” but each Mediterraneantype climate region exhibits subtle differences in rainfall patterns and not-so-subtle differences in soils and fire regimes that all would be expected to contribute to differences in plant traits and community assemblages (Keeley, Bond et al. 2012). Recent studies of water relations, particularly xylem structure and function, have found California chaparral shrubs to be convergent with South African fynbos shrubs when sites were matched (Jacobsen et al. 2009). However, shrubs from the two regions also differ in life history characteristics (Pratt et al. 2012).

Principal Organisms Found in Chaparral Plants While many plant species can be found in chaparral, the dominant shrubs that structure of the ecosystem represent a few genera common to most sites throughout the state. Climatic patterns and soil heterogeneity sort local shrub dominance, which in turn modifies ecosystem processes like hydrology and biogeochemistry as well as animal communities. Genera with multiple species, including rare species, include manzanita, ceanothus, shrub oak, and silk tassel bush (Garrya) (Table 24.1). Chamise occurs throughout the range of chaparral. California red-shank, a sister taxon (Adenostoma sparsifolium), sporadically joins chamise from Santa Barbara County south, mostly away from the coast. Other important genera that are monotypic or have only a few species include mountain mahogany (Cercocarpus), cherry (Prunus), coffee berry (Frangula), sugar or lemonade bush (Rhus), toyon (Heteromeles arbutifolia), chaparral pea (Pickeringia montana), and laurel sumac (Malosma laurina). Trees commonly associated with chaparral in some areas are usually species of pine (Pinus), cypress (Hesperocyparis),

oaks, and big-cone Douglas-fir (Pseudotsuga macrocarpa). The species of pines and cypress in chaparral are generally serotinous (e.g., knobcone pine, Sargent’s cypress) or semiserotinous, with slow-dispersing, heavy, thick cones (e.g., ghost, Coulter, or torrey pine). Stands of these trees are open with a dense chaparral understory or mosaic. In more mesic regions other tree species also found in adjacent mixed evergreen or coniferous forests can intergrade with chaparral. In southern California big-cone Douglas-fir is commonly found in chaparral from Santa Barbara County to San Diego County. Coast live oak (Quercus agrifolia) and sometimes other oaks can also invade chaparral; in some maritime associations coast live oak is a co-dominant of the chaparral. While chaparral contains a highly diverse component of annuals and herbaceous perennials, with few exceptions older stands of chaparral lack significant herbaceous cover (Hanes 1981). This is generally because most of these species have deep seed dormancy that requires stimulation from wildfire (Sweeney 1956). For other annuals and many herbaceous perennials lacking this dormancy, growth is limited by significant herbivory from chaparral animals or by limited resources (Mooney and Dunn 1970a, Christensen and Muller 1975, Swanck and Oechel 1991). Nevertheless, postfire stands of chaparral are often dominated by diverse herbaceous annuals and perennials for several years before shrubs regain dominance (Table 24.2). Chaparral also contains a number of suffrutescent shrubs whose presence or absence depends upon a number of factors. They are common in postfire stands because some or all of their seed banks are responsive to wildfire. However, these plants may persist long afterwards in more open conditions, such as in rockier habitats, along trails, in poor soil conditions such as in serpentine, or along the edges of drier systems like deserts or coastal scrub. While frequent or common in chaparral systems, these shrubs sometimes dominate adjacent plant communities.

Mycota and Other Microbiota Chaparral harbors considerable diversity of fungi and microbes, but little is known about individual species or their overall ecological impacts. These organisms play key roles in decomposition, nitrogen fixation, and mineral cycling, and some are pathogens of dominant plants. Experiments indicate that they are critical in many ecosystem processes and can be affected by wildfire (Horton et al. 1998, Stoll 1998, Baar et al. 1999, Peay et al. 2009) and increased atmospheric CO2 (Allen et al. 2005). Surveys of mycota and DNA fingerprinting indicate that chaparral stands can have a high diversity of mycota (Bradford 1998, Blair 1999). Over 134 taxa of fungi in 33 families and 17 orders were found beneath stands of manzanita (Blair 1999); 5 fungal families had the highest representation, including 2 familiar as boletes or amanitas. Almost all of these species are thought to form mycorrhizal relationships with plant hosts. In another study that sampled manzanita root tips, diversity of mycorrhizal fungi was extremely high (Bradford 1998). Within chaparral, species of pines, manzanitas, and shrub oaks all associate with ectomycorrhizae and as a consequence might facilitate successional dynamics from shrub- to forest-dominated sites (Amaranthus and Perry 1994, Horton et al. 1999, Dunne and Parker 1999, Bode 1999). Other plant genera associate with arbuscular mycorrhizae, usually with taxa in the Chaparr al  483

TA B L E 2 4 .1 Common plants in California chaparral

Species

Occurrence and distribution

Common name

Notes

Life history

Pinus

Pine

Especially P. attenuata, P. radiata, P. muricata, P. sabiniana

Serotinous or slow seed dispersal (semiserotinous)

Hesperocyparis (formerly Cupressus)

Cypress

Especially H. sargentii, H. macnabiana

Serotinous

Quercus agrifolia

Coast live oak

Most common oak found in chaparral

Pseudotsuga macrocarpa

Big-cone Douglas-fir

TREES

Coast ranges from San Obligate sprouter Francisco Bay region south Occasional in southern California mountains

Obligate sprouter

SHRUBS Adenostoma sp.

Chamise

Especially A. fasciculatum

Common and widespread

Facultative seeder

Arctostaphylos sp.

Manzanita

96 taxa in California; many rare

Common and widespread, especially in more mesic ranges

Facultative seeder or obligate seeder

Ceanothus sp.

California lilac; buckbrush; etc.

60 taxa in California; many rare

Common and widespread, mostly chaparral

Facultative seeder or obligate seeder

Quercus sp.

Scrub oak

Especially Q. berberidifolia, Q. durata, Q. vaccinifolia, Q. wislizenii var. frutescens

Common and widespread

Obligate sprouter

Heteromeles arbutifolia

Toyon

Common and widespread

Obligate sprouter

Pickeringia montana

Chaparral pea

Garrya sp.

Silk tassel

Especially G. elliptica, G. fremontii, G. veatchii

Frequent

Obligate sprouter

Prunus sp.

Holly-leaved cherry

Especially P. ilicifolia, P. emarginata, P. subcordata

Frequent, southern California to southern North Coast ranges, deciduous species at high elevation

Obligate sprouter

Frangula sp. (formerly Rhamnus)

Coffee berry

Especially F. californica, F. rubra

Widespread and frequent

Obligate sprouter

Cercocarpus sp.

Mountain mahogany

Especially C. betuloides

Widespread

Obligate sprouter

Rhus sp.

Sugar berry, Lemonade bush

Especially R. ovata, R. integrifolia

Mostly southern California Obligate sprouter or facultative seeder

Obligate sprouter

Malosma laurina

Southern California

Obligate sprouter

Source: Sweeney 1956, Keeley and Keeley 1988, Soule et al. 1988, Keeley, Fotherington, and Baer-Keeley 2005.

fungal genera Acaulospora, Glomus, Gigaspora, or Scutellospora (Allen et al. 1999). Chamise, in the rose family, appears to associate with mycorrhizae that are usually arbuscular; it has also been observed associating with ectomycorrhizae that include seven different mushrooms or cup fungi (Allen et al. 1999). Other members of the rose family appear to also produce both types of mycorrhizae (Smith and Read 1997). The roles of microbes in chaparral, though relatively poorly understood, undoubtedly play a critical role in a number of 484  Ecosystems

ecosystem functions. For example, mycorrhizae can produce extensive hyphal networks that can link multiple individual plants. While mycorrhizae generally are the principal pathway for mineral uptake, these networks also have been implicated in the survival of seedlings through summer drought (Horton et al. 1999, Dunne and Parker 1999, Egerton-Warburton et al. 2003, Egerton-Warburton et al. 2007, Plamboeck et al. 2007). Nitrogen cycling within vegetation is regulated largely by microbial communities, both bacterial and fungal

TA B L E 2 4 . 2 Representative plants found in postfire chaparral areas in California

Species

Family

Range

Chaenactis artemisiifolia

Asteraceae

Common, southern California

Emmenanthe penduliflora

Boraginaceae

Common

Phacelia brachyloba

Boraginaceae

Frequent

Phacelia parryi

Boraginaceae

Common, southern California

Phacelia grandiflora

Boraginaceae

Common, southern California

Phacelia cicutaria

Boraginaceae

Common, southern California

Phacelia minor

Boraginaceae

Common, southern California

Phacelia suaveolens

Boraginaceae

Common, northern California

Eucrypta chrysanthemifolia

Boraginaceae

Frequent

Cryptantha microstachys

Boraginaceae

Common

Silene coniflora

Caryophyllaceae

Frequent

Lupinus succulentus

Fabaceae

Occasional

Lupinus bicolor

Fabaceae

Common

Acmispon maritimus

Fabaceae

Occasional

Salvia apiana

Lamiaceae

Frequent, central, southern California

Salvia columbariae

Lamiaceae

Common

Calandrinia ciliata

Montiaceae

Common

Calyptridium monandrum

Montiaceae

Common, central, southern California

Eulobus californicus (Camissonia)

Onagraceae

Common, central, southern California

Ehrendorferia chrysantha (Dicentra)

Papaveraceae

Common, northern California

Ehrendorferia ochraleuca (Dicentra)

Papaveraceae

Common, southern California

Papaver californicum

Papaveraceae

Occasional

Romneya coutleri

Papaveraceae

Occasional, southern California

Antirrhinum coulterianum

Plantaginaceae

Common, southern California

Allophyllum glutinosum

Polemoniaceae

Frequent, central, southern California

Gilia capitata

Polemoniaceae

Common

Saltugilia australis (Gilia)

Polemoniaceae

Frequent, southern California

Chorizanthe fimbriata

Polygonaceae

Frequent, southern California

Artemisia californicum

Asteraceae

Common

Baccharis pilularis

Asteraceae

Common, esp. northern California

Ericameria arborescens

Asteraceae

Common, northern California

Eriodictyon californicum

Boraginaceae

Common, central, northern California

Eriodictyon crassifolium

Boraginaceae

Common, southern California

Helianthemum scoparium

Cistaceae

Common

Acmispon glaber (Lotus scoparius)

Fabaceae

Common

Lepichinia calycina

Lamiaceae

Common, northern California

COMMON POSTFIRE ANNUALS

COMMON SUFFRUTESCENTS

(continued)

TA B L E 2 4 . 2 (continued)

Species

Family

Range

Salvia mellifera

Lamiaceae

Common, southern California

Mimulus aurantiacus

Phyrmaceae

Common

Eriogonum fasciculatum

Polygonaceae

Common, southern California

COMMON SUFFRUTESCENTS

Sources: Sweeney 1956, Keeley and Keeley 1988, Soule et al. 1988, Keeley, Fotherington, and Baer-Keeley 2005.

(Grogan et al. 2000). Increases in atmospheric CO2 concentrations appear to increase the importance of microbial regulation of nitrogen (Allen et al. 2005), as elevated CO2 increases nitrogen deficiency in chaparral soils.

Chaparral Animals Invertebrate diversity in chaparral is thought to be considerable, but relatively few studies have been conducted at the community level. Invertebrates are a key component of ecosystem processes of mineral and energy flow as detritivores and folivores. Also, most chaparral plants are pollinated by a diversity of insects (Mosquin 1971, Fulton and Carpenter 1979). Other invertebrates are key members of food webs as parasitoids and predators and in other trophic roles. One study of a single montane chaparral species, green-leaf manzanita, found over 500 arthropod taxa from 169 different families in 19 orders (Valenti et al. 1997). About 80% of these species were herbivores, predators, or parasitoids. Another study investigated insects associated with leaves and branches of 26 coastal manzanita species and found over 209 insect taxa, with over 85% of them folivores (Andres and Connor 2003). They found a density of approximately 350 individuals m-2 for just leaf miners, leaf gallers, sap-suckers, and chewing insects. These studies indicate the importance of small insects and other arthropods in energy and mineral cycling of chaparral food webs. Chaparral also is a habitat in which many insect lineages have evolved (Miller and Crespi 2003). For example, one aphid genus (Tamallia) has radiated on species of manzanita (Miller 1998a, 1998b). Chaparral also harbors a large number of vertebrates, from mammals like woodrats, chipmunks, and harvest mice (Table 24.3) and birds (Table 24.4) like Bewick’s wren and the wrentit to a variety of common reptiles like the western fence lizard (Scleroporus occidentalis) and the Pacific rattlesnake (Crotalus viridis). These animals are generally granivores (seed eaters), herbivores, insectivores, or other types of predators that modify and extend the food web found within chaparral. Because of their size and density, such animals can significantly influence the dominance and frequency of certain plants (Quinn 1994, Mills 1983, Frazer and Davis 1988). For example, grazing by deer or rodents in the first several years after fire tends to differentially impact certain plant species, shifting their dominance (Mills 1983, Quinn 1994, Frazer and Davis 1988, Ramirez et al. 2012). Seed predators can limit seed input to seed banks (Keeley and Hays 1976, Kelly and Parker 1990, Quinn 1994, O’Neil and Parker 2005, Warzecha and Parker 2014). On the other hand, scatter-hoarding rodents may be critical in the burial 486  Ecosystems

of seed for many persistent seed bank species, burying them deep enough to permit survival of high temperature wildfires (Parker 2010).

Life Histories and Wildfire Fires are a natural and critical ecosystem process in chaparral. Chaparral dynamics correspond to cycles of wildfire, postfire recovery, and stand maturation. Chaparral is resilient to fires at 30 to 150+ year intervals, and within this range communities quickly return to prefire conditions. This occurs because all components of the prefire state are present after fire and because colonization plays a limited role in reestablishing vegetation (Hanes 1971). Of critical importance is that chaparral is adapted to a particular fire regime of a range of frequency, intensity, and timing (see Chapter 3, “Fire as an Ecosystem Process”). A departure from the natural fire regime, either by excluding fire or adding too much fire, reduces the sustainability of this ecosystem (Zedler et al. 1983, Parker 1990, Zedler 1995, Parker and Pickett 1998, Jacobsen et al. 2004, Keeley et al. 2005a, Keeley et al. 2005b). Postfire plant regeneration involves dormant seed banks that germinate after fire and resprouting from persistent stem bases and lignotubers (Parker and Kelly 1989). Consequently, fire size often does not affect community recovery. Rather, an appropriate fire regime that allows stands to recover and begin reproduction is critical. Because canopy fires typify chaparral, intensity and heat penetration into the soil vary with temperature, wind, soil, and fuel moisture prior to the fire, as well as aspect, exposure, and other characteristics. Heat penetration into the soil critically influences outcomes because many plant species recover from seed in the soil that must survive heat pulses (e.g., Odion and Davis 2000, Odion and Tyler 2002). Different types of seed banks can be found among chaparral plants. They illustrate a range of adaptations based on the type and degree of seed dormancy. Species with seeds that lack any type of long-term dormancy and germinate within a year are described as having transient seed banks. For these species a period of time exists when no seed reserve is found in the soil. Should a wildfire occur during that time period, the population will be eliminated unless some other life history stage can survive the fire. Species with more extensive seed dormancy, barring predation, always have seeds present in the soil. These are referred to as having persistent seed banks, whose crucial characteristic is that a reserve of seed is typically available. Trees found in chaparral often have serotinous cones, considered a persistent canopy or aerial seed bank.

TA B L E 2 4 . 3 Common mammals found in chaparral areas in California

Species

Common name

Taxonomic subgroups

Occurrence in chaparral

LARGE MAMMALS Puma concolor

Mountain lion

Occasional

Lynx rufus

Bobcat

Common

Odocoileus hemionus

Mule deer

Common, esp. postfire

Canis latrans

Coyote

Common

Taxidea taxus

Badger

Occasional postfire

Procyron lotor

Raccoon

Occasional postfire

Urocyron cineroargenteus

Gray fox

Occasional

SMALL MAMMALS Tamias sp.

Chipmunks

Especially T. merriami, T. sonomae T. quadrimaculatus,

Frequent, common at higher elevations

Spermophilus sp.

Ground squirrel

Especially S. lateralis, S. beecheyi

Sylvilagus sp.

Brush rabbit

Especially S. bachmanii and S. audobonii

Common

Lepus sp.

Jackrabbit

L. californicus

Common

Chaetodipus sp.

Pocket mouse

Especially C. californicus and C. fallax

Frequent, especially in southern California

Dipodomys sp.

Kangaroo rat

D. venustus, D. agilis, D. heermanni

Coastal Ranges

Neotoma sp.

Woodrats

Especially N. fuscipes, N. lepida

Common

Peromycus sp.

Deer mouse

Especially. P. maniculatus, P. boylii, P. californicus

Common

Reithrodontomys sp.

Harvest mouse

Especially R. megalotis

Frequent

Vole

Especially M. californicus

Frequent postfire

Perognathus californicus Microtus sp.

Sources: Based on Lawrence 1966, Fellers 1994, Price et al. 1995, Schwilk and Keeley 1998, Laakkonen 2003.

Characteristic postfire annuals (or pyro-endemics) and suffrutescents usually produce a wholly or partially dormant, persistent soil seed bank that responds to wildfire by losing dormancy and germinating in the next growing season (Sweeney 1956, Hanes 1977, Keeley 1991). Dormant seed banks of these species are triggered by either intense heat shock or combustion products from smoke or charred wood (e.g., Keeley 1991, Keeley and Fotheringham 2000). Their relative dominance depends on site history and the rainfall and temperature pattern of the initial postfire year, but generally their cover is significant and they continue to expand their populations into the second year. Consequently, chaparral typically has the highest plant diversity in the first and second years after fire (Sweeney 1956, Keeley et al. 2005a). Plant diversity declines in later years, although this trend may be reversed by very high rainfall in early seral stages (Keeley et al. 2005b). While annuals decline, suffrutescents remain until overtopped by reestablishing sclerophyllous shrubs or trees. Within chaparral, some annuals and suffrutescents typically establish in gaps and tolerate drought and other envi-

ronmental conditions found in postfire habitats. The substantial restriction of these widespread native annuals and suffrutescents to postfire stands evinces the long history of wildfire shaping plant community dynamics in the Mediterranean-type climate. Woody plants can be grouped into three general, postfire life history categories based on combinations of seed dormancy and postfire resprouting (Keeley 1987; Keeley, Bond et al. 2012; Parker and Kelly 1989). One cluster of species survives fire as adults; their aboveground stems are killed, but they resprout from stem or root crowns afterwards. Because of their transient seed banks, these species have no postfire seedling recruitment; consequently these types of plants are considered obligate resprouters. The seeds and seedling establishment patterns of obligate resprouters reflect no specific response to fire and are similar in reproductive characteristics and patterns to close relatives in other vegetation types. Two groups of plants, however, produce seeds that are dormant at maturity and create persistent soil or aerial canopy seed banks. Their seeds are wholly or principally stimuChaparr al  487

TA B L E 2 4 .4 Common birds found in chaparral areas in California

Species

Common name

Occurrence in chaparral

COMMON CHAPARRAL BIRDS Callipepla californica

California quail

Common

Thryomanes bewickii

Bewick’s wren

Common

Chamaea fasciata

Wrentit

Common

Toxostoma redivivum

California thrasher

Common

Psaltriparus minimus

Bushtit

Frequent

Aphelocoma californica

Western scrub jay

Common

Piplio maculatus

Spotted towhee

Common

Passerina amoena

Lazuli bunting

Occasional

Melozone crissalis

California towhee

Common

Melozone fuscus

Canyon towhee

Frequent

Polioptila californica

California gnatcatcher

Occasional in southern California

Geococcyx californianus

Road runner

Occasional

Calypte costae

Costa’s hummingbird

Occasional

Calypte anna

Anna’s hummingbird

Occasional

Artemisiospiza belli

Sage sparrow

Occasional

Aimophila ruficeps

Rufous-crowned sparrow

Occasional

Spizella atrogularis

Black-chinned sparrow

Occasional

Zenada macroura

Mourning-dove

Occasional, especially postfire

Columba livia

Rock dove

Occasional, especially postfire

COMMON POSTFIRE PREDATORS Buteo jamaicensis

Red-tailed hawk

Occasional postfire

Accipiter cooperii

Cooper’s hawk

Occasional postfire

Accipiter striatus

Sharp-shinned hawk

Occasional postfire

Falco sparverius

American kestrel

Occasional postfire

Bubo virginianus

Great horned owl

Occasional postfire

Corvus corax

Raven

Occasional postfire

Sources: Based on Lawrence 1966, Soule et al. 1988.

lated by wildfire and they germinate and establish in postfire stands. Of these plants, many can also resprout after fire and are considered facultative seeders, reflecting the survival of the adults and the postfire potential for a flush of new recruits. Finally, a third group of species does not resprout after fire, and their adults are killed by fire. Their populations persist exclusively through seed banks and seedling recruitment to reestablish their populations. These types of plants are called obligate seeders and in California are made up primarily by manzanitas and ceanothus among the shrubs and by pines and cypresses among the trees. Common obligate resprouters are toyon and shrub oak

488  Ecosystems

species (see Table 24.1). Toyon produces fleshy, bright-red, bird-dispersed fruit that mature in early winter. Once deposited in the droppings of birds, toyon seeds generally germinate quickly. The seedlings tolerate the shaded conditions in the understory of a mature chaparral canopy. Similarly, oak acorns mature in the fall and are dispersed by scatter-hoarding birds or rodents; acorns not buried shortly after falling from shrubs dry out and lose viability over a few weeks. The seeds lack complex dormancy mechanisms and germinate following initial fall rains or after stratification, resulting in yearly germination under chaparral canopies. Obligate resprouters thus build up small to extensive seedling banks in

Optimal growth period

Suboptimal growth period

Moisture limitation

100 Percentage of maximum

the understory of older-growth chaparral (e.g., Keeley 1992) and to some extent depend on longer fire-free intervals in chaparral to provide the occasional and necessary canopy gap for seedling emergence. Facultative and obligate seeders both produce persistent soil seed banks. Such seed banks derive from physiological dormancy mechanisms (manzanitas), have thick seed coats that prevent the entry of water until modified by heat pulses (physical dormancy; ceanothus species), have combinations of physical and physiological dormancy mechanisms (chamise), or retain seeds in thick woody cones that remained closed until opened by age or wildfire (pines and cypresses). In each of these ways, recruitment is restricted to the first year following wildfire. Fire opens canopies, increasing available light energy and soil surface temperature; ashes organic matter, making minerals available; and limits leaf area, reducing overall water loss from the soil by transpiration. These conditions are relatively ideal for seedling establishment in these genera, although the overall success of establishment then depends on survival of summer drought and herbivory. Woody plants that fall into these seed bank categories predictably reflect other sets of adaptive characteristics. Surveys of chaparral over large regions indicate that chaparral is dominated by species with persistent soil seed banks (facultative and obligate seeders) (Parker and Kelly 1989). Shrubs with persistent seed banks average over 80% cover across both coastal and interior conditions (Vasey et al. 2014). Patterns are occur within sites. Obligate resprouters, for example, are often more mesophytic (moisture loving) and tend to dominate more mesic sites like north-facing slopes or ravines. Obligate seeders and some facultative seeders are more xerophytic (adapted to dry conditions) in structure and dominate drier sites such as ridgelines and south-facing aspects. These persistent seed bank species might also be seen as occupying a gradient of population dynamic responses to fire, from species with adults surviving most wildfires, to species losing a considerable number of adults in fires, to obligate seeders that lose all adults that burn. The trade-offs in retaining some adults vegetatively versus sexually reproducing a cohort of seedlings after a fire lead reflect complex combinations of allocational trade-offs, history, site productivity, spatial and temporal environmental variability at a site, and the frequency with which fire returns. Organisms besides plants also have to survive wildfire. Wildfires undoubtedly have substantial impacts on microbial communities (Kaminsky 1981), but soils tend to buffer them except in the uppermost levels (Taylor and Bruns 1999). Like many chaparral plants, microbes have resistant stages. For example, fungi build up spore banks containing dormant spores and sclerotia in the soil (Taylor and Bruns 1999, Baar et al. 1999). The heat and ash from wildfires can influence the dominance of particular microbial species in postfire stands, sometimes favoring particular groups of fungi (e.g., Ascomycetes, Stoll 1998) but differentially sorting microbial species based on their ability to tolerate heat and ash (Baar et al. 1999, Izzo et al. 2006, Peay et al. 2009, 2010). Reviews of the roles of soil microbes in ecosystem processes and the influence of wildfire on those soil communities can be found in Neary et al. (1999) and Cairney and Bastias (2007). Animal recovery following fire differs profoundly from plant responses. Most plant taxa regenerate endogenously from dormant seed banks and/or resprouting and are relatively insensitive to fire inten-

80 60 40 20 0

Daylight hours Temperature Rain N

D

J

F

M

A

M

J

J

A

S

O

Month FIGURE 24.3 Seasonality of the Mediterranean-type climate. Seasonal maxima of three key factors are illustrated: daylight hours, temperature, and rain. At top, suboptimal and optimal growth periods for plant growth are represented, followed by the season in which moisture limitation restricts growth potential. Source: Data calculated from long-term average monthly means at the Los Angeles Civic Center (NOAA).

sity and fire size (Keeley et al. 2005a). In contrast, many chaparral fauna are far more sensitive to fire behavior characteristics, including fire severity and fire size, as well as the extent to which land management practices have fragmented metapopulations and altered corridors (see Chapter 3, “Fire as an Ecosystem Process”).

Physiology of Chaparral Shrubs In the past fifty years, the physiology of evergreen chaparral shrubs has been more intensively studied than that of perhaps any other plant community. Many excellent reviews have been written of chaparral shrub physiology (Mooney and Parsons 1973, Mooney et al. 1977, Miller and Hajek 1981, Mooney and Miller 1985, Carlquist 1989, Field and Davis 1989, Davis, Kolb, and Barton 1988, Keeley 2000). Chaparral shrub physiology is strongly shaped by the seasonality of the Mediterranean-type climate. Winter and spring rains recharge soil moisture, with most of the rain falling when temperatures are cool and days are short. This is the time when growth and photosynthesis of shrubs and ecosystem processes are most active (Figure 24.3). The evolutionary implications of this have not been well explored. Much of the interest in evolution of chaparral taxa has focused on the summer dry season, but many traits (particularly reproductive ones) are tied to winter rains, which likely have been present far longer than dry summers (Keeley, Bond et al. 2012). A second factor affecting chaparral physiology is the recurrent crown fires that occur during the summer or fall. These fires open space for seedlings of obligate seeders, gap specialists with a suite of life history and physiological characters linked to the postfire regeneration niche, to recruit (Figure 24.4). Obligate resprouters, in contrast, recruit seedlings during fire-free intervals and resprout after fire, and their physiology and life history traits diverge from the obligate seeders.

Chaparr al  489

FIGURE 24.4 Abundant seedlings of (left) Ceanothus megacarpus, a postfire obligate seeder, and (right) resprouts of chamise (Adenostoma fasciculatum). Photos: R. Brandon Pratt.

490  Ecosystems

8

24

Carbon assimilation Soil moisture

22 20

6

18 16

4

14

2 0 May

Soil moisture (%)

Vegetative growth of chaparral shrubs is affected by temperature, by photoperiod, and most strongly by available soil moisture (Miller and Hajek 1981, Miller 1983). During the wettest winter months, temperatures and photoperiod are at their lowest, producing suboptimal growth conditions (see Figure 24.3). The peak growing season for most shrubs is in spring, when temperatures and photoperiod rise and soils are still moist (Mooney et al. 1975, Mooney et al. 1977). During the dry summer and autumn months, growth is limited by low soil moisture availability (Davis and Mooney 1986). Timing of flowering diverges among shrubs, with some species flowering in the winter (e.g., big-berry manzanita and bigpod ceanothus), some in the spring (e.g., sugar bush and scrub oak), and still others during the summer and autumn (e.g., chamise and toyon) (Bauer 1936). Winter- and spring-flowering species tend to produce flower buds in the prior year and thus flower on old growth prior to current-year leaf production, whereas summer-flowering species flower on new growth following leaf production. The evergreen leaves of chaparral shrubs can photosynthesize year-round, but rates are moderate in the winter and early spring and increase in early summer (Figure 24.5). The chief factors affecting photosynthesis, ordered from most to least important, are available soil moisture, photoperiod, and temperature (Mooney et al. 1975). During the winter-growing season, chaparral species are able to photosynthesize at nearmaximum rates over a broad range of temperatures (85% of maximum rates can be achieved between approximately 10°C and 30°C) (Mooney et al. 1975, Oechel et al. 1981). This broad temperature response is adaptive because daily and seasonally temperatures can fluctuate broadly. During the summer and autumn months, shrubs restrict stomatal apertures to conserve water and avoid desiccation, which limits diffusion of CO2 to chloroplasts. The cool and short days during growing season favor C3 photosynthesis and C4 species virtually absent from chaparral communities (Sage et al. 1999). Evergreen and deciduous leaf habits represent two different strategies for photosynthetic carbon assimilation. Longerlived evergreen leaves have lower maximum photosynthetic rates than shorter-lived deciduous ones (Table 24.5). For evergreen chaparral leaves, net carbon gain accrues more slowly

10

Net carbon assimilation rate (µmol m-2 s-1)

Growth and Photosynthesis

12 10 Jul

Sep

Nov Jan

Mar

May

Jul

Sep

Month FIGURE 24.5 Seasonal patterns of soil moisture and maximum photosynthesis (carbon assimilation) for a typical chaparral shrub. Unpublished data from a southern California site are for Rhus ovata and soil moisture content at 2 meter depth. Source: Modified from Mooney and Dunn 1970b.

over a longer period, whereas deciduous leaves have a faster return rate over a shorter period (Mooney and Dunn 1970b, Harrison et al. 1971, Orians and Solbrig 1977, Field et al. 1983, Field and Mooney 1986, Ackerly 2004b, Wright et al. 2004). Evergreen leaves have lower specific leaf area (SLA) and are more sclerophyllous and mechanically stronger and stiffer, providing a more protected and durable leaf with greater longevity than deciduous leaves (Mooney and Dunn 1970b, Balsamo et al. 2003, Wright et al. 2004). In addition, deciduous leaves tend to have higher nitrogen levels and greater SLA (because they are less sclerophyllous) than evergreen leaves. In more arid sites evergreen chaparral gives way to sage scrub and desert scrub communities dominated by deciduous species. This pattern has been explained in the context of leaf economics (Mooney and Dunn 1970b, Mooney 1989). Because stomatal closure limits evergreen photosynthesis during the dry season, the costs of long-lived evergreen leaves exceed the return they can achieve at sites with more protracted dry seasons such as desert ecotones and coastal areas (Poole and Miller 1981, Mooney 1989). In these more arid sites the net carbon gain of deciduous leaves exceeds evergreens, giving the former a competitive advantage.

TA B L E 2 4 . 5 Leaf function of deciduous (n=6) and evergreen (n=6) chaparral shrubs grown in a common garden

Leaf habit

Anet area (µmol m-2 s-1)

Deciduous

Evergreen

Anet mass (µmol kg-1 s-1)

Nitrogen (%)

Specific leaf area (m2/kg)

Tensile strength A (N/mm2)

31.5 (3.6)

414.1 (69.1)

4.16 (0.40)

12.71 (0.80)

0.72 (0.15)

23.6 (4.1)

191.5 (3.5)

2.61 (0.16)

8.88 (1.30)

1.45 (0.24)

Source: Pratt, unpublished data. Note: Data are means with 1SE in parentheses. A. This is the modulus of rupture, or maximum force before breaking per unit leaf cross-sectional area of leaf.

TA B L E 2 4 . 6 Leaf stress traits: Adaptive leaf traits for coping with the stressful summer and autumn dry season typical of a Mediterranean-type climate

Leaf traits

Function

References

Sclerophyllous leaves

Structural support when turgor is lost and prevention of cell implosion; durability

Oertli et al. 1990, Brodribb and Holbrook 2005, Balsamo et al. 2003

Xeromorphic leaves

Aid in water retention

Cooper 1992

Stomatal response to drying soils

Reduce transpiration

Poole and Miller 1975, Jacobsen et al. 2008

Leaf angling during dry season

Reduce interception of solar radiation; thermal balance; reduced transpiration and photoinhibition

Comstock and Mahall 1985, Ehleringer and Comstock 1989, Valladares and Pearcy 1997, Valiente-Banuet et al. 2010

Heat shock proteins

Aid in stress tolerance

Knight 2010

Xanthophyll cycling and carotenoid accumulation

Protect photosynthetic pigments

Stylinski et al. 2002

Water Stress The protracted summer/fall rainless season creates hot, dry, and stressful conditions. South-facing slopes are often more arid and dominated by drought-tolerant species such as ceanothus, manzanitas, and chamise. North-facing slopes experience lower evaporation; however, north-facing slopes may dry out more rapidly and to a greater degree than south-facing slopes in mature stands because of a higher leaf area index and stand transpiration (Ng and Miller 1980). The effect of aspect may be especially important after fire when sensitive seedlings and resprouts are in the early establishment stage. Fog along the coast can help to mitigate the summer dry season (Vasey et al. 2012) because fog reduces evapotranspiration, and in some coastal areas because fog drip can lead to significant soil inputs of precipitation (Corbin et al. 2005). In southern California’s coastal areas fog inputs to the soil apparently do not provide a significant water source for shrubs (Evola and Sandquist 2010). Chaparral shrubs must cope during the protracted summer and fall rainless season with water stress, high temperature stress, and solar radiation that exceeds photosynthetic needs. The leaves of chaparral shrubs have a host of traits that equip them to manage these stressful conditions (Table 24.6).

Shrubs can be categorized along a continuum by the degree of water stress they experience during an average dry season, which is measured as the minimum seasonal water potential (Bhaskar and Ackerly 2006). At one end of the continuum are water stress tolerators that experience low water potentials during the dry season because they are relatively shallowly rooted (e.g., some ceanothus and manzanita species) (Hellmers, Horton et al. 1955; Poole and Miller 1975; Miller and Poole 1979; Poole and Miller 1981; Thomas and Davis 1989; Ackerly 2004b; Jacobsen et al. 2007a). At the other end of the continuum are deeply rooted water stress avoiders that experience a narrower range of water potential declines seasonally. This group is exemplified by Anacardiaceae including laurel sumac, sugar bush, lemonade bush (Miller and Poole 1979, Poole and Miller 1981, Thomas and Davis 1989, Jacobsen et al. 2007a). Most species experience water potentials in between the tolerators and avoiders, indicating that they have intermediate rooting depths or those varying from intermediate to deep depending on local edaphic conditions (e.g., chamise, ceanothus, toyon, scrub oak). Even water stress avoiders with deep roots experience a drop in water potential during the dry season, because all shrubs have some roots in shallow soil layers to acquire nutrients (Kummerow, Krause et al. 1978; Marion and Black 1988). Chaparr al  491

TA B L E 2 4 .7 Tolerator/avoider strategies for coping with water stress

Stress tolerators

Intermediate

Stress avoiders

Ceanothus spp. subgenus CerastesA

Heteromeles arbutifolia

Malosma laurina

2.4

n/a

>13.2

1

Minimum seasonal water potential (MPa)

-6.9

-4.3

-1.8

2, 3

Stem cavitation resistance (MPa)

-9.1

-6.2

-1.6

2, 3

Water potential at stomatal closure (MPa)

-5.5

-3.5

-2.2

4

Turgor loss point (MPa)

-5.6

-4.0

-2.3

5, 6, 7

Traits Maximum rooting depth (m)

B

Density of stem xylem (kg/m3) Stem mechanical Strength (N mm ) -2

Stem xylem starch storage (%) Vessel implosion resistance (t/b)h Vessel density (#/mm ) 2

Vessel diameter (µm) Stem hydraulic efficiency

C

2

References

674

611

462

2, 8

251

238

168

2, 3

2.9

3.0

4.8

0.041

0.035

0.019

9 2, 8

244

182

129

10

25

26

53

2, 8

1.5

2.3

5.7

2

Sources: 1. Thomas and Davis 1989; 2. Jacobsen, Pratt, Ewers et al. 2007; 3. Pratt, Jacobsen, Ewers et al. 2007; 4. Poole and Miller 1975; 5. Pratt et al. 2005; 6. Roberts 1982; 7. Calkin and Pearcy 1984; 8. Pratt et al. 2008; 9. Pratt, unpublished data; 10. Anna Jacobsen, unpublished data. A. Data are from Ceanothus crassifolius, C. cuneatus, C. gregii, or C. megacarpus. Data taken from references 2, 3, and 8 reported multiple species and data reported here are means. B. This is water stress–induced cavitation estimated as the water potential causing 50% loss in hydraulic conductivity. C. This is the maximum stem hydraulic conductivity (in the absence of emboli) divided by the sapwood area (xylem specific conductivity).

Stress tolerators diverge from the avoiders in a suite of traits that allow them to maintain a broader range of physiological function at more negative water potentials (Table 24.7). Key among these is greater resistance to water stress–​induced xylem cavitation (Kolb and Davis 1994, Davis, Kolb, and Barton 1988, Davis et al. 2002). Cavitation describes the process by which air is pulled into xylem conduits and displaces water with air emboli (Tyree and Sperry 1989). When this happens, emboli reduce flow (hydraulic conductivity) through the vascular system. If emboli occur in many conduits, a positive feedback loop can ensue, leading to runaway cavitation and desiccation of leaves, dieback of branches, or whole plant mortality (Davis et al. 2002, Paddock et al. 2013, Pratt et al. 2008). Species that experience more negative minimum seasonal water potentials have greater cavitation resistance (Davis, Kolb, and Barton 1988; Davis, Ewers et al. 1999; Bhaskar et al. 2007; Jacobsen et al. 2007a; Pratt, Jacobsen, Ewers et al. 2007) ( Figure 24.6). Moreover, greater cavitation resistance is correlated with greater survival of drought in chaparral seedlings (Pratt et al. 2008) In mixed stands, cooccurring species often have widely divergent cavitation resistances that reflect differences in access to soil moisture during the dry season and selection that occurs at the seedling stage and during episodic drought (Thomas and Davis 1989, Davis, Kolb, and Barton 1988, Jacobsen et al. 2008). The only way for plants to recover hydraulic conductivity following cavitation is to refill conduits or regrow new xylem, both of which cannot occur during the protracted dry season when water potentials are low (Kolb and Davis 1994, Williams et al. 1997). Thus, avoiding cavitation by resisting it is the most effective strategy for evergreen shrubs whose leaves 492  Ecosystems

require a sustained water supply. Evergreen chaparral shrubs do have greater cavitation resistance of stems than deciduous shrubs that occur in the chaparral community (Figure 24.7). However, deciduous species are not necessarily strict drought avoiders (Gill and Mahall 1986, Jacobsen et al. 2008). Cavitation and reduced hydraulic conductivity can occur in roots, stems, or leaves, and a process similar to xylem cavitation can occur in the rhizosphere. Among stress tolerators, the rhizosphere’s hydraulic conductivity may be most limiting during the dry season because of the high resistance of stems and roots (Davis et al. 2002; Pratt, Jacobsen, Golgotiu et al. 2007). This allows stress tolerators to extract the maximum amount of water from a limited volume of drying soil and also to rapidly resume water and nutrient uptake when winter rain falls (Poole and Miller 1975; Mooney and Rundel 1979; Gill and Mahall 1986; Pratt, Jacobsen, Golgotiu et al. 2007). At least two other traits are important for stress tolerators to maintain physiological function. The leaves of stress tolerators close their stomata at more negative water potentials than the stress avoiders, allowing them to continue to photosynthesize at greater water deficits (Poole and Miller 1975, Miller and Poole 1979, Oechel et al. 1981, Jacobsen et al. 2008). They also have lower turgor loss points than stress avoiders (Roberts 1982). In spite of a lower turgor loss point, stress tolerators lose turgor during the dry season, so it is not avoidance of turgor loss that is important but maintenance of turgor at more negative water potentials and tolerance of turgor loss (Saruwatari and Davis 1989). Additional traits are associated with the stress-tolerator strategy (see Table 24.7). Stem vascular tissue is denser and mechanically stronger (Wagner et al. 1998, Jacobsen et al.

0

0

A

–2

–4

P50 (MPa)

P50 (MPa)

–2

–6 –8 r2 = 0.75 P < 0.001 slope = 1.33

–10 –12

–8

–6

–4

Seedling mortality due to water stress (%)

–4

–6

–8

–2

–10

Minimum seasonal water potential (MPa)

Evergreen

80 70

P = 0.05

Deciduous

Leaf habit

B

FIGURE 24.7 Stem resistance to cavitation estimated as the water potential at which 50% of conductivity is lost (P50, n=16 for deciduous and n=18 for evergreen). Data are from chaparral shrubs growing at four field sites located in the San Gabriel, Santa Monica, San Bernardino, and the San Jacinto Mountains. Source: R. Brandon Pratt, unpublished data.

60 50 40 30

r2 = 0.80 P = 0.002 slope = 11.51

20 10 –12

–11

–10

–9

–8

–7

P90 (MPa)

FIGURE 24.6 Species that experience more negative minimum seasonal water potentials have greater cavitation resistance. Source: Data modified from Jacobsen et al. 2007a and Pratt et al. 2008. A Stem xylem cavitation resistance estimated as the water potential

at 50% loss of hydraulic conductivity (P50) plotted against the seasonal low in water potential measured during the peak of the summer rainless period for adult chaparral shrubs. B Greater cavitation resistance is associated with lower levels of

mortality due to water stress for seedlings.

2005, Jacobsen et al. 2007a). The vessels of stress tolerators are more resistant to implosion when the xylem is under tension (Jacobsen et al. 2007a; Pratt, Jacobsen, Ewers et al. 2007) and have narrower diameter vessels (Jacobsen et al. 2007a). Shoots of tolerators have lower leaf area per unit sapwood area of their branches (Ackerly 2004b). Finally, the stem xylem of stress tolerators store less carbohydrate than the stress avoiders during the dry season (see Table 24.7). This is hypothesized to be due to the ability of stress tolerators to photosynthesize at more negative water potentials and rely less on stored carbohydrates than stress avoiders (McDowell et al. 2008). Species dominating the more arid sage scrub and Mojave Desert communities are less resistant to water stress–​induced xylem cavitation than chaparral communities (Jacobsen et al. 2007b), even though species in these communities do not differ in minimum seasonal water potential they experience (Jacobsen et al. 2008). The costs associated with greater cavitation resistance may limit cavitation resistance in these species (Pratt, Jacobsen, Ewers et al. 2007). Leaf shedding during

dry periods, common among species in the more arid communities, may mitigate water stress compared to evergreen chaparral. However, drought-deciduous species are not necessarily water stress avoiders (Gill and Mahall 1986, Kolb and Davis 1994, Jacobsen et al. 2008). At a Sonoran Desert/ chaparral ecotone, some desert species—​for example, jujube (Ziziphus parryi)—​a re fully deciduous, quite vulnerable to cavitation, have low water-storage capacity, and experience highly negative water potentials during the dry season, suggesting that their roots remain in contact with dry soils. Perhaps species with this trait combination are able respond to seasonal rains or rain pulses by rapidly refilling xylem conduits and growing new high SLA leaves.

Freezing and Distribution Freezing temperatures can limit the distribution of frost-sensitive species at higher elevations and at low-elevation basins that fill with cold air. The coldest air temperatures are commonly caused by radiation frosts on calm and clear nights in the months of December and January (Davis et al. 2007). At freezing ecotones, species turnover can be abrupt and the turnover occurs where the minimum air temperatures drop to levels that cause damage to sensitive species (Ewers et al. 2003, Davis et al. 2007). In the Santa Monica Mountains of southern California, laurel sumac and ceanothus species dominate on coastal exposures and warmer sites outside cold air drainage zones. On the lower slopes and the valley floors these species are replaced by other more frost-tolerant ­species: sugar bush and other species of ceanothus (Davis et al. 2007). Some cold-sensitive species like laurel sumac suffer direct damage to their living leaf cells by freezing temperatures (Boorse et al. 1998, Pratt et al. 2005) and suffer freeze/thawinduced cavitation (Langan et al. 1997). The mechanism of cavitation caused by freeze/thaw-induced cavitation is different from that caused by water stress (Jarbeau et al. 1995), but they both result in emboli that reduce xylem hydraulic conductivity. The emboli typically form in the distal branches Chaparr al  493

100

Total survival after fire (%)

and can lead to desiccation of leaves in the days and weeks that follow the frost (Langan et al. 1997, Pratt et al. 2005, Davis et al. 2007). Overnight freezing temperatures are frequently followed by warm sunny days, thus the combination of evergreen leaves and a highly embolized vascular system can lead to dieback of branches even during the moist winter (Davis et al. 2005). For a plant under water stress, freeze/ thaw-induced cavitation leads to more extensive formation of emboli, dieback, and plant mortality (Langan et al. 1997, Davis et al. 2005, Davis et al. 2007). Species with largerdiameter vessels are more vulnerable to freeze/thaw-induced embolism (Davis, Sperry et al. 1999).

80 60 40 20 0 0

The Link between Life History Type and Physiology Recurring crown fires are important for understanding the physiology of chaparral shrubs, particularly at the critical seedling stage. An important framework for understanding the nexus between fire and physiology is to consider the different chaparral life history types. The three different life history types recruit seedlings in different environments, and that recruitment environment selects for different physiologies (Keeley 1998, Pratt et al. 2012). This coupled with physiological and allocation trade-offs leads to divergence in the suite of functional traits of the different life history types. Obligate seeders recruit seedlings in the most arid and open canopy microsites. Consistent with this, obligate seeders are more water stress tolerant and the least shade tolerant than the other life history types (Pratt et al. 2008). Facultative seeders recruit seedlings in open environments, but have greater survival in shadier or moister microsites where stress is ameliorated (Pratt et al. 2008). Accordingly, these facultative seeders have lower water stress tolerance and greater mortality of seedlings during the first dry season following fire (Thomas and Davis 1989). Obligate resprouter seedlings recruit during fire-free intervals in the shady understory of mature chaparral canopies (Keeley 1992). These species have the highest degree of shade tolerance, and they are of similar or greater water stress tolerance to the facultative seeders, which may be related to competition for water in a mature chaparral stand (Pratt et al. 2008). This divergence in seedling recruitment environments (Figure 24.8) is linked to other characteristics as well, such as differences in rooting patterns (Thomas and Davis 1989, Keeley 1998). Obligate seeders are commonly shallow-rooted, whereas facultative seeders are variable and some are relatively shallow-rooted (e.g., some populations of chamise and manzanita), yet others are among the most deeply rooted (e.g., laurel sumac). Instructively, the more shallowly rooted facultative seeders tend to be the ones that recruit more seedlings after fire and often resprout more weakly (e.g., some chamise populations)—​that is, they trend towards the obligate seeder end of the spectrum. Successful seedling recruitment in gaps after fire is achieved by both the ability to rapidly acquire resources when they are abundant after fire and by greater levels of stress tolerance when water is limiting. Evolving this ability and the associated necessary traits appears to compromise resprouting ability (Pratt et al. 2014). Stress tolerators have trait combinations that include both stress-tolerance traits and rapid-resource acquisition traits such as high net carbon assimilation rates per unit leaf area (Ackerly 2004b, Pratt et al. 2012), transpiration rates (Parker 1984), and hydraulic efficiency at the whole plant level (Pratt 494  Ecosystems

2

4

6

Seedling-to-parent ratio FIGURE 24.8 Greater seedling production (seedling-to-parent ratio) is significantly associated with lower levels of resprout survival two years after fire (r= –​0.81; P=0.03). All species co-occurred at a Santa Monica Mountains site. Source: Pratt et al. 2014.

et al. 2010). This is likely because during the wet season in the postfire environment, water and nutrients are readily available and seedlings must compete for these resources with other seedlings, including those of herbaceous fire-followers, and resprouting shrubs (McPherson and Muller 1967). At this time, they also must grow fast enough to establish a root system that will prevent them from desiccating during the hot and dry summer and fall. Once the summer and fall season arrives, seedlings experience considerable stress and mortality is often high at this time (Thomas and Davis 1989). These circumstances select for seedlings to rapidly grow and acquire resources when they are available and also to tolerate low levels of resources as they become scarce (Keeley 1998).

Chaparral in the Long Absence of Fire Because chaparral ranges from arid southern California to more mesic maritime regions and northern California mountain ranges, chaparral dynamics similarly can vary. Thus the phrase “the long absence of fire” has to be defined in the local climatic context. We can reiterate that climate, topography, soils, and wildfire are the principal determinants of the distribution of chaparral across the California landscape. As these principal influences vary among sites and when the betweenfire period increases, chaparral dynamics might be expected to change. Earlier studies of chaparral in southern California led to a paradigm that little succession occurred in chaparral and that stands older than sixty years were decadent (Hanes 1971). The idea was that stands reached some age at which plants began to senesce, productivity dropped, and no seedling establishment occurred (Hanes 1971, Hanes 1977, Vogl 1977, Reid and Oechel 1984). Ceanothus species, especially from subsection Cerastes, were thought to live only thirtyfive to fifty years at best. Succession was seen as the eclipsing of coastal sage scrub species by evergreen sclerophyllous species in the first few decades. While some thought that woodland might eventually invade and convert the site if fire was excluded (Horton and Kraebel 1955, Horton 1960, Wells 1962), others thought no further succession would occur (Hanes 1971). More detailed studies shifted the concepts of chaparral dynamics in the long absence of fire. Old-growth stands of big-pod ceanothus were found to have dead individuals

within them, but the larger living plants were still growing and mortality patterns suggested little change in the stand as it aged (Montygierd-Loyba and Keeley 1987). When a large number of sites were compared along an age gradient, species diversity was found to be stable although some composition change occurs as stands age (Keeley 1992). The greatest mortality rates were among species that fail to grow tall fast enough as they age, such as nonsprouting ceanothus species; other evergreen species showed almost no mortality across a diverse latitudinal and elevation gradient (Keeley 1992). One significant difference between these studies and the earlier paradigm was the evidence that obligate resprouters were establishing seedling banks in the understory of these older stands at fairly high densities (0.1-3.7 m-2). This work confirmed and extended earlier studies reporting the establishment of obligate resprouter seedlings in chaparral understory (P. H. Zedler 1981, P. A. Zedler 1982, Parker and Kelly 1989, Lloret and Zedler 1991). These seedling patterns in southern California chaparral tend to reinforce the life history categories found among the woody plants, those with persistent seed banks recruiting new individuals in the year or two after fire (obligate seeders, facultative seeders), while obligate resprouters require longer time intervals for their shade-adapted seedlings to establish in the understory of chaparral. Succession beyond these demographic shifts depends upon the proximity of other vegetation to old-growth chaparral. Conifers have limited dispersal ranges, and if chaparral stands are quite extensive, initially they have little ability to invade beyond the edges. Similarly, oaks and other hardwoods may benefit from animal dispersal in terms of farther potential dispersal distances, but establishment patterns remain low because of the low productivity and resource limitations of chaparral in this part of the state. Moving toward central and northern California or shifting to higher elevations into montane chaparral yields a different pattern. In these regions chaparral forms mosaics with forests with the sizes of each type of vegetation patch depending on local climate, topography, soils, and wildfire history. In areas near urban sites or where forest timber is economically critical, fire suppression has generally been the practice for the past century. As a consequence, chaparral stands often are invaded by conifers or other trees and the extent of chaparral may be quite reduced (Vankat and Major 1978, Conard and Radosevich 1982, Nagel and Taylor 2005, Collins and Stephens 2010, van Wagtendonk et al. 2012). One study in the Sierra reported that over the past sixty to seventy years, comparisons of aerial photographs reveal a loss of chaparral spatial coverage by over 60% (Nagel and Taylor 2005). Conifer invasion of chaparral does not always require long-term lack of fire in northern California, however, where, if conditions are appropriate, conifers may invade immediately after fire (Sparling 1994, Horton et al. 1998). For example, following a 1945 wildfire, one chaparral stand adjacent to a forest was invaded the first growing season after the fire, followed by pulses of invasion over the next thirty years (Sparling 1994, Horton et al. 1998). The conifer forest–​ montane chaparral mosaics have received considerable attention because of the economic importance of the forests. Studies indicate that forest and chaparral burn at different time intervals and at substantially different fire intensities; forests burn at relatively low intensities while chaparral fires are high intensity. Early researchers considered forest and chaparral patches as alternative vegetation states across the landscape, each maintaining itself due

to their different interactions with fire (Leiberg 1902, Show and Kotok 1924, Wilken 1967). Similar patterns are found in other Mediterranean-climate systems (Mount 1964, Jackson 1968). The differential distribution of these two vegetation types contributes to the self-reinforcing nature of their interaction with fire. High-intensity wildfires tend to occur on steeper or more south-facing aspects in the northern part of the state (Weatherspoon and Skinner 1995, Taylor and Skinner 1998, Alexander et al. 2006). The abiotic conditions combined with further resource limitations imposed by chaparral reduce the rate of conifer invasion and productivity. The cumulative effects of multiple high-intensity fires reduces soil carbon and site productivity, further slowing tree growth rates (Waring and Schlesinger 1985). Such differential impacts by chaparral and forest were modeled by Odion et al. (2010) to test the concept of alternative vegetation states. Their models supported co-occurring but different stands of vegetation due to different selfreinforcing relationships with fire. Chaparral vegetation burned at higher intensity, particularly on steep or equatorial-facing slopes, and its fire regime kept soil productivity from increasing through time. Forests maintain themselves in sites with gentler slopes and greater soil depth and tend to dominate polar-facing slopes. Their conclusion was that conifer invasion of chaparral sites resulted in conifers being in a “fire trap.” Chaparral sites burned at high enough intensity to remove conifers that had established since the last fire. This pattern has been found frequently in the western United States (e.g., Thompson et al. 2007, Holden et al. 2010). Focusing on the central Sierra Nevada, van Wagtendonk et al. (2012) found that these different vegetation patches were maintained as long as fire intensities were low to moderate; high fire intensities resulted in dominance of montane chaparral regardless of the prior vegetation. Another aspect to the dynamics between chaparral and forest patches is that some types of chaparral create conditions favorable to the facilitation of forest establishment. For example, most conifer seedlings benefit from chaparral canopies providing shade and higher levels of moisture, even though their growth rates may be restricted by resource competition (Conard and Radosevich 1982, Dunne and Parker 1999). Coast live oak, for example, is facilitated by chaparral and coastal scrub (Callaway and D’Antonio 1991) and populations differentially expand within shrub-dominated areas in central California (Callaway and Davis 1998). Furthermore, in the more mesic regions of chaparral distributions in which chaparral exists in mosaics with forests, the dominants of chaparral tend to be species of manzanita or shrub oak, both of which have mycorrhizal mutualisms with fungal species that are shared with conifers in the Pinaceae (Horton et al. 1998). The mycelial networks of fungal mutualists facilitate the establishment of conifer seedlings in conditions that otherwise prevent establishment (Dunne and Parker 1999, Horton et al. 1998). A study testing this idea randomly sampled chaparral stands from northern and central California that were adjacent to forests; chaparral was invaded by conifers in those stands at a rate and density proportional to the percentage of ectomycorrhizal species in the chaparral stand (Bode 1999). In the long absence of fire, chaparral begins to shift demographically; early successional species including those that may be prominent in coastal or sage scrub vegetation tend to disappear in the first few years to the first decade. As individuals increase in size, compositional shifts or proportions Chaparr al  495

TA B L E 2 4 . 8 Adaptive traits for nutrient-poor soils

Traits

Function

References

Root nodulesA

Symbioses with bacteria that fix atmospheric nitrogen

Pratt, Jacobsen, Ewers, et al. 2007 Kummerow, Alexander, et al. 1978

Long-lived leaves

Greater nitrogen use efficiency

Aerts and Van der Peijl 1993

Evergreen sclerophyllous leaves

Storae compartment for nutrients captured during nutrient pulse following first fall rains

Mooney and Rundel 1979, Shaver 1981

Luxury consumption

Consume nutrients beyond immediate need and store for later use

Rundel and Parsons 1980, Gray 1983

Mychorrizae

Aid in phosphorous extraction and uptake

Allen et al. 1999

A. Examples include Ceanothus spp., Cercocarpus betuloides, and legumes such as Acmispon glaber (formerly Lotus scoparius).

of dominance also change. In the southern ranges of chaparral, fire usually returns prior to any further shifts; however, in more mesic regions, chaparral dynamics involve adjacent vegetation. Some chaparral species, for example, may facilitate forest invasion because of their shared mutualisms with mycorrhizal fungi and the modification of environmental extremes that chaparral canopies may provide. Wildfires balance this process. Chaparral sites burn at an intensity too high for conifers or most forest species to survive, creating fire traps that provide maintenance of chaparral stands over the long term. These changes in chaparral result from differences among chaparral species in their rates of growth, height, or other aspects. While shared mutualisms with forest species may facilitate invasion of chaparral, wildfire, climate, topography, and soil conditions generate the conditions that retain chaparral as a dominant vegetation type in California.

Biogeochemical and Hydrological Dynamics Carbon Exchange Community productivity indicates the amount of energy captured and converted to biomass and has important implications for trophic interactions and biotic diversity. At the community scale, chaparral stands can be quite productive annually. For example, stands dominated by big-pod ceanothus have productivity rates of 850 g m-2 yr-1 (Schlesinger and Gill 1980) and have been reported to be as high as 1,056 g m-2 yr-1 (Gray 1982). The higher value is comparable to some temperate forest communities and is generally higher than nearby coastal sage communities (Schlesinger and Gill 1980, Gray 1982). Other studies have documented productivity values somewhat lower to much lower at more arid chaparral sites where chamise is abundant (Rundel and Parsons 1979, Vourlitis et al. 2009). Productivity broadly varies on an annual basis depending on temperature and seasonal rainfall (Hellmers, Bonner et al. 1955; Li et al. 2006; Vourlitis et al. 2009). The net ecosystem exchange of CO2 of old-growth chaparral stands indicates that they can be substantial carbon sinks (Luo et al. 2007). This suggests that gross primary productivity can remain high in these communities and does not support the idea that the productivity of chaparral stands declines as they age (Hanes 1971). The amount of CO2 that old-growth stands accumulate or give off annually depends 496  Ecosystems

on current and previous year precipitation and environmental conditions (Luo et al. 2007) and the interaction between the time since the last fire and available nutrients (see “Mineral Nutrition,” below).

Mineral Nutrition Unburned chaparral soils are poor in available nutrients, and chaparral shrubs have a host of traits that are adaptive in this context (Table 24.8). Nitrogen, and in some cases phosphorous, are in limiting supply, leading to reduced levels of productivity, whereas exchangeable cations are of secondary importance (Rundel 1983). When long unburned plots are fertilized with nitrogen, growth of plants is generally stimulated, which demonstrates the influence of nitrogen on productivity (Hellmers, Bonner et al. 1955; McMaster et al. 1982; Vourlitis 2012). At the system level, most of the nitrogen exists in unavailable standing biomass and litter (Rundel 1983). This reservoir is converted by periodic crown fires and results in a pulse of available nitrogen to the soil (Christensen 1973, Christensen and Muller 1975, Rundel and Parsons 1980, Vourlitis et al. 2009). Immediately after fire, the ammonium significantly increases in the upper soil layers, whereas nitrate is little affected (DeBano, Eberlein et al. 1979; Rundel 1983; Fenn et al. 1993). In the months after fire, the soil pH increases towards neutral and this stimulates nitrification, leading to an increase in soil nitrate (Dunn et al. 1979, Rundel 1983). Although fires greatly increase available forms of nitrogen in the soil, nitrogen is volatilized by fire, resulting in significant losses from the system (DeBano and Conrad 1978; DeBano, Eberlein et al. 1979; Rundel 1983). The intensity of the fire is an important determinant of how much nitrogen is volatilized (DeBano, Rice et al. 1979; Marion et al. 1991). Other postfire losses include leaching, surface runoff, dry erosion, and biogenic emissions (Christensen 1973, DeBano and Conrad 1978, Rundel and Parsons 1980, Gray and Schlesinger 1981, DeBano and Dunn 1982, Rundel 1983, Mooney et al. 1987). Losses would be greater if nutrients were not immobilized by a dense growth of fire-following annual herbs and short-lived shrubs such as deer weed (Acmispon glaber, formerly Lotus scoparius) (Nilsen and Schlesinger 1981). After successive fires, chaparral soils would eventually be depleted of nitrogen without new inputs. Atmospheric nitro-

4000

Aboveground biomass (g m-2)

gen fixation by microorganisms that are free-living and symbiotic with plants are an important new nitrogen input that offsets some losses (Kummerow, Alexander et al. 1978; Dunn et al. 1979; Ellis and Kummerow 1989; Ulery et al. 1995). Other smaller nitrogen inputs come from precipitation and dry deposition of nutrients on leaves, which reach the soil in solution as a pulse following the first fall rains (Christensen 1973, Schlesinger and Hasey 1980). These atmospheric inputs have increased due to pollution caused by the burning of fossil fuels (see below). Decomposition converts complex organic substances into simpler forms, and it affects nitrogen availability in between fires. Decomposition by soil animals and microorganisms occurs primarily during the wet season and at a much lower level during the summer dry season (Quideau et al. 2005, Li et al. 2006). Rates are also affected by local conditions such as soil texture and type, slope, and elevation, and rates differ under different evergreen species (Quideau et al. 1998, Quideau et al. 2005). The chief litter input is from leaves, thus their nutrient content and turnover rate are dominant factors in nutrient cycling and are key reasons why the litter of a given species affects decomposition. Much of the nitrogen and phosphorous of leaves are reabsorbed prior to abscission, thus the nutrient quality of the litter in chaparral stands is poor (Mooney and Rundel 1979, Schlesinger 1985, Quideau et al. 2005). The poor quality of the litter (high C:N and C:P ratios) slows decomposition (Schlesinger 1985). Moreover, decomposition is slowed by the high levels of carbon forms that are resistant to decomposition, such as lignin, cutin, and phenolics, which are typically higher in evergreen sclerophylls than deciduous species (Schlesinger 1985, Aerts 1997). Decomposition of leaves occurs in three phases: first is the leaching of soluble components (K, Mg, carbohydrates, phenolics) during rainstorms (Schlesinger and Hasey 1981). This phase is rapid (greater than one year) and is determined by the amount of soluble components in the litter. The second phase is slower (approximately one to five years) and results from litter fragmentation and slow release initially of nonsoluble components (e.g., N, P, Ca, and lignin) upon breakdown by soil microbes. Turnover rates of litter organic matter during this phase ranges from 2.8 to 7.7 years, with litter bag studies tending to yield higher values than mass balance methods (Schlesinger 1985). A three-year study found that the nitrogen and phosphorous becomes immobilized in the litter layer, making it unavailable over this time period (Schlesinger 1985). A third, slower decomposition phase occurs when organic matter is mixed with mineral soil and chemically altered and breakdown products are leached (Chapin et al. 2002). Studies of this phase are lacking, but there have been studies of nutrients across chronosequences that provide some insight. Total nitrogen and potentially mineralizable nitrogen decline in stands older than fifty years (Marion and Black 1988). In some systems, nitrogen is immobilized in microbial products; however, this does not appear to be the case in chaparral topsoil (Fenn et al. 1993). The decline of available soil nitrogen as stands age occurs because the amount sequestered into biomass and recalcitrant soil compounds is greater than inputs. The tendency for stands to become nutrient poor in the decades following fire, particularly in nitrogen, has been suggested to lead to declines in productivity and senescence of plants in old stands (Rundel and Parsons 1980, Marion and Black 1988; however, see Fenn et al. 1993) (Figure 24.9). Many aspects of decomposition are poorly understood in the chaparral. One in particular is phase three dynamics

North-facing South-facing

3500 3000 2500 2000 1500 1000 500 0

20

40

60

80

100

Age of stand (years) FIGURE 24.9 Relationship between aboveground biomass of shrubs on north- and south-facing slopes and age of chaparral. Source: Data modified from Table 24.1 in Marion and Black 1988. To simplify chronosequence, same-age sites are averaged.

and the controls over longer-term decomposition processes. Another area that is little studied is decomposition in the rhizosphere. Fine root mass has significant turnover in the top 10 to 20 centimeters of soil on a seasonal basis (Kummerow, Krause et al. 1978). Many fine roots in these shallow soil layers die during the summer dry season and then presumably decompose during the following wet season (Kummerow, Krause et al. 1978). Phosphorous may be a limiting nutrient for some species in long unburned stands as suggested by fertilization studies (Hellmers, Bonner et al. 1955; McMaster et al. 1982). In particular, species that form symbioses with nitrogen-fixing bacteria may have adequate nitrogen but may become phosphorous-limited (Schlesinger and Gill 1980, McMaster et al. 1982, Schlesinger 1985). The majority of phosphorous in chaparral systems is in unavailable forms in the soil (DeBano and Conrad 1978, Rundel 1983). Like nitrogen, but to a lesser degree, available phosphorous increases following fire (Christensen and Muller 1975, Marion and Black 1988). Unlike nitrogen, little phosphorous is lost from the system after fire (DeBano and Conrad 1978, Marion and Black 1988). Phosphorous inputs are from weathering, inorganic solutes in precipitation, dry deposition, and, to a lesser degree, decomposition (Marion and Black 1988). Available phosphorous in older stands (older than fifty years since the last fire) is low and availability is determined by absorption and desorption reactions between soluble phosphorous and soil surfaces (Marion and Black 1988). Chaparral species form associations with mycorrhizal fungi that likely aids in extraction of phosphorous from soils (Allen et al. 1999). Burning of fossil fuels has led to the production of numerous air pollutants that contain nitrogen (Bytnerowicz and Fenn 1996). Some of these pollutants are dry deposited on chaparral canopies and are then leached to the soil during rains. Areas that have fog may also experience significant wet deposits of nitrogen. An area especially affected is the Los Angeles Air Basin. Because chaparral communities are nitrogen-limited, deposition can lead to higher rates of net productivity (Vourlitis 2012). In areas of high and chronic deposition, the system may become nitrogen-saturated, as has been documented in the portions of the San Gabriel and San Bernardino Mountains (Fenn et al. 1998). Compared to unsaturated systems, nitrogen-saturated systems have elevated leaf Chaparr al  497

and tissue nitrogen content, nitrate losses from streams, NO loss from the soil, N mineralization and litter decomposition, and lowered pH and base saturation of soils (Fenn et al. 1996, Vourlitis and Fernandez 2012).

Hydrology Studies have examined the precipitation inputs and losses of chaparral watersheds at the San Dimas Experimental forest in the San Gabriel Mountains in Los Angeles County and at Echo Valley in San Diego (Hamilton and Rowe 1949, Rowe and Colman 1951, Poole et al. 1981). The chief input of precipitation is about seven to fifteen rainstorms during the winter wet season (Cowling et al. 2005). About 78% to 80% of precipitation reaches the soil as throughfall, which falls directly to the soil or drips off the canopy to the soil (Hamilton and Rowe 1949, Poole et al. 1981), and another fraction of water reaches the soil by flowing down stems (Poole et al. 1981). Stemflow is affected by canopy architecture, and species with erect branches and smooth bark have greater stemflow (Hamilton and Rowe 1949). In one study, chamise had greater stemflows than other co-occurring chaparral shrubs (Poole et al. 1981). Losses of precipitation occur via canopy interception, surface runoff, evaporation from the soil, and transpiration. Water may be lost through subsurface drainage into fractures in the soil depending on site, and such losses are greater in heavier rainfall years (Rowe and Colman 1951, Hill 1963, Ng and Miller 1980, Poole et al. 1981). Some precipitation intercepted by the canopy evaporates from plant surfaces. The amount of loss from the canopy is determined by the number of rainfall events, the intensity of storms, and the canopy architecture (Poole et al. 1981). Previous studies have found that 5–​41% of precipitation is lost due to canopy interception with average losses higher in southern California (Hill 1963, Hill and Rice 1963, Poole et al. 1981). Greater losses may occur when a storm occurs in the warm summer, when canopies have higher levels of leaf area, and when precipitation falls as snow (Poole et al. 1981). Greater numbers of smaller storms leads to greater losses from the canopy (Hamilton and Rowe 1949). With needle leaves and erect branches, chamise has low levels of canopy interception (Poole et al. 1981), whereas species with more leaf surface area incept more precipitation (Poole et al. 1981). Surface runoff is the difference between precipitation inputs, soil storage, and evapotranspiration. The amount of precipitation that makes it into the soil storage is dependent on frequency and size of rainfall events, slope, postfire factors, as well as soil structure and chemistry. Soil structure is partially affected by species composition, as is stemflow, thus species-level effects can influence infiltration of precipitation. Surface runoff is generally low but can be high during high rainfall years (Rowe and Colman 1951, Meixner and Wohlgemuth 2003). At a site in San Diego County, surface runoff in average and low rainfall years was low (0–​4%) with the greater runoff on north-facing slopes (Ng and Miller 1980, Poole et al. 1981). Farther north in Los Angeles County, runoff of a chaparral watershed has been estimated by streamflow, and the fifteen-year average of streamflow yield was 11% of the total rainfall input (Hill and Rice 1963). During the wettest year, streamflow yield was higher at 21% and during the driest year streamflow yield was trace. More direct smaller-scale measurements of runoff have been made at this site in large 498  Ecosystems

lysimeter studies (Patric 1961), which found about 14.5% of

precipitation inputs was lost as runoff averaged across five consecutive dry years, and 40% was lost during a wet year (Hill and Rice 1963). Over the short-term, crown fire leads to large increases in surface runoff and streamflow (Meixner and Wohlgemuth 2003). This effect is due to a decline in evapotranspiration and the formation of hydrophobic soils after fire (DeBano et al. 1977, Valeron and Meixner 2010). The chief loss of water is due to evapotranspiration. In lysimeter studies virtually all of the water that enters the soil is lost to evapotranspiration during below-average rainfall years, and more than 80% of it is lost during years when rainfall is substantially above average (Patric 1961). A chamise-dominated south-facing slope can show 80% loss due to evapotranspiration, whereas a north-facing slope of mixed chaparral with greater leaf area can lose virtually all of the precipitation input to evapotranspiration (Ng and Miller 1980). Estimates of the separate contribution of evaporation and transpiration have been made in a chamise-dominated stand, in which the amount of water transpired is about equal to the amount evaporated (Poole et al. 1981). However, in mixed chaparral stands with greater canopy cover, transpiration is about three times greater than evaporation (Poole et al. 1981). Hydraulic redistribution refers to the movement of water by plant roots from one soil compartment to another down a water potential gradient. Such water fluxes may be an important factor affecting the hydrology of chaparral sites, particularly in its effects on evapotranspiration. Studies of hydraulic redistribution are currently lacking for chaparral sites.

Ecosystem Services Ecosystem services describe the ways that ecosystems benefit people. Such services can be categorized as regulating (e.g., climate, flooding), provisioning (e.g., food, fuel, fresh water), supporting (e.g., nutrient cycling and carbon sequestration), and cultural (e.g., aesthetic, educational, recreation) (Millennium Ecosystem Assessment 2005). Chaparral provides services in each of these categories. For regulating services, chaparral vegetation absorbs sunlight and transpires water—​both of which help to regulate temperature during the hot summer months compared to highly urbanized areas that experience the “heat island effect” (LaDochy et al. 2007). The growth of shrublands on steep hillsides helps to reduce flooding, erosion, and mudslides that can occur during heavy rains that commonly occur each winter (Gabet and Dunne 2002). This service is especially apparent after chaparral has been removed by fire, and heavy winter rains cause costly and lethal mudslides (Ren et al. 2011). Provisioning services of chaparral includes filtration of rainwater, which helps to maintain fresh drinking water in aquifers and reduce eutrophication in the ocean and reservoirs that receive runoff. This is important in some areas where nitrogen deposition is high and nitrogen pollutants such as nitrate might be more prone to leach into groundwater supplies and collect in downstream bodies of water. Areas that were formerly chaparral that have been converted to grassland are not as effective at filtering water and they yield greater nitrate runoff (Riggan et al. 1985). In areas where nitrogen pollution is the most severe, such as some watersheds in the San Bernardino and San Gabriel Mountains northeast of Los Angeles, runoff of nitrate is among the highest in the United States (Fenn and Poth 1999).

Supporting services include carbon sequestration with stands of chaparral, even very old ones, acting as carbon sinks (Luo et al. 2007). The pollination services provided by native bees are associated with the amount of nearby natural habitat (including chaparral) where these bees reside (Kremen et al. 2004). As discussed already, chaparral vegetation has a substantial impact on the hydrology of a watershed. Studies conducted in the Transverse mountain range at the San Dimas Experimental Station manipulated the vegetation in an effort to increase usable water. Chaparral and riparian vegetation was removed from a portion of the watershed and this increased water yield from the watershed (Hill and Rice 1963, Meixner and Wohlgemuth 2003). This result confirms many studies that removing chaparral or any vegetation type that is deeply rooted reduces transpiration of the system. Increases in water yield by removing chaparral are likely to come at the expense of water quality, a reduction in temperature regulation services, and a loss in cultural services. Cultural services provided by chaparral systems are high. Chaparral systems are located in some of the largest metropolitan areas in North America such as Los Angeles. This ensures that there are many millions of visitors to these shrublands for recreation activities such as hiking, biking, horse riding, and camping. The presence of chaparral on the low-elevation slopes beautifies the landscape. The educational impact of chaparral is high due to the large numbers of parks, colleges and universities, and organizations that operate near chaparral systems. For example, the Santa Monica Mountains National Recreation Area is a park in southern California that receives about thirty-five million visitors annually with outreach and educational programs that target people of all ages.

The Future of Chaparral Predictions are that the climate in California will be increasingly warmer and drier in the coming decades (Hayhoe et al. 2004); however, there is uncertainty in this prediction because rainfall may increase in some regions of California (Neelin et al. 2013). The water deficits for chaparral shrublands will depend on interplay between temperature, the amount and timing of rainfall, and local soil water storage dynamics. Of paramount importance will be how the changing climate affects extreme weather patterns (maximum and minimum temperatures and drought intensity) and wildfire (Westerling et al. 2006). Extreme events such as the record droughts since 2012 and heat waves will likely have direct impacts on chaparral communities. These effects are already evident at the arid ecotones where chaparral mixes with desert scrub communities and some adult chaparral species have experienced significant levels of mortality (Paddock et al. 2013). Increases in minimum nighttime temperature (Crimmins et al. 2011) will affect species whose current distributions are limited by subzero temperatures (Ewers et al. 2003, Davis et al. 2005, Davis et al. 2007). Higher nighttime temperatures may lead to greater rates of respiration with implications for carbon balance. Climate will likely interact with fire to drive change. For example, a recent study found that postfire resprouts of some species suffered high levels of mortality in the first year after a fire during an intense drought, whereas adjacent unburned plants did not suffer mortality (Pratt et al. 2014). One of the most immediate and devastating affects to chaparral communities is alteration of the fire regime. Chaparral

stands are generally not resilient to short fire-return intervals less than about fifteen to twenty years (Zedler et al. 1983, Jacobsen et al. 2004, Keeley et al. 2005b). Such short return intervals have become more common due to anthropogenic ignitions and the abundant fine fuels produced by annual alien grasses in disturbed areas (Brooks et al. 2004). If short fire-return interval fires continue or increase in the future, increasing areas of chaparral communities will be converted into homogenous savannahs where only the most vigorous resprouters persist (e.g., laurel sumac in southern California). Other aspects of climate change, such as increasing CO2, may increase water-use efficiency and alter patterns of fuel moisture in ways that potentially could offset increasing fire ignition hazard due to warmer temperatures (Oechel et al. 1995); however, CO2 may also stimulate biomass accumulation and lead to an increase in fuels and high-intensity fires. Increased CO2 will also subtly shift the importance of the less well-known microbial communities as already low levels of nutrients may become more limiting and cascade along food webs (Oechel et al. 1995). Because of its presence at or near the boundaries of urban developments and metropolitan centers, conflicts between the impacts of chaparral wildfire and human life and structures likely will increase without intelligent regional development policies. Historically the primary management focus on chaparral has been one of fuels and fire hazard (Parker 1987, 1990; see Chapter 3, “Fire as an Ecosystem Process”); indeed that was the primary motivation for the development of the vegetation-type mapping project begun in the late 1920s (Keeley 2004). Today fire hazard and watershed hydrology are the primary foci of management, and the maintenance of chaparral cover for its critical role in hydrology will increase in importance in a future with potentially less precipitation and warmer temperatures.

Summary Chaparral shrublands are biotically diverse and the most abundant vegetation type in the state. These shrublands are dominated by evergreen species and occur in areas with hot, dry summers and cool, moist winters. The species that inhabit the chaparral are adapted to a Mediterranean-type climate. Chaparral is a dynamic ecosystem, and wildfires (see Chapter 3, “Fire as an Ecosystem Process”) facilitated by the summer rainless period are more predictable than in many other fire-prone landscapes; this is reflected in many evolutionary responses to fire. While diverse in microbes and animals, the dominant plants exhibit characteristics explicitly selected by wildfire in their development of persistent soil or canopy seed banks. Chaparral ecosystems provide many services, such as the stabilization of steep slopes, filtration of drinking water, and myriad recreational opportunities. They also beautify the landscape for many millions of California inhabitants and visitors. Increasingly, chaparral is being managed for its intrinsic value to resource conservation and even community restoration programs. There are many threats to chaparral ecosystems from climate change, altered fire regimes, development, nonindigenous invasive species, and poor management practices.

Acknowledgments V. Thomas Parker was supported by the San Francisco State University Office of Research and Sponsored Programs; Chaparr al  499

R. Brandon Pratt by a National Science Foundation CAREER grant (IOS-0845125); and Jon E. Keeley by the U.S. Geological Survey Fire Risk Scenario Project.

Folivore  An herbivore that specializes in eating leaves.

Recommended Reading

Lysimeter  A device used to measure the evapotranspiration from plant/soil systems as the difference between precipitation inputs and water lost through the soil.

Keeley, J. E. 1989. A chaparral family shrub—A genealogy of chaparral ecologists. Pages 3-6 in S. Keeley, editor. California Chaparral: Paradigms Re-examined. Natural History Museum of Los Angeles, Science Series No. 34, Los Angeles, California. Keeley, J. E. 1993. Proceedings of the symposium interface between ecology and land development in California. International Journal of Wildland Fire, Fairfield, Washington. Pincetl, S. S. 2003. Transforming California: A political history of land use and development. Johns Hopkins University Press, Baltimore, Maryland. Stephenson, J. R., and G. M. Calcarone. 1999. Southern California mountains and foothills assessment: Habitat and species conservation issues. General Technical Report, PSW-GTR-172. USDA Forest Service, Pacific Southwest Research Station, Albany, California. Tenhunen, J. D., F. M. Catarino, O. L. Lange, and W. C. Oechel. 1987. Plant response to stress. Functional analysis in Mediterranean ecosystems. Springer, New York, New York.

Glossary Arbuscular mycorrhizae  Mycorrhizae are mutualistic associations between higher fungi and vascular plants. In arbuscular mycorrhizae the fungus forms hyphae that penetrate the cell walls of plant roots, where they form structures (arbuscules) with large surface areas that facilitate the exchange of minerals and carbohydrates. The fungi involved in arbuscular mycorrhizae are from a lineage in the Zygomycetes. Cations  A positively charged atom or molecule such as potassium and calcium. Cavitation  The breaking of the water column in vessels or tracheids whereupon liquid water changes to water vapor ultimately leading to emboli. Plants differ in their susceptibility to cavitation, but it is common in the vascular tissues when water content of the soil is low or during episodes of freezing and thawing. Detritivores  These heterotrophic organisms consume detritus (which is decomposing plant and animal parts) to obtain energy and minerals, and consequently they contribute to decomposition and the nutrient cycles. Ectomycorrhizae  These mycorrhizae (mutualistic associations between higher fungi and vascular plants) are characterized by the presence of a fungal mantle covering the plant host roots with some fungal hyphae penetrating between root cells (called a Hartig net). The fungi are usually from two fungal groups, Basidiomycetes and Ascomycetes. Emboli (sing. embolus)  These gas bubbles form following cavitation in the water and transport cells (vessels or tracheids) within plant vascular tissue. Eutrophication  The response of an aquatic system to pollutants contained in runoff. In the case of nitrogen pollutants, this response is often a bloom of algal growth that can choke waterways and lead to the reduction in oxygen levels to levels lethal to some species. Evapotranspiration  The sum of the water lost via evaporation and transpiration. Facultative seeders  Shrubs or trees that survive wildfire and sprout new shoots after the fire. These types of plants also have persistent soil seed banks that fire stimulates, and 50 0  Ecosystems

they recruit new individuals from seedlings that successfully establish in the postfire environment.

Mycorrhiza (pl.: mycorrhizae)  A symbiotic, mutualistic (but occasionally weakly pathogenic) association between a fungus and the roots of a vascular plant. Mycorrhizae are critical in bringing water and minerals to their host plant, which in turn provides carbon energy to the fungus. Mycota  Refers to species from the kingdom Fungi. Nitrification  The oxidation of ammonia into nitrate by microorganisms. Obligate resprouter  Shrubs or trees that can survive wildfire and sprout new shoots after the fire. Generally no seedlings are recruited in the postfire environment. Obligate seeders  Shrubs or trees that are killed by wildfire and persist in the habitat because they have persistent soil seed banks that fire stimulates. Obligate seeders recruit new individuals after wildfire from the seed bank and establish new populations in the postfire environment. Persistent seed bank  Refers to viable seed constantly being found in the soil unless stimulated by a strong environmental event such as a wildfire. Physical dormancy  When referring to seeds, this generally means that there is a thick seed coat or other structure that does not allow water or gasses to enter a seed, keeping the seed in a dormant state. Physiological dormancy  When referring to seeds, this generally means that there are physiological processes that have to be met before normal metabolism will stimulate germination; examples are light, temperatures, and chemicals from smoke. Pyro-endemics  Refers to annual or short-lived plants that are only found in postfire stands of vegetation. Sclerophyllous  Refers to plants or to a vegetation indicating plants have generally small tough leaves with thick cuticles; usually an adaptation to low water availability and nutrientpoor soils. Seed bank  Viable plant seed stored in the soil or in serotinous cones or woody fruit in the canopy of a tree or shrub. Serotinous  Having seed held in woody cones or fruit rather than releasing at seed maturation. Release occurs in response to an environmental trigger, usually fire, or death of the stem. Specific leaf area (SLA)  The fresh leaf area divided by the oven-dry mass; SLA is an index of sclerophylly with lower values being more sclerophyllous. Stratification  Some seeds require a chilling period (generally several degrees above freezing) that lasts a minimum time period (weeks to months) before dormancy is broken and germination can occur; the chilling is called stratification. Suffrutescent  A small shrub having a stem that is woody only at the base or some of the main stems; generally the wood is light. Transient seed bank  Refers to all seed germinating or losing viability within a year such that there is a fraction of the year with no seed stored in the soil. Transpiration  Loss of water through the tiny pores (stomata) of plant leaves. Water potential  The potential energy of water relative to pure water. It is used as a measure of plant water status

with more negative values indicating that tissues are more dehydrated. Xylem  The vascular system of plants that conducts water. Cells are usually dead and hollow at maturity.

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T WENT Y-FIVE

Oak Woodlands

ECOS

FR AN K W. DAVIS, DENN IS D. BALDOCCH I , and CL AUD IA M . T YLER

Introduction For thousands of years California’s oak woodland and savanna ecosystems have provided a bounty of ecological goods and services to human societies. Oak woodlands harbor high plant and animal diversity and supply livestock forage and firewood. They occur in climates and on soils that can produce many valuable crops, including some of the world’s finest wine grapes. They are also among the state’s most iconic and attractive landscapes and are prized for rural residential housing (Pavlik et al. 1993). In the 250 years since European settlement, as human values and socioeconomic systems changed and population growth increased pressure on these systems, oak woodlands have undergone significant changes in human use and associated landscape and ecosystem structure, composition, and function. Some keystone species, like the California grizzly bear, have been eliminated. Domestic livestock, exotic diseases, and invasive plants and animals have transformed natural communities. Fire, once a dominant force, is now suppressed. Climate, watershed hydrology, and nutrient inputs have changed rapidly over the past century. Today some large tracts of oak woodland still retain their rural character and ranching economy, but many others are fragmented and degraded by

residential development, agricultural conversion, and dense road networks. In this chapter we summarize current knowledge about the distribution, structure, and ecosystem ecology of California oak woodlands and savannas. Another recent review covers the distribution, composition, and general ecology and management of oak woodlands (Allen-Diaz et al. 2007). We focus on biodiversity and ecosystem processes in foothill oak woodlands and particularly blue oak (Quercus douglasii) woodlands, where ecosystem processes have been studied in more detail. We discuss historical trends and current patterns of human use of foothill oak woodlands and some ecosystem services they provide. We conclude by considering the future of oak woodlands under projected climate and land use change.

Distribution By modern convention, “woodland” is defined as vegetation with 20% to 60% tree cover (Sawyer et al. 1995, 2009; Anderson et al. 1998), but California inventories and wildlife habitat assessments have tended to also include vegetation with 509

10% to 20% tree cover, pooling oak “savannas” and “woodlands” into a single type (Mayer and Laundenslayer 1988, Gaman and Firman 2006). We will broadly define oak woodlands as vegetation with 10% to 60% tree canopy cover, dominated by one to several deciduous or evergreen arborescent oaks, and with an herbaceous understory often dominated by exotic annual grasses, native and exotic forbs, and scattered shrubs. For some oak woodland types, the understory can alternatively be comprised of a dense layer of shrubs associated with chaparral or coastal sage scrub. Oak woodlands extend across coastal, foothill, and montane environments, but foothill woodlands are most extensive (Figure 25.1). Southern coastal oak woodlands are dominated by coast live oak (Q. agrifolia) or the more geographically restricted Engelmann oak (Q. engelmannii). Interior foothill woodlands that occupy the periphery of the Central Valley and Interior Coast Ranges are dominated by blue oak, interior live oak (Q.wislizenii var. wislizenii), and valley oak (Q. lobata). Foothill pine (Pinus sabiniana) often co-dominates the tree layer with blue oak in interior foothill woodlands. Oregon white oak (Q. garryana) woodlands increase in importance in the northwestern part of state, whereas canyon live oak (Q. chrysolepis) and black oak (Q. kelloggii) dominate montane oak woodland and forest vegetation across large areas of the Sierra Nevada and Coast Ranges. According to a recent statewide inventory, oak woodlands occupy roughly 3.5 million hectares (8.6 million acres) or 8–​9% of the California land area. Foothill oak woodlands account for slightly over 2 million hectares (5.1 million acres), or 60% of the woodland acreage. Blue oak woodland is by far the most widespread type, comprising 37% of all oak woodlands and 62% of foothill oak woodland area (Table 25.1).

depending on local rainfall inputs, soil water storage capacity, and losses to actual evapotranspiration (AET), surface runoff, and groundwater. This means that for deciduous oaks like blue oak, valley oak, and Oregon oak, soil water supply is greatest several months before the tree canopy has fully developed. While many understory plants, notably now-dominant non-native annual grasses, have winter-spring growing seasons to exploit the early pulse of shallow soil moisture, foothill oaks must be adapted to withstand hot and dry summer-growing seasons. The asynchrony between water supply and demand in oak woodlands is illustrated by trajectories of cumulative climate water deficit (CWD) between December and November (Figure 25.3). This deficit is the difference between potential evapotranspiration (PET), which is the amount of water that would have been lost had it been available, and actual evapotranspiration (AET) as limited by available soil water (Stephenson 1998). Given a soil’s water-holding capacity and wilting point, soil water available for evapotranspiration (AW) can be estimated as (Flint and Flint 2007):

Oak Woodland Climate

When CWD is positive, evaporative demand exceeds available soil water. In blue oak and coast live oak woodlands, modeled cumulative CWD is close to zero between November and April but increases rapidly thereafter (see Figure 25.3). Averaged across the ranges of the species, cumulative CWD for coast live oak and blue oak totals over 800 mm by the end of the year. Cumulative CWD is less than 500 mm for black oak, which occupies wetter and cooler environments. The CWD trajectories suggest that readily available soil water, at least as modeled using mapped soil water-holding capacity and climate grids, is depleted by late spring, during or shortly after the production of new foliage. This implies that oaks must also be tapping deeper groundwater or water in rock fractures to remain active throughout the hot summer months, a conclusion supported by tracer studies (Lewis and Burgy 1964), water balance models, and observed water table fluctuations (Miller et al. 2010, Gou and Miller 2013). In summary, the various oak woodland ecosystems in California occur across a fairly wide range of temperature and rainfall regimes. Additional variation is created by local topography and soil properties that exert a strong control on radiation and temperature, evapotranspiration, and available soil water supply. The resulting local environments exert strict and multiple constraints on oak woodland function, structure, and metabolism, as is discussed below.

Oak woodlands experience a Mediterranean-type climate with cool wet winters and hot dry summers. Foothill woodland types develop in locations experiencing 400–​8 00 millimeters (mm) of winter rainfall and mean annual temperatures of 14°C to 17°C (Figure 25.2). Montane woodland and forest types typically receive 1,200–​1,400 mm of precipitation as rain and snow and experience mean annual temperatures of 10°C to 11.5°C. Chaparral vegetation is widespread in the lower montane settings between these foothill and montane zones (see Chapter 24, “Chaparral”). Foothill woodland climates combine extreme precipitation and temperature seasonality. Over 90% of precipitation falls between December and March when days are cool and nighttime frosts are common. During the period 1911–​2000, winter daily minimum air temperatures at 3,273 sample plot locations supporting blue oak averaged 1.4 ± 1.4°C with over one hundred frost days per year. Summer maximum daily temperatures at these same sites averaged 32.9 ± 3.4°C. High yearto-year variation in precipitation occurs in relation to threeto seven-year El Niño and La Niña climate oscillations and other atmospheric dynamics (Haston and Michaelsen 1994, Jones 2000). Soil moisture in oak woodlands is highest between January and March and lowest in late summer or early autumn, Photo on previous page: Blue oaks at the University of California Sedgwick Reserve near Los Olivos covered with the epiphytic lichen, Ramalina menziesii. Photo: C. Tyler. 510  Ecosystems

AW = P + Sm –​PET –​Sa + Ss

(Eq 25.1)

Where,

P = precipitation Sm = snowmelt PET = potential evapotranspiration, estimated based on modeled hourly solar radiation and air temperature (the Priestley-Taylor equation) Sa = snow accumulation and snow storage carried over from previous time period Ss = soil water storage carried over from the previous time period

Biodiversity and Characteristic Species The rich flora and fauna of California’s oak woodlands contribute considerably to the designation of the California

FIGURE 25.1 Generalized distributions of foothill oak woodlands in California. Montane oak woodlands are not mapped. Data from U.S. Geological Survey, Gap Analysis Program (GAP). Map: Parker Welch, Center for Integrated Spatial Research (CISR).

TA B L E 2 5 .1 Areal extent of oak woodlands in California, circa 2003, based on 30-meter satellite imagery and ancillary sources.

Woodland type

Acres

Hectares

Phenology

General distribution

Semideciduous

Narrowly distributed endemic, Peninsular Ranges, mainly San Diego and Riverside Counties

FOOTHILL WOODLANDS Engelman oak* Quercus engelmannii

20,367

8,246

930,534

376,734

Evergreen

Coastal plain, valleys, and foothills of the Central Coast Ranges, Transverse Ranges and Peninsular Ranges

Valley oak* Q. lobata

85,882

34,770

Deciduous

Interior valleys, foothills, and riparian zones, Santa Monica Mountains north to Oregon

Blue oak* Q. douglasii

3,184,018

1,289,076

Deciduous

Interior foothills of the Central Coast Ranges and foothills surrounding the Central Valley

869,380

351,976

Evergreen

Foothill to lower montane zones surrounding the Central Valley and in southern ranges, from the upper Sacramento Valley to northern Baja California

Coast live oak Q. agrifolia

Interior live oak* Q. wislizeni var. wislizeni

MONTANE AND NORTHERN WOODLANDS Black oak Q. kelloggii

692,507

280,367

Deciduous

Lower to mid-montane zone, widespread throughout mountain ranges west of and including the Sierra Nevada

Oregon oak Q. garryana

639,449

258,886

Deciduous

Valleys and foothills, western Sierra Nevada, Cascades, and North Coast Ranges north to British Columbia

1,016,373

411,487

Evergreen

Lower and mid-montane zones throughout California

Tanoak Notholithocarpus densiflorus

388,695

157,366

Evergreen

Coastal and lower montane, mainly Coast Ranges with scattered population in Cascades and northern Sierra Nevada

Mixed oak

738,455

298,970

Mixed

Foothill and montane zones, especially upper Sacramento Valley, North Coast Ranges, and Central Coast Ranges

8,565,660

3,467,879

Canyon live oak* Q. chrysolepis

Total

Source: Gaman and Firman 2006. * Indicates species whose native ranges are restricted or nearly so to California.

­ loristic Province as a “biodiversity hotspot” (Myers et al. F 2000). Only a handful of tree species define and shape these communities, but this relatively simple structure belies the diversity within. Oak woodlands support thousands of understory plant species, over three hundred species of vertebrates, and thousands of invertebrate species. This diversity creates complex networks of species interactions and interdependence that have yet to be extensively studied. Here we focus on some of the principal organisms and interactions that play key roles in these communities.

Plants Blue oaks can be the only overstory species over large areas of blue oak woodland. Other tree species present in these woodlands can include foothill pine, California buckeye 512  Ecosystems

(Aesculus californica), valley oak, interior live oak, and coast live oak (Pavlik et al. 1993, Sawyer et al. 2009). Co-dominant tree species in valley oak woodlands include trees common to riparian habitats such as box elder (Acer negundo), white alder (Alnus rhombifolia), Oregon ash (Fraxinus latifolia), western sycamore (Platanus racemosa), Fremont cottonwood (Populus fremontii), and California walnut (Juglans californica). In drier upland habitats, valley oaks can co-occur with coast live oak, interior live oak, or blue oak (Pavlik et al. 1993, Sawyer et al. 2009). Coast live oaks often occur in monospecific woodland and forest stands. Other tree species present can include bigleaf maple (Acer macrophyllum), box elder, madrone (Arbutus menziesii), California walnut, western sycamore, Fremont cottonwood, blue oak, valley oak, and bay laurel (Umbellularia californica) (Pavlik et al. 1993, Sawyer et al. 2009). The herbaceous understory is the locus of oak woodland plant diver-

1600 1400

P

K

-

-

Q. agrifolia (1430) Q. chrysolepis (8163) Q. douglasii (3273) Q. engelmannii (161) Q. garryana (1233) Q. kelloggii (8289) Q. lobata (838) Pinus ponderosa (9892) Q. wislizenii (3689)

1000

1200

CG

W

800

MAP (mm)

A C D E G K L P W

600

D

L A

400

E

10

11

12

13

14

15

16

17

MAT (°C) FIGURE 25.2 Modeled median annual temperature (MAT) and median annual precipitation (MAP), 1971–​2 000, at locations supporting eight dominant species of California oaks. Ponderosa pine is included for comparison. The 270 meter climate grids are produced as described in Flint and Flint 2012 and downloaded from . Climate values extracted for species localities in the database as described by Viers et al. 2006. Numbers in parentheses indicate the number of sample localities for each species. Lines intersect at the median MAT and MAP values for species identified by a letter code (placed to the upper left of median values). They span the first to third quartiles of the observed distribution. Foothill woodland species cluster in the lower right portion of the figure, and montane woodland species cluster in the upper left. Illustration created by Frank W. Davis.

sity. Although relative cover is dominated by a few non-native annual grasses (e.g., Avena barbata, A. fatua, Bromus diandrus, B. hordeaceous, B. madritensis, or Hordeum murinum), the herb layer can support high species richness. Many species are native, including annual grasses (e.g., Bromus carinatus, Vulpia microstachys), perennial grasses (e.g., Elymus glaucus, Stipa sp., Poa secunda, Koeleria macrantha), and numerous perennial and annual forbs. The annual forbs are particularly diverse, accounting for the preponderance of rare and threatened plant species associated with oak woodlands. The herb layer also includes many non-native forb species, some of which are invasive and undesirable (e.g., yellow starthistle [Centaurea solstitialis], tocalote [Centaurea melitensis]). Others nonnatives are ubiquitous but less noxious (e.g., Cerastium glomeratum, Erodium cicutarium, Stellaria media). The species composition of the herbaceous understory vegetation is influenced by a variety of factors that have been investigated either in oak habitats or adjacent grasslands. These factors include climate, tree cover (Parker and Muller 1982, Borchert et al. 1991, Maranon and Bartolome 1993, Rice and Nagy 2000, Roche et al. 2012), insolation (Borchert et al. 1991), present and historical land use (Stromberg and Griffin 1996, Knapp et al. 2002, Keeley et al. 2003), and annual variation in rainfall (Knapp et al. 2002, Suttle et al. 2007),

fire, and disease (Brown 2007). Oak woodlands support many moss and lichen species. A distinctive constituent of relatively cool and humid oak woodlands, especially stands of blue oak near the coast, is the epiphytic lichen, Ramalina menziesii (see Lead Photo). This lichen plays an important role in biomass turnover and nutrient cycling in these ecosystems, as it captures rainfall and dry deposition of atmospheric nitrogen and phosphorus (Boucher and Nash 1990, Knops et al. 1996). Experiments conducted at the University of California’s Hastings Reserve demonstrated that trees with lichens present had higher nutrient levels (e.g., N, Ca) in the throughfall beneath the canopy as well as slower decomposition rates of oak leaf litter (Knops et al. 1996)—​factors that could affect abundance or composition of understory species. Shrubs and woody sub-shrubs can be present in oak woodland or savanna but are generally at low density and cover. Frequent species include poison oak (Toxicodendron diversilobum), California coffeeberry (Rhamnus californica), coyote bush (Baccharis pilularis), and various species of Ceanothus, Arctostaphylos, Ribes, and Eriogonum. The shrub species represented in a particular oak habitat are governed by soil moisture and texture, amount of canopy cover, and climate. However, the cover or relative abundance of woody understory species is often a function of disturbance (Tyler et al. 2007). Oak Woodlands   513

0

Cumulative CWD (mm) 400 800

Coast live oak (MAP = 530 mm, MAT = 16.9 ºC) Blue oak (MAP = 657 mm, MAT =15.2 ºC) Black oak (MAP =1229 mm, MAT =11.2 ºC)

N

D

J

F

M

A

M

J

J

A

S

O

Month FIGURE 25.3 Modeled trajectories of cumulative climate water deficit (CWD) for the period 1971–​2 000 for the ranges of two foothill oak species (coast live oak, blue oak) and a montane oak species (black oak). Values are averages of modeled CWD (270 meter resolution) at the same vegetation plot locations used to produce Figure 25.2. Grids of CWD produced by L. and A. Flint using methods detailed in Flint and Flint 2007, Flint and Flint 2012. These data downloaded January 2013 from . Also shown are estimated median annual precipitation (MAP) and median annual temperatures (MAT) across all localities for each species. Illustration created by Frank W. Davis.

Common Vertebrates and Invertebrates The native fauna within California’s oak woodland habitats is remarkably diverse, with over three hundred species of vertebrates (Guisti et al. 1996, CDFW 2008) and thousands of invertebrates (Swiecki et al. 1997). Some are species of conservation concern with special status designation by the California Department of Fish and Game and/or the U.S. Fish and Wildlife Service. The overwhelming majority of animal species present depends directly or indirectly on the food or shelter that oaks provide (CDFW 2008). The diversity of fauna reflects the abundance and diversity of those resources: acorns, leaves, twigs, sap, leaf litter, and a complex physical structure including canopy, shaded and open branches, cavities, bark, and standing, dead, or downed logs. Acorns are a nutrient-rich food resource for many species. The acorn crop varies greatly from year to year, especially white and red oak species like blue oak, valley oak, and coast live oak, whose acorns mature in one year (Koenig et al. 1994, Koenig and Knops 2013). As in other temperate and Mediterranean oak woodlands and forests, oak masting can drive population dynamics of animals dependent on them and throughout the food webs (Koenig and Knops 2005, Ostfeld et al. 2006, Koenig et al. 2009).

VERTEBR ATE S

Mammal species using oak woodland habitats include carnivores such as cougar (Puma concolor), bobcat (Lynx rufus), American badger (Taxidea taxus), coyote (Canis latrans), gray fox (Urocyon cinereoargenteus), and long-tailed weasel (Mustela frenata), as well as large omnivores and browsers such as American black bear (Ursus americanus) and mule deer (Odocoileus hemionus). Insect, seed, or herb consumers include tree and ground squirrels, Botta’s pocket gopher (Thomomys 514  Ecosystems

bottae) and other rodents, brush rabbit (Sylvilagus bachmani), and bats. Grizzly bears (Ursus arctos californicus) were historically abundant and may have been important not only as predators but also through their extensive digging for food resources in the ground layer (Pavlik et al. 1993). Now nonnative wild pigs (Sus scrofa) are an important agent of soil disturbance in many oak woodlands and grasslands, but their activities are associated with increases in exotic plant species that recover from pig disturbance faster than native species (Tierney and Cushman 2006). A few of the mammal species commonly found today significantly reduce the rate of acorn, seedling, and sapling survival, especially gophers and California ground squirrel (Otospermophilus beecheyi), which limit seedling emergence and survival, and deer, which can limit sapling establishment (Tyler et al. 2006, Davis et al. 2011). Relative abundances of rodent species are correlated with microhabitat such as shrub and downed wood cover. Duskyfooted woodrats (Neotoma fuscipes), for example, are more abundant where there is shrub cover, whereas a sparsely vegetated ground cover is associated with higher numbers of deer mice (Peromyscus spp.) (Tietje and Vreeland 1997). Birds are the most diverse group of vertebrates in these habitats, with over one hundred species represented. Some of these rely on acorns as a primary food source, notably the acorn woodpecker (Melanerpes formicivorus), western scrub-jay (Aphelocoma californica), yellow-billed magpie (Pica nuttalli), oak titmouse (Baeolophus inornatus), Nuttall’s woodpecker (Picoides nuttallii), and band-tailed pigeon (Patagioenas fasciata). Others, such as phainopepla (Phainopepla nitens) and western bluebird (Sialia Mexicana), rely on food sources associated with oaks, feeding seasonally on the fruits of the epiphytic oak parasite, mistletoe (Phoradendron spp.). Many birds that breed in oak habitats are cavity-nesting species and thus benefit from the matrix of living and dead branches in old large oak trees, as well as downed or standing dead trees (Tietje et al. 1997). Raptors such as red-tailed hawk, red-shouldered hawk, Cooper’s hawk, great horned owl, and golden eagle, though not restricted to oak habitats, nest in these woodlands and can play a major role in the community trophic structure as predators on small mammal populations (Tietje et al. 1997). Turkey vultures (Cathartes aura) are an important carrion feeder in these ecosystems. Reptiles and amphibians benefit from shelter provided by leaf litter, downed wood, or hollows in oak trunks, as well as abundant food sources—​especially insects and small mammals (Tietje and Vreeland 1997, Tietje et al. 1997, Block and Morrison 1998). Lizards and snakes are important predators in oak communities and include skinks (Plestiodon spp.), western fence lizard (Sceloporus occidentalis), California legless lizard (Anniella pulchra), alligator lizards (Elgeria spp.), racers (Coluber spp.), common kingsnake (Lampropeltis getula), gopher snake (Pituophis catenifer), and western rattlesnake (Crotalus oreganus). Gopher snakes and western rattlesnakes, though not always apparent, may significantly impact small mammal populations. For example, research indicates they can remove 14–​35% of juvenile ground squirrels annually (Fitch 1948, Diller and Johnson 1988). Amphibian abundance and diversity vary considerably with site conditions and associated microhabitats. Salamanders and frogs are among the most common amphibian species, notably slender salamanders (Batrachoseps spp.), ensatina (Ensatina spp.), and chorus frogs (Pseudacris spp.) (Tietje and Vreeland 1997, Block and Morrison 1998). Most species depend on moist microsites under downed logs and brush piles or

in litter, riparian elements, or vernal pools (Block and Morrison 1998).

in production and maintenance of soil and in nutrient flow, including microfauna (nemotodes), mesofauna (mites, springtails, larval insects), and macrofauna (earthworms, beetles, flies, ants, snails).

INVERTEBR ATE S

Thousands of invertebrate species are associated with California’s oak woodland habitats, including an estimated five thousand arthropod species, mainly insects (Swiecki et al. 1997). Some depend on and affect oaks directly. Nearly all parts of an oak provide food for some insect, which can be classified based on its niche: acorn feeder; foliar feeder; insect gall former; sap feeder; or twig, bark, and wood borers (Swiecki and Bernhardt 2006). While some of these insect species are significant pests (e.g., the introduced goldspotted oak borer), most native species seldom inflict severe or fatal damage to the host tree, even those that have conspicuous impacts (Swiecki and Bernhardt 2006). One species with dramatic though rarely fatal impact on oaks is the California oak worm (Phryganidia californica), which is the larval stage of the California oak moth. Outbreaks of the oak moth fluctuate annually, with up to three generations per year. At high population densities the larvae can defoliate one to many trees within a stand. Following an outbreak, the oak moth population may collapse due to predation and larval parasites (Swiecki and Bernhardt 2006). Both environmental and density-dependent factors contribute to population limitation (Milstead et al. 1987). Interactions between oaks and oak moth larvae strongly affect cycling of macronutrients such as nitrogen and phosphorus because input via leaf litter is reduced, while input from insect fecal material is increased, during outbreaks (Hollinger 1986). Filbert weevils (moth larva, Curculio sp.) and filbertworm (beetle larva, Cydia latiferreana) are the most common invertebrate consumers of acorns. Levels of infestation vary annually and among and within trees (Lewis 1992). Reported infestation levels range from 38% to 62% on coast live oak acorns (Lewis 1992, Dunning et al. 2002) but can be as high as 75% to 80% (Swiecki and Bernhardt 2006). In spite of these high rates, it is not clear that infestation results in loss of germinability in the acorns affected. Dunning et al. (2002) found that most infested acorns suffered less than 20% damage, which might not prevent germination if tissue is intact at the growing tip (Swiecki and Bernhardt 2006). However, in valley oak, over an eight-year period Griffin (1979) estimated that 20% of viable acorns dropped were lost to insect damage. Among the noticeable, but relatively harmless, impacts of insects on their oak hosts are insect galls, induced most often by wasps in the family Cynipidae. California oaks support more than two hundred species of gall wasps, which each specialize on either the white oak or black oak group (Russo 2006). The large “oak apple,” produced by the wasp Andricus quercuscalifornicus, can reach a diameter of 8 centimeters (cm) and is one of the most commonly recognized galls, though a great diversity of shapes, colors, and sizes of these intriguing structures exists, reflecting the diversity of gall-making species. Perhaps even more fascinating are the complex associated food webs, as the galls themselves and developing larvae attract a myriad of predators, parasites, competitors, and mutualists (Russo 2006). Finally, not as well understood or described but of great importance to the oak woodland ecosystem is the invertebrate community residing in the litter layer and soil. In addition to serving as a prey base for small vertebrate consumers, many organisms here play vital roles

Oak Diseases A diverse and still poorly understood array of disease organisms variously afflict oak acorns, leaves and twigs, branches and trunks, or root systems (Swiecki and Bernhardt 2006). Most of these are of minor importance, but a few have serious ecological and economic impacts. Canker rot fungi, notably Inonotus andersonii and I. dryophilus, cause widespread mortality among oak. Sulfur fungus (Laetiporus gilbertsonii) is also a major cause of tree failure (Swiecki and Bernhardt 2006). Since its discovery in the San Francisco Bay Area of California in the mid-1990s, Phytophthora ramorum, the causal agent of the forest disease known as sudden oak death (SOD), has killed millions of tanoaks (Notholithocarpus densiflorus) and oaks in the mixed broadleaf evergreen forests and redwood forests of coastal California (Meentemeyer et al. 2008). P. ramorum is now established in coastal forests from the Big Sur coast northward to southern Mendocino County; disjunct introductions have also been detected in southern Humboldt County and Curry County, Oregon. This introduced disease infects at least thirty native California woody and herbaceous host species, causing nonlethal foliar and twig damage in some species (e.g., the important host and carrier California bay laurel, Umbellularia californica) and lethal twig and stem cankers in others. Not all oaks are affected. Red and black oaks, which are placed into the Erythrobalanus (“red oak”) section of the genus and include coast live oak, Shreve’s oak (Quercus parvula var. shrevei), and black oak, are especially vulnerable (Rizzo and Garbelotto 2003). The disease does not affect white oaks (section Quercus), like valley oak and blue oak, and is mainly associated with forest climates rather than drier foothill woodlands. However, where the disease occurs, it has significantly impacted coast live oak woodlands and forests (Davis et al. 2010). Annual mortality rates for coast live oak of 4.5% to 5.5% (an order of magnitude higher than background rates) have been reported in infested areas of Big Sur and the San Francisco Bay region (Brown and Allen-Diaz 2009).

Functional Roles: Foundation Species, Keystones, and Ecosystem Engineers To truly portray the role a species plays in its community requires an understanding of the complex web of direct and indirect interactions within that system. For California oak woodlands this level of understanding has not yet been achieved. Many of the “pieces” required to build models of these complex interactions are available: lists of species (as sketched earlier), and both population- and communitylevel studies of some key organisms. However, most of the latter focus on interactions involving pairs of or several species. Some notable exceptions in North American oak woodland research include the studies conducted in eastern oak forests illuminating the complex web linking acorn production to gypsy moth outbreaks to Lyme disease risk (Ostfeld et al. 1996, Jones et al. 1998). In spite of our limited knowledge of interaction webs, certain species do appear to play particularly important roles in Oak Woodlands   515

the function of California’s oak woodlands. Next we propose terms that describe some of these key roles and suggest that this is an area of oak woodland ecology with great potential for future research. A better understanding of the functional roles played by various species in these systems may improve our ability to predict consequences of species losses and potential extinction cascades (Zavaleta et al. 2009, Colwell et al. 2012).

OAKS AS FOUNDATION SPECIE S

Obviously, without oaks there is no oak woodland. They define the structure of the community both literally and functionally, and the vast majority of animal species in these systems depend directly or indirectly on the resources that oaks provide. Although it has not been calculated or even estimated, the “community importance” value (sensu Power et al. 1996) of oaks is clearly greater than any other species in this ecosystem. We propose that the term “foundation species” best describes the role of oaks in California. As discussed by Ellison et al. (2005), a foundation species is one that “controls population and community dynamics and modulates ecosystem processes,” whose loss “acutely and chronically impacts fluxes of energy and nutrients, hydrology, food webs, and biodiversity” (Ellison et al. 2005, p. 479). Oaks act as foundation species whether they are “dominant” in terms of biomass or abundance (oak forests and dense woodlands) or less common (oak savanna).

siderable impacts, such as reduced species diversity (Mills et al. 1993). In oak woodlands, large carnivores such as cougars (Puma concolor) may function as keystone predators by reducing herbivore populations that limit oak establishment. For example, mule deer (Odocoileus hemionus) have been found to limit establishment of oak saplings (Ripple and Beschta 2008, Tyler et al. 2008, Davis et al. 2011), but their impacts are reduced where cougars are active (Ripple and Beschta 2008). A review of large predators and trophic cascades in five national parks in the western United States supports the hypothesis of top-down control in oak woodlands (Beschta and Ripple 2009). The authors report that where large predators have been displaced or locally extirpated, ungulates have had major impacts on dominant woody species (e.g., black oaks in Yosemite) and ecological processes. Similarly, medium-sized carnivores such as the bobcat (Lynx rufus), American badger (Taxidea taxus), and coyote (Canis latrans) may function as keystone predators by reducing populations of pocket gophers and ground squirrels—​ small mammal species that significantly limit oak establishment at the seedling stages (Tyler et al. 2006, Tyler et al. 2008). The importance of these carnivores in structuring oak woodland communities has not been studied, and impacts of their removal on oak populations may be difficult to discern given the long life span of oaks.

ECOSYSTEM ENGINEERS

Several groups of oak woodland species function as ecosystem engineers as defined by Jones et al. (1994). These organMUTUALIST SEED -DISPERSERS

Oaks do not require seed dispersal by animals, as evidenced by the cohorts of young seedling that establish in some years under the tree canopy or at the drip-line. However, most acorns that drop under adult trees are consumed by animals, and those that remain on the soil surface succumb to desiccation or heat stress. Acorns that are shallowly buried have a higher probability of producing emergent seedlings (Borchert et al. 1989, Nives and Weitkamp 1991), and while some may be buried naturally (e.g., by falling into deep leaf litter), acorn-caching animals likely play an important role in dispersing acorns to microsites that promote their survival and germination. Of singular importance is the western scrubjay (Aphelocoma californica), described by Grinnell (1936) as the “uphill planter” responsible for the distribution of black oak seedlings upslope from adult trees. The western scrub-jay may cache, or “plant,” up to five thousand acorns in a season but only recover and consume half (Carmen 1988), leaving thousands of buried acorns across the landscape. Other acorn-caching animals include the yellow-billed magpie (Pica nuttalli) and the California ground squirrel.

isms impact the availability of resources to other species by causing physical changes to the environment and thereby influence biodiversity within the community. “Soil engineers” include pocket gophers and California ground squirrels as well as non-native pigs and badgers. These species influence the composition of soil biota, herbaceous vegetation, and potentially the establishment of oaks by moving or removing acorns. The work of ecosystem engineers in the oaks is also evident, as they modify the tissue of the trees. Some of the most apparent ecosystem engineers include woodpeckers (e.g., acorn woodpecker, Nuttall’s woodpecker) and gall-forming insects (e.g., cynipid wasps). With the exception of feral pigs, the aforementioned “engineers” most likely enhance biodiversity in these systems.

Ecosystem Structure and Processes The ecosystem ecology of Mediterranean-climate oak woodlands is distinctly different from that of temperate oak woodlands (Baldocchi and Xu 2005, 2007). In this section we describe the structure and function of the woodlands and their metabolism in terms of energy capture, water use, carbon assimilation, and respiration.

KEYSTONE PREDATORS

The concept of a keystone species has been broadly applied and debated in the ecological literature (Paine 1969, Mills et al. 1993, Power et al. 1996, Cottee-Jones and Whittaker 2012). Here we use the term to refer to a species high in the food web that has large effects on the community that far exceed its abundance (Power et al. 1996). Accordingly, the removal of such a species from the community is expected to have con516  Ecosystems

Stand Structure Tree crowns in woodlands are typically isolated and nonoverlapping, but the vegetation structure of oak woodlands varies widely depending on environmental setting (climate, topography, and soils), tree composition, and history of tree cutting and burning. The tree canopy usually consists of a single layer

0

250

500 m

FIGURE 25.4 August 2012 orthophoto of the San Joaquin Experimental Range near Fresno, California. Dark dots are individual tree crowns and are mainly blue oaks with scattered foothill pines. The ground layer is predominantly residual thatch from annual grasses that senesced several months earlier. This landscape illustrates the open nature of blue oak woodlands and the tendency for trees to cluster at multiple scales. Source: Google Earth, image date August 26, 2012.

TA B L E 2 5 . 2 Structural characteristics of widespread foothill woodland types Numbers generalized from multiple sources

Typical canopy height [max] (m)

Representative adult tree density (number per hectare)

Coast live oak

6–12 [24]

Valley oak

Woodland type

Total tree basal area (m2/hectare) Average

Range

20–50

9–23

4–59

12–18 [36]

15–100

11–17

6–59

Blue oak

6–18 [28]

100–500

8–12

4–30

Blue oak / Foothill pine

6–18 [35]

100–500

11–14

5–30

Interior live oak

6–18 [28]

10–50

8–14

5–24

Source: Allen-Diaz et al. 2007, Bolsinger 1988, Holzman and Allen-Diaz 1991, Kertis et al. 1993, Whipple et al. 2011.

6–​18 meters (m) in height (Table 25.2); pines, when present, create a second layer 20–​30 m in height. The isolated oak canopies create understory light, temperature, soil moisture, and nutrient conditions that can contrast strongly with those in the surrounding grassland. Part of the challenge in neatly summarizing oak woodland structure is that trees are usually clustered at multiple scales so that properties such as stand composition, tree canopy closure, and leaf area are highly sensitive to the scale of spatial sampling (Figure 25.4). This scale-dependence affects any conclusions drawn from monitoring as well as spatially explicit modeling of ecosystem status and processes based on remotely sensed imagery. Aboveground woody biomass has not been reported for California oak woodlands, so we rely here on published e­ stimates of tree basal area in mature stands (see Table 25.2) and relate those to basal area and aboveground biomass estimates for other temperate and Mediterranean-climate oak woodlands.

Basal area in mature foothill oak woodland ecosystems averages 20–​30 m2 ha-1 but can range from less than 10 m 2 ha-1 to greater than 40 m 2 ha-1 or higher for coast live oak, valley oak, and black oak woodlands. These values are comparable to or slightly lower than those reported for oak woodlands of the Mediterranean Basin, where basal areas average 25–​30 m 2 ha-1 and associated aboveground biomass estimates generally range between 150–​250 t ha-1 (Ibanez et al. 1999, Salvador 2000, Johnson et al. 2002). By comparison, basal area for mature temperate oak forests of eastern North America is typically 25–​30 m2 ha-1 with an associated aboveground biomass of 120–​140 t ha-1 (Held and Winstead 1975, Keddy and Drummond 1996, Johnson et al. 2002). Thus despite growing in a semiarid climate, foothill oak woodlands can support woody biomass comparable to that reported from wetter, second-growth temperate forested systems, although the biomass in California oak woodlands tends to be concentrated in fewer, larger trees. Oak Woodlands   517

0.9 0.8 0.7 0.6

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Canopy Architecture and Phenology California oaks range from broadleaved evergreen and deciduous trees to small-leaved desert and subalpine shrubs, with species’ canopy architecture and phenology adapted to particular light, temperature, and moisture regimes. In foothill oak woodlands, evergreen sclerophyllous, and broad-leaved deciduous trees often co-occur, although where they do, the evergreen live oaks tend to occur in more mesic sites such as north-facing slopes, lower slopes, and swales. By comparison, large-leaved deciduous valley oaks associate with alluvial soils close to the water table, and smaller-leaved deciduous blue oaks increase in importance on upper slopes, spurs, and hilltops (White 1966, Griffin 1973, Hollinger 1992, Robinson et al. 2010). The deciduous blue oaks and valley oaks exhibit similar phenology. Leaf budburst occurs between late February and early April, and leaves develop while trees are producing male and then female flowers (Koenig et al. 2012). Leaves are dropped during October through November, although early leaf abscission can occur under severe drought (Callaway and Nadkarni 1991). The time course of the development of leaf area index (LAI) of deciduous oaks is very dynamic (Ryu et al. 2012), and the start and end of the growing season may vary by thirty days from year to year (Figure 25.5). Leaf area index (LAI) and specific leaf area (SLA) are important ecosystem variables related to energy and nutrient exchange, light absorption, and photosynthesis. Leaf area index, which is defined as the one-sided area of green leaf surface per unit of ground area, is a dimensionless quantity often used to predict primary production and evapotranspiration. Specific leaf area is the ratio of leaf area to leaf dry weight (e.g., mm2 mg-1) and is an important plant functional trait related to photosynthesis, leaf longevity, nutrient retention, and water use. Neither LAI nor SLA has been systematically analyzed across a range of oak woodland sites. However, the potential leaf mass of individual oaks can be estimated with allometric relationships between diameter at breast height (dbh) and tree leaf area (Karlik and McKay 2002). Because leaf mass scales with specific leaf area, one can esti-

0.8

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FIGURE 25.5 Seasonal variation in plant area index (PAI) of deciduous blue oak woodland. The dynamics of canopy structure and PAI were measured using the probability of light transmission, as detected with a network of digital cameras. PAI is almost identical to leaf area index (LAI) but derived from different measurements (Ryu et al. 2012). In two of the four years, windstorms accelerated leaf drop.

518  Ecosystems

Photosynthesis (µmol g-1 s-1)

Plant area index at ecosystem scale

2012 2011 2010 2009

1.0

FIGURE 25.6 Relationship between leaf nitrogen (N) and leaf photosynthesis on a mass basis for the evergreen coast live oak (Q. agrifolia) and the deciduous valley oak (Q. lobata), r2 =0.77. Source: Figure copied from Hollinger 1992.

mate leaf area per tree and the leaf area index of the stand. One set of data indicates that blue oak trees hold about 4 m 2 of leaves per m2 of ground. By inference, taking into consideration the spatial extent among trees, oak woodlands support tree canopy LAI on the order of 0.9 (Karlik and McKay 2002, Ryu et al. 2010).

Oak Physiology Plants face trade-offs and compromises in adopting either an evergreen or deciduous leaf habit. Ecological scaling theory predicts that evergreen life forms are associated with leaves possessing lower nitrogen levels and lower photosynthetic capacity compared to deciduous forms of similar genera (Hollinger 1992, Reich et al. 1997). For example, the deciduous valley oak attains greater rates of photosynthesis by acquiring more nitrogen than the co-occurring evergreen coast live oak (Figure 25.6). Mahall et al. (2009) also documented higher rates of photosynthesis in valley oak than coast live oak. The habit of an oak leaf and its fate of enduring a long, dry summer have a marked impact on the dynamics and capacity of stomatal conductance (the rate of carbon dioxide or water vapor exchange between the leaf and the atmosphere), leaf photosynthesis, and transpiration (Goulden 1996, Xu and Baldocchi 2003). For example, the maximum photosynthetic rate (A max) experiences great seasonal variability in blue oak, peaking at approximately 25 mmol m-2 s-1 just after full leaf expansion during the spring when soils are wet and declining thereafter as stomata close gradually with a depleted soil reservoir (Xu and Baldocchi 2003). It is noteworthy that deciduous blue oaks attain very high A max values compared to evergreen oaks (Goulden 1996), eastern deciduous oaks (Wilson et al. 2000), and fertilized crops (Wullschleger 1993); the evergreen coast live oak, for example, attains peak rates near 10 mmol m-2 s-1 in the spring and drops to below 2 mmol m-2 s-1 in September. The ability for blue oaks to achieve very high rates of photosynthesis rapidly in the spring is probably highly adaptive given their need to acquire enough carbon to sustain the foliar canopy during the short/wet and long/dry periods of the spring and summer, respectively.

Carbon, Water, and Energy Exchange Ecophysiological research on plants occurs across multiple organizational levels and scales from individual leaves to whole plants to multiple plants at the site or ecosystem scale. Knowledge of carbon, water, and energy exchange of California oak woodlands at the ecosystem scale has been produced by three methods. One uses long-term and continuous eddy covariance flux measurements; a second is based on water balance of gauged catchments; and the third uses models driven with remote sensing. Two sets of direct, eddy covariance flux measurements have been made over oak woodlands during the past decade. One set is near Ione, California, in the central foothills of the Sierra Nevada (Baldocchi, Chen, et al. 2011). At this site blue oak woodlands receive about 565 mm yr-1 of rain and evaporate about 390 mm yr-1 of water vapor against a demand of 1,429 mm yr-1 by potential evaporation. On the basis of this water balance, woodlands assimilate 1007 +/- 193 gC m-2 yr-1 of carbon and respire 907 +/- 189 gC m-2 yr-1 (Baldocchi, Chen, et al. 2011). Thus, on average and over many years, this blue oak woodland is a net carbon sink, taking up about 100 gC m-2 yr-1. Another set of carbon and water flux data has been collected along the foothills of the southern Sierra Nevada (Goulden et al. 2012). At this site, at an elevation of 400 m, oak woodlands receive about 500 mm y-1 of rain, evapotranspire about 500 mm yr-1, and assimilate about 1,100 gC m-2 yr-1. Although it is not possible from the eddy covariance measurements to partition water loss from the tree layer versus the understory grasses, evapotranspiration decreases from roughly 3 mm per day in May to 2 mm per day in midsummer to less than 1 mm per day in September, suggesting that the tree layer is responsible for around two-thirds of springtime evapotranspiration. These direct measurements of evapotranspiration from oak woodlands are consistent with water balance studies. In one study the conversion from an oak woodland watershed to grassland found evaporation to decrease from 513 to 378 mm yr-1 (Lewis 1968). Evaporation rates deduced using the Thornthwaite equation (Thornthwaite 1948) and soil moisture accounting range between 280 and 382 mm yr-1 across the oak woodlands of California (Major 1988). Based on a seventeen-year watershed study at the Sierra Foothill field station in the northern portion of the foothill woodland distribution, mean annual evaporation was estimated to be 368 mm +/- 89 mm yr-1 from 708 mm yr-1 of rainfall (Lewis et al. 2000). In a two-year study at Sierra Field station catchment evapotranspiration was found to vary from 296 to 498 mm yr-1 as rainfall ranged between 498 and 610 mm (Swarowsky et al. 2011). A study in Sonoma County of water balance before and after the conversion of oak woodlands (40–​60% canopy cover) to irrigated vineyards concluded that oaks used about 270 mm of water and the grapes used about 300 mm of water per year (Grismer and Asato 2012). The carbon balance of foothill woodlands is the net result of uptake of atmospheric carbon dioxide through photosynthesis and carbon dioxide loss to the atmosphere through ecosystem respiration. These fluxes are difficult to measure at the whole ecosystem level over large areas. Data assimilation modeling using flux measurements, satellite remote sensing information, and statistical analysis indicates that blue oak woodlands collectively take up 53.8 teragrams of carbon per year (TgC yr-1 ), lose 45.2 TgC yr-1 through respiration, and thus accumulate 8.6 TgC yr-1. On an area basis the annual sum

of carbon assimilation is 932 gC m-2 y-1 and the net carbon exchange is -150 gC m-2 y-1 (Baldocchi, Chen, et al. 2010). On a unit land area basis, evergreen oak woodland, with its lower photosynthetic capacity, tends to acquire similar amounts of carbon and use similar amounts of water to a more productive deciduous stand with a similar LAI (Hollinger 1992, Baldocchi, Ma, et al. 2010). In other words, the longer the growing season for the species, the lower the photosynthetic capacity.

Ecosystem Processes in Oak Woodlands: Synthesis Ultimately, oak woodland ecosystems must regulate water use such that carbon gain at least offsets respiratory needs and evaporative demand. How does oak woodland achieve a balance between function, structure, and metabolism given these distinct climatic constraints? We attempt to answer this question for a deciduous oak woodland site by inspecting a set of figures showing links between the fluxes of carbon and water and their dependence on rainfall, soil moisture, and the leaf area index of the canopy (Figure 25.7). First and foremost, the oak woodland must assimilate more carbon through photosynthesis than it loses through respiration (see Figure 25.7a). This is not as trivial as it may seem given the annual asynchrony between the periods of water supply and water demand. The period of highest photosynthetic capacity is relatively short (see Figure 25.7b); peak photosynthesis of deciduous oaks is generally constrained to a hundredday period between late spring (approximately day 80) and early summer (approximately day 180). The soil water reservoir is depleted during the long summer drought period (see Figure 25.3), which induces stomatal closure and down-regulates transpiration and photosynthesis (Goulden 1996, Xu and Baldocchi 2003). Respiration, on the other hand, persists year-round and has the potential to be elevated during hotter periods unless it is also down-regulated by reduced photosynthesis (Atkin et al. 2005). This brings us to the next question: How much carbon assimilation is potentially possible? Sunlight is not a limiting factor in California. The oak woodlands receive about 6.5 gigajoules (GJ) m-2 y-1 of sunlight (Ryu et al. 2008). If this energy was converted to carbon assimilation, we would expect annual carbon assimilation to reach about 1,400 gC m-2 y-1. This computation is based on the assumption that the ­photosynthetic ­e fficiency of an ecosystem is 2%, that half of sunlight is available to photosynthesis, and that the energy content of photosynthetic fixation is 496 kilojoule (kj) mole-1 CO2. The actual amount of woodland photosynthesis is one-half to three-­ quarters of the potential amount based on available sunlight (Ma et al. 2007). The key limiting factor on the annual sum of carbon assimilation by oak woodland is the amount of available water. Consequently, annual photosynthesis, or gross primary productivity (GPP), scales with annual evaporation (see Figure 25.7c). For a deciduous oak woodland, annual photosynthesis (GPP) increases from about 700 to 1,100 gC m-2 y-1 as evaporation increases from about 350 to 520 mm yr-1. Oak woodlands do not achieve much higher rates of carbon assimilation as rainfall increases. During dry years, annual evapotranspiration of oak woodland approaches and is proportional to total rainfall. In wet years evapotranspiration falls well below total rainfall and appears to level out at rainfall levels greater than 600 mm (see Figure 25.7d). It appears that the trees are unable to exploit the surplus rainfall, especially early season rainfall. While it is more convenient to relate Oak Woodlands   519

B. Photosynthesis is inhibited during the summer growing season due to soil moisture deficits.

A. Photosynthesis > respiration 10 Oak woodland

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Leaf area index (m2 m-2)

Oak savanna, 2002 D105-270

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Ppt (mm yr-1)

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E/ Eeq

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weighted by roots (cm3 cm-3)

Mature ecosystems

7 6 5 4 3 2 1 0 1

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[N] P/E eq

FIGURE 25.7 Summary of key relationships that govern carbon and water use in oak woodland ecosystems. See chapter for detailed explanation. Data are annual averages. Sources: (A, B, E) Baldocchi, Chen et al. 2010; (C) Baldocchi et al. 2004; (D) Baldocchi, Ma et al. 2010; (F) Baldocchi and Meyers 1998. A Relationship between gross primary productivity (GPP) and respiration (Reco, ecosystem respiration). B Annual variation in GPP, from January 1 (Julian day=1). C Relationship between evaporation (E, from ecosystem, estimated from net radiation measurements) and GPP. D Relationship between annual precipitation (ppt, mm yr-1) and evaporation (E). E Relationship between soil moisture deficits and transpiration. F Relationship between ecosystem water balance and woodland-scale leaf area index (LAI, m leaves/m soil), r2 =0.67.

annual carbon and water fluxes to annual sums of rainfall, researchers in California and the Mediterranean Basin find that carbon assimilation (GPP) is better correlated with total spring (March–​May) rainfall (Ma et al. 2007, Allard et al. 2008). Oak trees must overcome severe moisture deficit stresses during the hot dry summer—​volumetric soil moisture is often below 5% by volume, causing oak trees to experience predawn water potentials as low as –​7.0 megapascals (MPa) (Baldocchi et al. 2004). They accomplish this in two ways. First, soil moisture deficits induce gradual stomatal closure, yielding a marked reduction in transpiration (see Figure 25.7e). But stomatal closure is not enough. To survive the hottest and driest portion of the summer, oak trees must tap into the groundwater (Lewis and Burgy 1964), where they can extract up to 50% of the water they need to survive during that dry period (Miller et al. 2010). On annual to decadal time scales, the water balance of the system limits the leaf area index of the woodland (see Figure 25.7f). A prominent feature of oak woodlands is its savanna nature, with open canopies and leaf area indices below 2 (Ryu et al. 2010). This openness reduces the amount of energy absorbed by the vegetation and used to drive evaporation.

Ecologists define a disturbance as a relatively discrete event that disrupts ecosystem, community, or population structure and changes resources, substrate availability, and/or the physical environment (Davis and Moritz 2001). Disturbances tend to free up limiting resources such as space, light, and nutrients, triggering successional processes of ecosystem and community recovery. Resilience is a measure of an ecosystem’s ability to recover to its original state when subjected to disturbance. Here we focus on important disturbance processes that operate over stand-to-landscape scales in oak woodlands, including fire, woodcutting, and livestock grazing. Other disturbances such as flooding (notably for valley oak), extreme wind events, snow and ice storms, and local landslides can have important local effects but are not as pervasive in oak woodlands and are not discussed here. We discuss ecosystem conversion to cropland and residential and urban development in a later section.

quently as every one to two years in some areas (Greenlee and Langenheim 1990), although much longer intervals of twenty to thirty years have also been reported (Mensing et al. 1999). Native American burning was probably also responsible for the relatively short three- to seven-year fire return intervals reported for black oak woodlands, where fire was used to promote oaks over pines and perhaps to control acorn pests such as Tilbert worm and Filbert weevil (Mensing 2006; see Chapter 10, “Indigenous California”). The modern fire regime in foothill woodlands is controlled by active suppression and indirectly by livestock grazing that reduces fuel loads. Frequent human ignitions result in numerous small wildfires that are quickly extinguished, especially along busy roads. Long fire-free periods in coastal woodlands can allow the development of shrub understories so that when fires do occur, they tend to be much more severe (Davis and Borchert 2006). Fire suppression in black oak woodlands has dramatically reduced fire frequency and may be leading to an increase in conifers over oaks as well as to an increase in fire severity when fires occur (Mensing 2006, Miller et al. 2009). Oak woodlands are relatively resilient to burning, with coast live oaks the most fire-tolerant and interior live oak the least fire-tolerant species (Lathrop and Osborne 1991). In coast live oak, most large trees can produce new canopies after burning, and most saplings and even some seedlings resprout after being top-killed by fire (Holmes et al. 2008). Blue oak trees will survive low-moderate fires but can be killed in more severe burns. A high proportion of blue oak seedlings and saplings are top-killed but resprout, so that the main effect of burning is to retard progression of saplings to the tree layer (Swiecki and Bernhardt 2002, Allen-Diaz et al. 2007). In montane woodlands, black oaks can survive fires once they attain a 16 cm stem diameter, and top-killed trees usually resprout vigorously after burning. Oak woodland herb communities also appear to be quite resilient to fires. The non-native annual grasses that now dominate the understory produce enormous quantities of seed, enough of which survive lowseverity grass fires to provide full recovery in the growing seasons that follow. Native perennial grasses and bulbs resprout after burning, and populations of native annuals may be relatively neutral, responding more to postfire temperature, rainfall variation, and livestock grazing than to fire occurrence (Wills 2006).

Fire

Tree Cutting for Firewood and Range Management

Lightning is uncommon in foothill woodlands; humans set most fires. This generalization probably applies to prehistoric as well as modern fire regimes. Lighting is more common in montane black oak woodlands, and fire suppression now exerts a strong influence on fire regime in those systems. In both foothill and montane woodlands, fires readily ignite in the dry grass layer during hot dry summer months and tend to be fast moving and of low intensity, with 0.5–​1 m flame lengths (Skinner and Chang 1996). We have scant information from a few analyses of fire scars, pollen cores, and early historical narratives on prehistoric fire regimes in foothill woodlands. Available evidence suggests that some blue oak woodlands were regularly burned by native Californians. Estimates of mean fire return intervals range from seven to fourteen years (McClaran and Bartolome 1989, Mensing 2006, Standiford et al. 2012). Coast live oak and mixed oak woodlands in the Coast Ranges may have been burned as fre-

Because acorns of foothill oaks were a food staple for many native Californian groups, deliberate tree removal seems unlikely prior to European settlement (Rossi 1980, Mensing 2006). During the Spanish Mission era (1770–​ 1830) and Mexican rancho era (1830–​1850), oak was occasionally cut locally for fuelwood and fencing, but no evidence suggests any major effects on oak distribution or abundance (Mensing 2006). Rather, the main impact of Europeans prior to 1850 was to dramatically reduce the size of native Californian populations who for millennia had occupied and, through burning, influenced oak woodlands and specific plant and animal resources in them. The second major impact of early Europeans was the introduction of domesticated livestock. These impacts registered more strongly on the understories than on the tree layer of oak woodlands. After the Gold Rush of 1849, the rapid increase in the number of towns and agricultural expansion probably claimed

Disturbance and Ecosystem Resilience

Oak Woodlands   521

large tracts of oak woodlands near major rivers and on arable valley soils. None of the California oaks has much value for lumber, especially compared to native pines such as ponderosa pine (Pinus ponderosa) or sugar pine (P. lambertiana), but oaks were harvested for fuelwood and charcoal and cleared for cropland. Agricultural clearing was especially widespread in the Central Valley and San Francisco Bay region, where large tracts of valley oaks and live oaks were felled. Further south on the coastal plain and the inland valleys of the Coast, Transverse, and Peninsular Ranges, coast live oaks, valley oaks, and blue oaks were thinned or cleared for row crops and orchards (Kelly et al. 2005, Mensing 2006, Whipple et al. 2011). No evidence of widespread oak cutting in foothill woodlands for fuelwood or range improvement appears until the twentieth century. Beginning around 1930 and continuing for at least another fifty years, a deliberate campaign of woody plant removal in California rangelands, largely conceived and promoted by University of California rangeland specialists, resulted in controlled burning for shrub removal and tree cutting to increase forage production and water yield over millions of acres. The extent of tree removal for range improvement is poorly documented, but between 1950 and 1973 perhaps 900,000 acres of foothill woodlands were cleared (Alagona 2008). The policy of oak clearing has since been reversed on both scientific grounds, as evidence mounted that oaks can improve soil fertility and forage production and as a result of public pressure to protect oaks for wildlife habitat and other amenity values. Oak clearing for range improvement is now discouraged, especially in areas receiving less than 50 cm annual precipitation (Allen-Diaz et al. 2007). Oak cutting for firewood continues but is highly localized. A four-year statewide survey conducted between 1988 and 1992 documented firewood harvesting on roughly 25,000 acres per year, only 0.11% of woodland acreage (Standiford et al. 1996). Intensive coppice management of oak woodlands for fuelwood production has long been practiced in the Mediterranean Basin but is uncommon in California. California oaks appear to be moderately resilient to tree cutting. Most trees including blue oak and interior live oak, which are the primary sources of firewood for commercial harvest, produce basal sprouts after cutting. The rate of resprouting in blue oak declines with tree size, from nearly 70% for trees less than 10 cm dbh to under 20% for trees greater than 50 cm dbh (AllenDiaz et al. 2007, Standiford et al. 2011). Pillsbury et al. (2002) reported 30–​45% resprouting from cut stumps in experimentally thinned stands of coast live oak. Because browsing by livestock and deer promotes multistemmed “hedged” shrub morphology and slows sprout extension to the tree canopy, some form of protection may be needed to promote recovery of the tree layer. Low recruitment of new oaks in most oak woodlands and savannas means that these systems are not very resilient to mechanical tree removal or to tree cutting with herbicide treatment to prevent resprouting. Many areas where trees were deliberately cleared more than a half-century ago remain grasslands today (Brooks and Merenlender 2001). Stand thinning may have a more subtle effect on oak demography by increasing the distance between trees, thereby reducing gene flow by pollination as well as acorn production. Sork et al. (2002) documented short-range pollen dispersal and small effective population sizes in open stands of valley oak. Knapp et al. (2001) reported lower acorn production by more isolated blue oaks. 522  Ecosystems

Tree removal can strongly affect local climate, hydrology, and ecosystem processes. Trees absorb more sunlight than annual grassland, exert a greater effect on air flow and sensible heat transfer, and evaporate considerably more water (Baldocchi, Chen, et al. 2010). Isolated trees in oak woodlands and savannas create microenvironments distinctly different than the surrounding grasslands. The tree canopy reduces light reaching the understory by 25% to 90% and can intercept 5% to 50% of rainfall (Maranon et al. 2009). Conversely, areas of grassland adjacent to oak woodland can produce edge effects in oak woodland microclimates that extend tens of meters into the woodland and affect woodland wildlife communities (Sisk et al. 1997). Tree removal can also have significant effects on soils and nutrient cycling. Soil nutrient levels, soil organic matter, and nitrogen cycling are considerably higher in oak understories than in surrounding grassland, especially in deeper soils where tree roots are able to access deeper pools of water and nutrients than are more shallow-rooted herb species (Callaway et al. 1991). Tree understories also support higher microbial activity, higher canopy dry deposition and animal deposition (perching birds, shadeseeking cattle, deer, and other animals), and reduced nutrient losses from leaching and soil erosion (Callaway and Nadkarni 1991, Knops et al. 1996, Dahlgren et al. 1997, Herman et al. 2003, Maranon et al. 2009).

Livestock Grazing Grazing animals have been part of California grasslands and woodlands for several million years (Jackson and Bartolome 2007), although many large grazers of the late Pleistocene such as bison, llamas, and horses went extinct near the close of the last Ice Age more than ten thousand years ago. Immediately prior to European settlement, the main grazers were pronghorn antelope (Antilocapra americana) and tule elk (Cervus canadensis nannodes). Those species have been eliminated over most of their former California range and now play a negligible role in oak woodlands except in a few areas. Cattle grazing has been the dominant use of oak woodlands since European settlement (Huntsinger et al. 2010). Since that time, cattle operations have depended on natural forage production and thus have been highly sensitive to interannual variability in rainfall and, to a lesser degree, temperature. During the Mission Era, the number of cattle in California rose to approximately 400,000 head. The first major boom in livestock grazing accompanied the Gold Rush and lasted from 1849 to 1868, when the number approached 1,000,000 animals (Olmstead and Rhode 2004). Livestock numbers statewide plummeted between 1868 and 1871 due to severe drought and then partially recovered over the next several decades to a statewide herd of about 470,000 animals. The livestock industry boomed again after World War II to a peak of 3.2 million head in 1976 (Alagona 2008). Numbers have continued to decline, and only 710,000 beef cows and heifers grazed California rangelands in 2011, down from 920,000 in 2001 (see Chapter 37, “Range Ecosystems”). Though commercial ranching has declined in recent decades, more than 80% of oak woodland properties larger than 80 hectares continue to be grazed by livestock, and roughly 90% of those properties practice year-round cattle grazing (Huntsinger et al. 2010). In many ways, today’s oak woodlands are structurally, compositionally, and functionally a legacy of two centu-

ries of continuous livestock grazing. The advent of cattle (and sheep) grazing in the eighteenth century initiated large changes to oak woodland ecosystems due to a combination of factors including the introduction of invasive exotic grasses, reduction in fire frequency, and introduction of new ungulate herds. Cattle trampling increases bulk density of the topsoil, affecting both water infiltration and root penetration (Dahlgren et al. 1997, Herman et al. 2003, Allen-Diaz et al. 2007); cattle also concentrate and redistribute nutrients and increase nutrient fluxes by promoting more rapid turnover of organic matter (Herman et al. 2003).

CAT TLE GR A ZING AND OAKS

Structurally, livestock grazing promotes a more open tree layer by reducing seedling and sapling recruitment and, through repeated browsing, delaying the development of established saplings into the tree layer (Borchert et al. 1989, Tyler et al. 2006, Davis et al. 2011). Grazing impacts on seedlings depend on seasonal grazing patterns and, at least for blue oaks, are especially pronounced during spring and summer (Hall 1992). Over many decades, grazing can lead to declining tree densities if the rate of new tree recruitment is less than adult tree mortality (Callaway and Davis 1993, Davis et al. 2011).

GR A ZING AND WOODL AND UNDERSTORY VEGETATION

Grazing disturbance can suppress shrub recruitment and maintain an understory of annual herbs (Callaway 1992, Huntsinger and Bartolome 1992, Callaway and Davis 1993). Shrubs can facilitate recruitment of some foothill oak species, notably blue oak and coast live oak; over time the suppression of a shrub layer could thus also contribute to declining tree densities (Callaway 1992, Callaway and Davis 1998). Given sufficient time and lack of fire and grazing, shrubs or woody sub-shrubs commonly establish in the grasslands between oaks in savanna or under oaks where canopy cover is incomplete. This has been observed as rapid colonization by coyote brush (Baccharis pilularis) in moist northern California sites where cattle have been excluded (McBride and Heady 1968, Williams et al. 1987, Williams and Hobbs 1989). Grazing exerts a strong influence on the height and cover of both the growing and senesced herb layer, with associated effects on surface energy exchanges and water balances. Contrary to reports from other savanna and grassland ecosystems, grazing does not appear to increase plant growth. Furthermore, current year herbage production declines when grazing pressure (measured as residual dry matter) is higher in the previous year, especially in areas of higher rainfall (Jackson and Bartolome 2007). Contemporary herb communities in oak woodlands and grasslands are relatively resilient to continued light-to-moderate livestock grazing. Grazing can have species-specific effects on understory species and resulting plant community composition and richness, but these effects are hard to disentangle from the effects of large year-to-year weather variation (Jackson and Bartolome 2007). Studies in California grasslands suggest that moderate grazing can increase plant diversity compared to nongrazed or heavily-grazed areas (Bartolome et al. 1994). Native annual forbs may be favored by grazing that reduces competition with taller introduced

annual grasses (Safford and Harrison 2001, Gelbard and Harrison 2003, Hayes and Holl 2003). Livestock grazing can be a viable method for reducing some noxious weeds (Reiner and Craig 2011) as well as potential fire severity.

Land Use Conversion and Ecosystem Fragmentation Agricultural Conversion As described earlier, large-scale woodland conversion to cropland and orchards began in the latter half of the nineteenth century but was mainly limited to arable valley and floodplain soils. By 1900, 6 million to 7 million acres were already in cultivation—​a number that increased to 8 million acres by 1950 and has ranged between 7 million and 8.5 million acres since that time (Olmstead and Rhode 2004). Thus most lowland oak woodland conversion to cropland occurred over a century ago. Rapid growth of California’s wine industry during the last quarter of the twentieth century produced a second wave of agricultural conversion that mainly affected foothill oak woodlands. For example, between 1976 and 2010 vineyard acreage jumped from 52,609 hectares to 198,296 hectares (USDA NASS n.d.). Large-scale vineyard expansion into oak woodlands occurred in many coastal counties, notably Sonoma, Mendocino, San Luis Obispo, and Santa Barbara Counties. Merenlender (2000) estimated that 2,925 hectares of oak woodland were converted to vineyard between 1990 and 1997 in Sonoma County alone.

Urban and Residential Development Since 1950, suburban and rural residential development has transformed many oak woodland landscapes, notably in the San Francisco Bay region, the “Gold Country” of Nevada, Placer and Eldorado Counties, coastal counties from San Luis Obispo to San Diego, and parts of western Inyo and Riverside Counties. Perhaps two-thirds of the new housing development has occurred at urban and suburban densities (0.2–​2 hectare lots) on agricultural lands, many that were formerly oak woodlands (FRPP 2010). However, large-lot suburban and rural residential housing (2–​8 hectare lots) pervades many oak woodland landscapes. Most coast live oak woodlands and forests in coastal foothills from Santa Barbara County to San Diego County have been transformed by residential development (Davis et al. 1995). Rural housing development has been imprinted on foothill oak woodlands of the central Sierra Nevada in what Duane (1999, p. 200) characterized as “low-density, land-intensive, large-lot exurban sprawl” that has altered “both the ecological and the social landscape.” This same exurban sprawl increasingly affects montane oak woodland communities as well. Spero (2001) analyzed census data from 1950 through 1990 to document historical residential development in oak woodlands. Based on population projections, he then modeled future development to 2040 (see Figure 25.7). Recent trends in rural residential development are consistent with or exceed his projections. For example, Gaman and Firman (2006) estimated that by 2006 more than 400,000 hectares of oak woodland habitat had been developed. Stewart et al. (2008) estimated that 84% of new residential development in oak woodlands between 1990 and 2000 occurred in lot Oak Woodlands   523

sizes between 2 and 8 hectares. The woodlands in and around these dispersed residences can be additionally fragmented by roads and fencing, degraded by heavy modification of understory vegetation associated with rural ranchettes, and impacted by domestic pets, impaired air quality, altered surface and groundwater hydrology, and water quality. Fire is half as frequent in rural residential areas as in undeveloped areas (Spero 2001). These and other factors lead to systematic shifts in the bird, mammal, and plant communities of rural residential areas compared to extensive woodlands (Maestas et al. 2001, 2003; Merenlender et al. 2009)

Ecosystem Services The Millennium Ecosystem Assessment (2005) distinguished four types of ecosystem services including provisioning, regulating, supporting, and cultural services. We have already described the importance of oak woodlands for provisioning services such as livestock production, firewood, and game species as well as oak contribution to soil fertility, a supporting service. Here we focus on the role of oaks in climate regulation (carbon storage, hydrology, local energy balance), and on cultural services such as nongame wildlife and aesthetic values (Millenium Ecosystem Assessment 2005, Kroeger et al. 2010, Plieninger et al. 2012).

Carbon Storage, Hydrology, and Climate Regulation Badocchi, Chen, et al. (2010) calculate that blue oak woodlands serve as a small net carbon sink (−92 ± 48 gC m−2 yr−1). This is roughly half the average net carbon uptake reported for other ecosystems in the U.S. and reflects a close balance between carbon uptake via photosynthesis (1,031 gC m−2 year−1) and loss through respiration (939 gC m−2 yr−1). These values are only approximate and vary considerably between years as a function of spring rainfall and seasonal water balance (Ma et al. 2007, Baldocchi, Chen, et al. 2010). Though small on a per area basis, this uptake becomes tangible when integrated over the extent of blue oak woodlands (1.19 Tg yr-1, or about 0.2% of California’s total greenhouse emissions in 2006). Kroeger et al. (2010) suggest that reforestation of cleared oak woodlands could sequester up to 1% of the state’s current emissions annually when averaged over a seventy-five-year period. Other activities with relatively large carbon benefits include restoration of riparian vegetation in oak woodland landscapes and restoration of perennial grasses (Kroeger et al. 2010). As described earlier, oaks can exert a significant effect on site hydrology by intercepting rainfall lost by evaporation from the canopy and by using soil and groundwater below the rooting depths of herbaceous understory plants. For example, Miller et al. (2010) estimated summer groundwater uptake by blue oaks of 15 mm to 23 mm per month. To the extent that soil water, local groundwater, and associated streamflows are desired for other purposes, oaks could be seen as reducing local hydrological services. This could depend on how much groundwater taken up by deep roots is transported to shallow soil layers through hydraulic redistribution during the dry summer season (Gou and Miller 2014). Oaks have complex effects on local climate that could offset their positive effects on the climate system through carbon uptake. Oak canopies

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absorb more sunlight than open grasslands and reduce the rate of airflow by increasing surface roughness, thereby reducing heat sensible heat exchange between the surface and the atmosphere. As a result, average air temperature in blue oak woodland is about 0.84°C warmer than surrounding grasslands, despite the fact that oak canopies provide cool microsites (Baldocchi, Chen, et al. 2010).

Cultural Services In recent decades oak woodlands have been increasingly valued for their environmental amenities such as aesthetic views, nonharvested plant and animal diversity, and the peace and quiet associated with rural rangelands. As documented by Huntsinger et al. (2010) based on surveys conducted between 1985 and 2004, oak woodlands have attracted a new generation of landowners with different values and land management practices than traditional ranching families. Over that period, the percentage of owners that relied on livestock ranching as a major source of income declined from 27% to 14%, and the number of landowners citing “living near natural beauty” as an important factor influencing their decision to live in oak woodlands increased from 46% to 71% of respondents. Furthermore, owners in 2004 reported more reasons to value and plant oaks than in previous surveys, and a lower proportion reported cutting oaks. These trends were especially pronounced among owners with properties less than 80 hectares. Huntsinger et al. (2010) point out that owners on small properties may be satisfied with cultural ecosystem services such as natural beauty, but that these small lots cannot support other kinds of ecosystem services such as maintaining viable populations of harvested and nonharvested wildlife, regulating water quality, or sequestering carbon. Owners of large properties tend to be interested in undertaking environmental improvements that increase ecosystem services at a meaningful scale, but these same owners typically lack financial resources to do so. Compared to other ecosystems, oak woodlands on large ranches appear to allow fuller bundling of (rather than trading off between) provisioning, regulating, and amenity services (Plieninger et al. 2012). However, many of these benefits from extensive woodlands such as water quality, wildlife habitat, and scenic quality accrue off-site rather than to the landowners. As a result, ranchers have not had much incentive to invest in management actions such as oak restoration or riparian protection where private costs may exceed private benefits (Kroeger et al. 2010). Increased private market prices for carbon sequestration and water quality could induce ranchers to supply more of these public benefits. Both Kroeger et al. (2010) and Huntsinger et al. (2010) emphasize that maintaining and managing large properties for provisioning and regulating services will require new funding sources and programs that incentivize conservation. To support such policy formulation, research is needed to better assess the market value of ecosystem services on oak woodlands. Amenity values are challenging to estimate, especially given the diffuse nature of benefits such as nonharvestable biodiversity, viewshed quality, and open space. Standiford and Huntsinger (2012) present several analytical approaches for doing this, including contingent valuation and hedonic regression. Standiford and Scott (2008) used hedonic regression to analyze the effect of oak woodland open space on

urban property values in southern California and showed that both land and home value decreased significantly as the distance from open space boundaries, trailheads, and local stands of native oak habitat increased. These studies indicate the tangible value of oak woodlands to property owners but do not necessarily translate into public funding for new oak woodland conservation. To the contrary, available empirical studies of public willingness to pay indicate that large areas of oak woodland cannot be protected by land acquisition with public funds (Thompson et al. 2002).

Scenarios of Oak Woodlands in the Mid-twentieth Century Changing land use and climate are likely to be the most important drivers of ecosystem change in oak woodlands. These primary drivers could bring other kinds of changes, such as new fire regimes associated with increased population density and longer fire seasons, or local extinctions of species from habitat loss and fragmentation that affect nutrient cycling and ecosystem services. While such changes cannot be forecasted with any certainty, it is possible to construct plausible scenarios for oak woodlands over the next several decades. In this section we focus on the two primary drivers: climate change and land use change.

Climate Change Since 1950, the most significant trend in the climate of the California foothills has been an increase in minimum nighttime temperatures (ca. +0.4°C decade-1) and to a lesser extent in mean annual temperature (ca. +0.3°C decade-1) (Christy et al. 2006, LaDochy et al. 2007, Crimmins et al. 2011). This warming trend is due to both regional land use change, including the expansion of urban areas and irrigated agriculture, and global climate change. Precipitation has also increased slightly during this time at a rate of 3 mm to 4 mm decade -1 (Crimmins et al. 2011), with considerable spatial variation in trends. Interannual climate variability has also increased. These changes in climate have been accompanied by detectable trends in many ecological processes and species dynamics such as increasing wildfire activity (Westerling et al. 2006), changes in plant phenology (Cayan et al. 2001), higher tree mortality rates related to warm drought (Van Mantgem and Stephenson 2007, Van Mantgem et al. 2009), and elevational shifts in species distributions (Kelly and Goulden 2008, Moritz et al. 2008). Few studies have focused on the direct effects of climate change on oak woodlands. McLaughlin and Zavaleta (2012) surveyed sapling and adult valley oak populations across the southwestern part of the species range and found, in a pattern consistent with changing climate, that sapling recruitment was more constricted around surface water than were adult populations. Global climate models all predict that California will continue to warm over the next century, with increases in midcentury mean annual temperatures of 1°C to 3°C compared to the second half of the twentieth century (Hayhoe et al. 2004, Cayan et al. 2008). End-of-century increases range from 2°C to almost 6°C depending on the climate model and emissions scenario used. Precipitation projections have varied

considerably among the climate models, although the most recent Coupled Model Intercomparison results (CMIP5) show much higher agreement than previous model intercomparisons (Neelin et al. 2013). Depending on the region, at least eleven and as many as fifteen out of fifteen CMIP5 models currently project increased precipitation over areas now occupied by oak woodlands (Neelin et al. 2013). Whether this increase translates into lower climate water deficits has not been evaluated. Climate exerts multiple, interacting controls in oak woodlands. Winter temperatures control seed survival, germination, and seedling development. Early spring temperature and humidity affect pollen production and transport. Air temperatures near the ground may become lethally hot for new seedlings. Rainfall timing and amount exert a strong control on seasonal soil moisture availability, plant community composition, net primary production, and nutrient cycling. Many of these climate factors vary at the microclimate scale of centimeters to meters as a function of topography and soils, and are regulated by the vegetation canopy so as to create feedbacks between local climate and vegetation. The combination of very fine-scale controls on climate near the ground and vegetation-climate coupling pose a great challenge to predicting how modern climate change will affect species distributions and associated ecosystem processes. A variety of approaches have been used to investigate how climate change could affect the distributions of ecosystems and species in California. For example, Lenihan et al. (2008) used a dynamic general vegetation model (DGVM) to model coupled climate, vegetation, carbon dioxide, and fire scenarios under warmer/wetter and warmer/drier scenarios. Mixed evergreen forest and grassland increased by 40% to 80% in these scenarios, and mixed evergreen woodland decreased by 10% to 35%. These changes were partly driven by increased wildfire in all scenarios. Kueppers et al. (2005) produced climate envelope models for valley oak and blue oak and projected them forward to late-twenty-first-century climate scenarios using both a regional climate model and a downscaled global climate model. Based on the regional climate model, suitable habitat for both species shrank to less than 60% of current potential habitat, with contraction of the southern range and some expansion into northwestern California and higher elevations of the Sierra Nevada. Sork et al. (2010) sampled genetic variation in valley oak across the range of the species and found a strong association between nuclear multilocus genetic structure and climate gradients. They also showed that regional populations of valley oak occupy significantly different climatic conditions across the species’ range. McLaughlin and Zavaleta (2012) explored differences in modeled species distributions based on current climate envelopes of valley oak saplings versus adults and highlighted the potential significance of drought-mediating microrefugia for local persistence of the species under climate change. These examples all highlight the need for continued study and model improvement to assess the potential vulnerability of oak species (and plant species in general) to ongoing climate change (Morin and Thuiller 2009, Dawson et al. 2011). Projected changes in the distribution and composition of oak woodlands under the stress of future climate must be considered in the context of the variability that these woodlands have experienced historically (Klausmeyer et al. 2011). Stress could vary across the range of a species or ecosystem

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partly because climate change will vary across the area, but also because of geographic differences in the historical range of climate variability, which could influence both community composition and the amount of adaptive variation within populations of a species (e.g., Davis and Shaw 2001, Sork et al 2010). Unusual drought years and wet years could also play an important role in determining oak persistence and redistribution. Critical climate thresholds or triggers associated with oak woodlands are poorly understood, but evidence from other forest and woodland systems suggests they could be significant (e.g., Reyer et al. 2013, Knapp et al. 2008).

Land Use Change California’s Department of Finance projects that over the next thirty years, the state’s population will increase 27%, from thirty-seven million in 2010 to forty-seven million in 2040. Demand for new housing, continued demand from current urban residents for primary- and second-home rural housing, and associated economic pressure to subdivide large ranches all pose significant threats to oak woodland ecosystems. According to Gaman and Firman (2006), roughly 303,500 hectares of oak woodland are at high risk of development by 2040. A recent assessment by the Department of Forestry and Fire Protection classified 506,500 hectares of foothill oak woodland types at medium to high risk of development by 2040 (Figure 25.8). Woodlands at highest risk are in the Sierra Nevada foothills northeast of Sacramento, in Nevada, Placer and El Dorado Counties, foothills of Madera and Fresno Counties, and south of San Francisco Bay in Santa Clara, Contra Costa, and Alameda Counties. The ongoing fragmentation of oak woodlands, climate change, fire regime shifts, and changes in owner preferences and land management make the long-term sustainability of oak woodlands in these areas highly uncertain.

Management and Adaptation Strategies Since 1980, oak woodlands have been priority conservation targets for both state and county agencies and nongovernmental organizations like The Nature Conservancy, the Audubon Society, the California Oaks Foundation, the California Cattleman’s Association, and numerous county land trusts.

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The California Legislature passed the Oak Woodland Conservation Act in 2001, and at least forty-one counties have now adopted policies focused on conservation and sustainable development of oak woodlands (http://ucanr.edu/sites/ oak_range/Description_of_County_Oak_Conservation_Policies/). Public bond funds have been used to acquire highprofile sites such as the Ahmanson Ranch, a 1,200 hectare property purchased in 2003 for $135 million. Some large tracts of oak woodlands have been protected from development by conservation easements and others by multispecies habitat conservation plans and Natural Community Conservation Plans engendered by the federal Endangered Species Act and the California Endangered Species Act. A noteworthy recent example is the 57,870 hectares Tehachapi Uplands Multispecies Habitat Conservation Plan, which encompasses thousands of hectares of diverse oak woodlands on the Tejon Ranch in the western Tehachapi Mountains. Despite some high-profile successes, conservation through land purchases can be contentious and difficult to achieve due to lack of funding or political support. In addition, lack of interest or economic incentive to maintain cattle ranching on large family ranches is leading to subdivision of large ranches, due in part to estate taxes encumbered during intergenerational transfers (Giusti et al. 2004). Easement purchase of development rights has now become the conservation tool of choice for protecting oak woodlands on large working ranches, although easements are not a panacea (Merenlender et al. 2004, Reiner and Craig 2011). Payments for ecosystem services such as carbon sequestration or watershed protection are of intense interest to conservation organizations, but they have had limited application to date, in part for reasons just discussed. Furthermore, many undeveloped oak woodland landscapes have already been subdivided into 20 to 40 acre lots, making large-scale conservation through acquisition, purchase of development rights, or ecosystem service payments infeasible or undesirable. In those areas the best course of action may be educating small-lot landowners about sustainable land management practices (Brussard et al. 2004). Though many techniques have been designed and management practices identified to help sustain or restore oak woodland ecosystems (McCreary 2004), we only touch on a few examples here. Low-cost shelters can be used to protect seedlings and sapling oaks from cattle and other large ungulates. Reduced livestock grazing during summer months also improves survival of young oaks. Progressive livestock ranching techniques such as fencing riparian areas, distribution of water troughs to allow better dispersion of grazing pressure, and retaining adequate end-of-season residual dry matter to conserve soil and promote grassland productivity (mentioned earlier) have all been shown to promote increased biodiversity and ecosystem sustainability (Huntsinger et al. 2007). This is not to say that sustaining oak woodlands is simple or without real costs. For example, control of invasive noxious weeds, feral pigs, and other nuisance species can be extremely expensive and challenging. Furthermore, management outcomes are increasingly uncertain when climate change, air pollution, and fragmentation associated with rural residential development are considered. In principle, the best strategy for managing under such uncertainty engages adaptive ecosystem-based management (Huntsinger et al. 2007, Millar et al. 2007). This requires explicit management goals, formal monitoring of management outcomes compared to control

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areas under a different management practice, and the adjustment of management as learning occurs over time. Such management is scientifically challenging. Time lags in ecosystem responses to management changes, interacting with largerscale changes in ecosystem drivers such as climate and air quality, create complexities that make it difficult to pin management practices to ecological outcomes. Furthermore, while attractive in theory, adaptive management in practice faces hurdles including lack of landowner training, added expense of design and implementation, and limited landowner ability or willingness to change practices (Lee 1999, Aslan et al. 2009).

Research Needs and Priorities Much is known about the ecology and management of California oak woodland, but the body of research is still relatively shallow compared to that for other California ecosystems such as mixed conifer forests or wetlands, or compared to other oak woodland ecosystems such as midwestern and eastern North American deciduous woodlands and forests (Johnson et al. 2002). We conclude by listing a few research areas that, based on our review of the scientific literature, are particularly in need of additional research.

FIGURE 25.9 Projected threat of development to current oak woodlands. The map was produced by combining a map of foothill woodland habitats (extracted from California Department of Forestry and Fire Protection multisource land cover data, 2006) with a map of localized and landscape development threat based on projected population growth, land ownership, current land use, roads, and county general plans.

Oak management and restoration require better knowledge of key demographic parameters including acorn dispersal distances and rates, species-specific age- and size-based sapling survival and growth rates, and adult tree acorn production and mortality rates. To improve climate change adaptation strategies for oak woodlands, research is needed on speciesspecific ecophysiology, microclimate factors controlling seedling establishment and early growth, climate controls on oak mortality, and patterns and scales of genetic climate adaptation. Managing water resources and oak woodland ecosystem services can be improved by ecohydrologic research on shrub-tree-grass interactions and associated spatio-temporal patterns of water availability, and on associated spatio-temporal variation in soil moisture in woodlands in complex rugged terrain. Biodiversity conservation planning and wildlife management lack community-level research on food webs and species interactions—​for example, the role of herbivores versus site physical factors in regulating oak regeneration as a function of climate and soils; the role of top predators such as mountain lions, bobcats, and coyotes on community composition and dynamics; impacts of invasive plants on animal community composition; and mechanisms and controls on community invasibility. For land use planning, fire management, and ecosystem services management, landscape and ecosystem models need Oak Woodlands   527

continued development and refinement to clarify the relationships between fire regime and woodland pattern, composition, and dynamics; effects of low-density residential housing on community composition and ecosystem processes; and the relationships among climate, vegetation structure and dynamics, and transfers of energy, water, carbon, and other nutrients. Finally, to devise effective policy for ecosystem services, socioeconomic research is needed on valuation of regulating, supporting, and amenity services associated with oak woodlands as a function of woodland location, structure, and composition; and on policy design to incentivize private landowners to conserve or increase provision of ecosystem services on their properties.

Summary California’s oak woodlands occupy a wide range of foothill and montane environments and harbor exceptional plant and animal diversity. The open canopy of scattered trees over grass or shrub understories creates high local variation in microclimates and soils associated with tree understories and canopy gaps. The acorn crop is an important food resource for many animal species, and the trees also supply important habitat elements such as trunk cavities and downed wood. Oaks serve as foundation species in these ecosystems in that they exert inordinate control on community and ecosystem processes. This chapter focuses on foothill oak woodlands in general and on blue oak woodlands in particular. The foothill climate combines high temperature seasonality, low winter and early spring rainfall, and summer and autumn drought. Both evergreen and deciduous oaks in these woodlands must cope with extreme summer heat and moisture stress. They do so by gradually reducing gas and water exchange between leaves and the atmosphere over the course of the summer and by tapping deep soil water and/or groundwater. Net primary production is largely controlled by the amount of available water, and in general uptake of carbon dioxide by photosynthesis is only slightly higher than loss of carbon dioxide through ecosystem respiration. Because oak woodlands occupy such a large area (roughly 3.5 million hectares), this small per-area carbon gain translates into significant carbon storage across the entire system. For example, blue oak woodlands are estimated to store around 8.6 teragrams of carbon annually on a statewide basis. Oak woodlands have been significantly reduced in extent by agricultural and residential development. Ongoing suburban and rural residential development poses the greatest immediate threat to remaining oak woodlands, but climate change is also a serious concern. Both land development and climate change threaten to diminish the many ecosystem services provided by oak woodlands such as forage for livestock, important habitat for game and nongame wildlife species, and highly valued scenery. These services are mutually compatible, allowing for bundling of multiple services when designing conservation strategies such as payments for ecosystem services or purchasing of development rights.

Acknowledgments Frank W. Davis was supported in part by the National Science Foundation Macrosystems Biology Program, NSF #EF528  Ecosystems

1065864. Dennis D. Baldocchi was supported in part by the U.S. Department of Energy Terrestrial Ecosystem Science Program.

Recommended Reading Baldocchi D., Q. Chen, X. Chen, S. Ma, G. Miller, Y. Ryu, J. Xiao, R. Wenk, and J. Battles. 2010. The dynamics of energy, water, and carbon fluxes in a blue oak (Quercus douglasii) savanna in California. Pages 135–​154 in M. J. Hill and N. P. Hannan, editors. Ecosystem Function in Savannas: Measurement and Modeling at Landscape to Global Scales. CRC Press, Boca Raton, Florida. Keator, G. 1998. The life of an oak. Heyday Books, Berkeley, California. Pavlik, B. M., P. C. Muick, S. G. Johnson, and M. Popper. 1991. Oaks of California. Cachuma Press, Los Olivos, California. Tyler, C. M., B. Kuhn, and F. W. Davis. 2006. Demography and recruitment limitations of three oak species in California. Quarterly Review of Biology 81(2):127–​152.

Glossary Arborescent  Used to describe a woody plant that at maturity assumes a tree growth form, typically one or a few stems and greater than 5 meters in height. Basal area  The total cross-sectional area of tree stems in a given land area, usually measured 1.3 meters above the ground surface, and often expressed in square feet per acre (ft2/ac) or in square meters per hectare (m2/ha). Climate water deficit (CWD)  An important bioclimatic variable, CWD is the difference between potential evapotranspiration (PET), which is the amount of water that would have been lost, had it been available, and actual evapotranspiration (AET) as limited by available soil water (Stephenson 1998). Coppice management  A method of woodland management in which trees are cut and allowed to resprout from the stumps for future reharvest. Ecosystem engineer  A species that directly or indirectly controls the availability of resources to other organisms by causing physical state changes in biotic or abiotic materials (Jones et al. 1994). Eddy covariance  Often referred to as the eddy flux method, eddy covariance is an important technique for measuring vertical turbulent fluxes in the lower atmosphere. Knowing rates of vertical and horizontal air movement as well as properties such as gas concentrations, temperature, pressure, and humidity, it is possible to estimate important ecosystem processes such as the exchange of water vapor and carbon dioxide between vegetated land surfaces and the atmosphere. Foundation species  A species that by virtue of its structural and/or functional properties plays a defining role in the biotic composition and ecosystem processes associated with an ecological community (Ellison et al. 2005). Insect gall  Abnormal plant growth induced by insects through injection of chemicals, usually into leaf or twig tissues, by larvae or adults. Keystone species  A species that is high in the food web and whose large effects on the community are highly disproportionate to its abundance. Leaf area index (LAI)  An important descriptor of ecosystem structure and predictor of ecosystem function, LAI is the onesided area of green leaf surface per unit of ground area and is a dimensionless quantity that ranges from less than 1 to greater than 7 for some forest types.

Masting  Used to describe the more or less synchronous production of large quantities of seeds by members of a plant population preceded by a long period of time without high production. Many oaks exhibit masting every few years. Photosynthetic capacity (Amax)  The maximum rate at which leaves are able to fix carbon from atmospheric carbon dioxide during photosynthesis, typically expressed as µmol m-2 s-1. Photosynthetic efficiency  The percentage of incoming light energy that is converted into chemical energy by photosynthesis. Typical values for plant canopies range from 0.1% to 2%. Specific leaf area (SLA)  The ratio of leaf area to leaf dry weight (e.g., square millimeters per milligram [mm 2 per mg]). SLA is an important plant functional trait related to photosynthesis, leaf longevity, nutrient retention, and water use. Stomatal conductance  The rate of exchange of either water vapor or carbon dioxide between a plant and the atmosphere through the stomata, or small pores of the plant. Usually expressed in mm s-1 or mmol m-2 s-1.

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T WENT Y-SIX

Coast Redwood Forests HAROLD MOONE Y and TODD E . DAWSON

Introduction The coast redwood forest of California is a major component of the state’s northern forest region, which extends northward from about 150 kilometers south of Big Sur into southern Oregon. As an ecosystem type, the coast redwood forest is both globally unique and historically fascinating. Blanketing the northwest corner of the state, it dominates the western faces of the Coast Ranges, the coastal marine terraces, and the alluvial plains that drain some of the most pristine watersheds in the region. Redwood forest is also found in the most southern and immediately coastal sections of the Klamath Mountains. In this chapter we do not cover in detail the variants of the redwood forest forest type that occur in drier sites away from strong coastal influences. We further cannot cover the remarkable and unusually diverse vegetation of the Klamath Mountains, which lie mostly between the immediate coastal region and the Cascade Range to the east. We point those interested to learning of the exceptional diversity of the vegetation of the Klamath Mountains to John Sawyer’s Northwest California: A Natural History ( 2006) and to the classic vegetation study of this region by Robert Whittaker (1960) as well as the recent resampling of Whittaker’s original analysis by J. Grace et al. (2011). The montane and subalpine forests even

further to the east, in the Cascade Range, are a continuation of the vegetation found in the Sierra Nevada to the south and are not included here but are reviewed in Chapters 27, “Montane Forests,” and 28, “Subalpine Forests.”

Physiographic Setting Geological History and Features The coastal zone of northwestern California has a diverse and complex geological history characterized by a legacy of tectonic and some volcanic activity and a predominantly sedimentary geomorphological influence (see Chapter 4, “Geomorphology and Soils”). These features have shaped the existing topography, watershed diversity and hydrology, and vegetation composition. Many abrupt boundaries between vegetation types of the region are clearly linked to its complex geology and soil development. Tectonic and volcanic activity first set the stage for the region’s diverse rock and soil types. However, sedimentary processes, marked metamorphic events, and the current moist climatic regime have largely 535

created patterns of erosion and associated landscape features, such as landslide-associated river systems.

SOILS AND SOME UNIQUE PL ANT ASSOCIATIONS

The region’s unique coastal zone geology has also left a distinctive mark on the soils and soil development of the region (see Chapter 4, “Geomorphology and Soils”). Soils are generally quite shallow in the steeper topographic settings but become deeper and well developed in valley bottoms and even on some of the coastal marine terraces. Soils that have developed on serpentine and peridotite have low fertility but do support a diverse assemblage of coniferous tree species (including some endemics) such as Sargent’s cypress (Hesperocyparis sargentii), Macnab’s cypress (Hesperocyparis macnabiana), and knobcone pine (Pinus attenuata). Some montane conifers—​Jeffrey pine (P. jeffreyi), sugar pine (P. lambertiana), ponderosa pine (P. ponderosa), and incense-cedar (Calocedrus decurrens)—​a lso can be found, but no hardwoods inhabit these infertile soils. Marine terraces of diverse ages caused by changes in sea level, just adjacent to the coast, provide a remarkable illustration of the impact of time on the development of soils. Growing on these soils is a diversity of vegetation types that are closely associated with specific soil properties that are derived from the same substrates and under the same climate regime. The time of origin of each marine terrace is the sole variable in the development of the vegetation types that are found. These range from grasslands on the youngest soils, to redwood–​Douglas-fir forests on more mature soils, to stunted trees like Pygmy cypress (Hesperocyparis pygmaea) and a subspecies of lodgepole pine, Bolander’s beach pine (Pinus contorta ssp. bolanderi), on the oldest soils. Over hundreds of thousands of years, accumulated nutrients are leached out of the soil and the clay minerals are moved downward, forming a hardpan that leads to very poor drainage and thus an unsuitable environment for normal tree growth (Jenny et al. 1969). This unique topographical setting, or “ecological staircase,” is found in coastal Mendocino County near the village of Caspar. Because of its uniqueness, it is now largely protected as the Jug Handle State Natural Reserve. At the extreme northern end of the California coast, the forest takes on a physiognomy that more closely resembles the great temperate rainforests of the Pacific Northwest, with vast, alluvial deposits along river courses supporting the tallest forests on the planet. Tree diversity is low and dominated by a mixture of the widespread conifers Douglas-fir (Pseudotsuga menziesii) and coast redwood (Sequoia sempervirens) with some Sitka spruce (Picea sitchensis), western hemlock (Tsuga mertensiana), Port Orford–​cedar (Chameacyparis lawsoniana), and even (at or on serpentine soil contacts) western white pine (Pinus monticola).

General Vegetation Extent and History of the Pacific Coastal Forest Northern coastal forests are part of a rich evergreen coniferous forest belt found running along the western edge of North America and the Pacific coast. It begins as patchy forPhoto on previous page: Coastal redwood forest in northwestern California near Humboldt Redwoods State Park. Photo: Anthony Ambrose. 536  Ecosystems

ests along stream channels of the Santa Lucia Mountains in south-central California about 200 kilometers south of San Francisco, then becomes more continuous over the land surface beginning in the Santa Cruz Mountains and extending northward through the coast ranges into the states of Oregon and Washington and finally into western Canada. In past geological times, this forest type possessed even greater species richness than today because it also included many more deciduous tree species closely related to those found today (e.g., Acer, Fraxinus, Salix, etc.). These taxa had the greatest alliance to trees from along the eastern coast of North America as well to many taxa still found in Japan and China. By the late Pliocene, however, deciduous tree diversity had been significantly reduced. This was probably associated with intensification of a winter-wet, summer-dry Mediterranean-type climate (MTC) (see Chapter 8, “Ecosystems Past: Vegetation Prehistory”). As in the other MTC regions of the world, evergreen plant life forms dominate the vegetation and thrive under the drought- and fire-prone conditions that prevail today. Coniferous trees dominate the forests, with a rich flora of other woody evergreen species—​often shrubs—​in the understory that are able to maintain positive carbon gain and necessary nutrient levels throughout the year. Even during the cool winter months when soil moisture is highest, nutrients are available and temperatures are above freezing. Because deciduous trees have no leaves during the winter months (November through March or April) when metabolism would be favorable, they experience a shorter carbon gain period than the rich evergreen elements in the flora. Further, the needle-shaped leaves of conifers track air temperatures more closely than do larger, deciduous leaves, allowing the former to maintain a more favorable energy balance. Moreover, when deciduous-leaved taxa experience water deficit, because of their generally larger leaf size, their leaf temperatures can rise above ambient temperatures by 3°C to 10+°C due to poor convective heat loss, resulting in tissue damage (Waring and Franklin 1979). One remarkable feature of the vast majority of conifers that dominate the Pacific coastal forest is the great age and size they attain (Table 26.1). Western red cedar, redwood, and Douglas-fir can all exceed twelve hundred years of age, and redwood and Douglas-fir grow to 75 meters to 110+ meters in height. In 2013 the redwood named Hyperion, which is the tallest known tree in the world and is found in Redwood National Park, was measured to be 115.8 meters (380 feet) tall, or 22 meters (72 feet) taller than the Statue of Liberty (S. Sillett, personal communication). These same coniferous species also can accumulate standing aboveground biomass that rivals or exceeds the great angiosperm-dominated forests of the Amazon, the Congo Basin, and Southeast Asia. Within the Californian northwest coastal region, in addition to the unusual forest types found on specialized soils noted earlier, are a variety of other ecosystem types including communities that contain a significant fraction of the companion species to redwood, Douglas-fir or tanoak (also called tan-bark oak) (Notholithocarpus densiflorus), or other coastal conifers such as Sitka spruce, Port Orford–​cedar, and grand fir (Abies grandis). At the southerly end of these great forests one can even find them intergrading with chaparral, northern coastal scrub, grassland, and oak woodlands and savannas (see Chapters 22, “Coastal Sage Scrub”; 23, “Grasslands”; 24, “Chaparral”; and 25, “Oak Woodlands”). In general, redwood-dominated forests give way to Douglas-fir and tanoak to the immediate interior and further to the east, where con-

TA B L E 2 6 .1 Age and dimensions of some conifers of California northwest forests

Typical

Maximum

Age (years)

Diameter (cm)

Height (m)

Silver fir Abies amabilis

> 400

90 to 110

44 to 55

Port Orford-cedar Chamaecyparis lawsoniana

> 500

120 to 180

60

Incense-cedar Calocedrus decurrens

> 500

90 to 120

45

> 542

368

Engelman spruce Picea engelmannii

> 400

> 100

45 to 50

> 500

231

Sitka spruce Picea sitchensis

> 500

180 to 230

70 to 75

> 750

525

Sugar pine Pinus lambertiana

> 400

100 to 125

45 to 55

Western white pine Pines rnonticola

> 400

110

60

615

197

Ponderosa pine Pines ponderosa

> 600

75 to 125

30 to 50

726

267

Douglas-fir Pseudotsuga menziesii

> 750

150 to 220

70 to 80

1,200

434

Coast redwood Sequoia sempervirens

> 1,250

150 to 380

75 to 100

2,200

501

Western red cedar Thuja plicata

> 1,000

150 to 300

> 60

> 1200

631

Western hemlock Tsuga heterophylla

> 400

90 to 120

50 to 65

> 500

260

Mountain hemlock Tsuga mertensiana

> 400

75 to 100

> 35

> 800

221

Species

Age (years) 590

Diameter (cm) 206 359

306

Source: Waring and Franklin 1979.

ditions are much less mesic. At the eastern edge of the more lush coastal forest belt one finds Douglas-fir and ponderosa pine with increasing dominance. With increasing elevation in the Coastal Ranges, one encounters white fir (Abies concolor) and Shasta red fir (Abies magnifica var. shastensis) in the montane regions. Further upward in elevation, mountain hemlock (Tsuga mertensiana) is an important dominant in the subalpine zone (Sawyer 2006) (see Chapter 28, “Subalpine Forests”).

The Redwood Forest Because of its iconic beauty, recreational and economic value, and biological uniqueness, the redwood forest has attracted considerable attention by the general public, biologists, conservationists, and the forest industry (Noss 2000) (see Chapter 36, “Forestry”). Despite global recognition of the unique properties of the coastal redwood forest, information on the

physiology, ecology, and forest-scale processes that characterize the dynamics of this system is not extensive. In recent decades our fundamental understanding about redwoods and the forest ecosystem has increased, especially in light of environmental pressures such as land use, urbanization, and climate change that pose real threats to the long-term viability of redwoods. Our focus in this chapter is on the nature and dynamics of the remnants of the pristine redwood forest. Redwood forests are distributed in a narrow coastal strip of central and northern California extending from the southwest corner of Oregon at latitude 42°09’N southward to southern Monterey County at 35°41’N (Olson et al. 1990) (Figure 26.1). This extensive latitudinal range includes areas with a mean annual rainfall from approximately 600 millimeters in the south to well over 3,000 millimeters (Olson et al. 1990). This fact along with its paleo-distribution, encompassing former summer rainfall climates, poses somewhat of an enigma. However, because of their great stature, redwood trees create their own local climate and favorable microcliCoast Redwood Forests   537

FIGURE 2 6.1. Distribution of redwood (Sequoia sempervirens) indicating locations of remaining old-growth stands. Data courtesy of the Save the Redwoods League and Cal Fire, Fire Resource and Assessment Program (FRAP). Map: Parker Welch, Center for Integrated Spatial Research (CISR).

mates that are more constant than the variability of rainfall received throughout its range would indicate. In the redwood forest mean annual temperatures are relatively cool all year and mild compared to other forest ecosystems in the state, with temperatures that range between 5°C to 25°C. Frost or snow is very rare and patchy in winter. Sumer heat waves that exceed 30°C are also very rare, though these have been more common since the mid-1980s (Fernandez et al. 2012; Johnstone and Dawson 2010).

Principal Organisms

A A

B B

Tanbark Tanbark oak oak

Madrone Madrone

PL ANTS

The redwood forest is dominated by its namesake. This forest does, however, include a set of commonly associated tree species that are generally more broadly distributed than redwood itself. Among these tree taxa are California bay (Umbellularia californica), tanoak, and Douglas-fir. The distributions of these trees overlap redwood’s range but are not coincident (Figure 26.2). Douglas-fir has quite a wide distribution and is an important component of the Pacific Coastal Forest. California bay and tanoak are predominately Californian and extend into many drier regions of the state than does redwood. Pacific madrone (Arbutus menzeisii) is most abundant from central coastal California north through Washington and British Columbia, with scattered populations in the Sierra Nevada as well as the mountains of Baja California in the south. This discordance in distributions is also apparent in the understory plants. For example, the very prominent and characteristic herbaceous perennial redwood sorrel (Oxalis oregana) is found not only with the coastal redwood but also throughout coastal Oregon and into Washington. The characteristic evergreen huckleberry (Vaccinium ovatum) occurs in the redwood forest as well as in fir-spruce, Douglas-fir, and lodgepole pine forests and is found north into British Columbia as well as in the Sierra Nevada.

C C

DouglasDouglasfir fir

D D

California California bay bay laurel laurel

FIGURE 26.2. Distribution of some of the major co-occurring tree species found in the redwood forest. (U.S. Geological Survey 1999). Note the disparate ranges of key “members” of the redwood forest.

EPIPHY TE S AND ENDOPHY TE S: UP IN THE CANOPY

Although the diversity of higher organisms is relatively low in redwood canopies, a dramatic diversity of epiphytic species (organisms growing with no direct connection to the ground) (Williams and Sillett 2007) occurs at these heights. The complex structure that reiterated trunks can provide can result in large catchments for leaves and canopy debris that develop into soils able to support a wide range of epiphytic plants. These canopy or arboreal soils have a high water-holding capacity (Ambrose 2004, Sillett and Van Pelt 2007). and can be colonized by leather fern (Polypodium scouleri) and the evergreen California huckleberry as well as a rich community of fungi, invertebrates, and even a salamander (Aneides vagrans) (Figure 26.3). In a study of 9 ancient redwood trees, Williams and Sillett (2007) found 256 species of epiphytes that included 100 species of microlichens, 69 macrolichens, 17 cyanolichens, 45 bryophytes, 19 liverworts, and 26 mosses. They attributed this richness to the great diversity of microhabitat types that these plants favor within these tree crowns. For example, some of these species were found associated principally with burned bark, others on burned wood. Although the species diversity of epiphytes was high, their

biomass was low compared to that found in other tree species of the coastal forests. Further, an ecologically important epiphytic form, the nitrogen-fixing cyanolichens, are scarce in redwoods although they are abundant in old-growth Douglas-fir forest and serve as in important nitrogen input to that system (Denison 1973, Pike et al. 1977). One possibility is that herbivore-deterring toxins, commonly associated with redwoods, could inhibit biomass growth of epiphytes as well as perhaps of such specialists as the cyanolichens (Williams and Sillett 2007). A class of microbes, endophytes, is also associated with the canopy of redwoods and is as species-diverse as in other coniferous forests. Some of these species could be endemic to redwood foliage. Fungi associated with the exterior and interior of redwood leaves are numerous but also show a very patchy distribution. These endophytes might play diverse roles in the redwood canopy; many have been found to be mutualists or latent pathogens (Espinosa-Garcia and Langenheim 1990). A recent investigation also shows that these leaf endophytes likely play a role in the absorption of water that accumulates on the leaf surfaces of redwood trees (see Burgess and Dawson 2004). Ongoing investigations are showing Coast Redwood Forests   539

FORESTS IN THE AIR Hundreds of feet above the ground, the crowns of ancient redwoods shelter another forest. Thickets of berry bushes, ferns, and other conifers--some large enough to bear cones--rise from dense mats of soil on broad limbs or in trunk forks

STELLER’S JAY

REDWOOD TRUNK GROWING FROM LIMB

The soil, as thick as three feet, forms from decayed leatherleaf ferns and redwood leaves and bark, nourishing an aerial ecosystem unknown until the 1990s, when scientists first climbed into the canopy.

RED HUCKLEBERRY

YELLOW CHEEKED CHIPMUNK

COPEPODS FUNGI EVERGREEN HUCKLEBERRY

CANOPY SOIL

MARBLED MURRELET SALAMANDER

FIGURE 2 6.3. A complete ecosystem high above the ground in the canopy of a mature redwood. Source: Boume 2009. Reprinted with permission from National Geographic.

that the fungi obtain a small fraction of carbon from the tree leaves themselves; this represents a new and undocumented type of plant-fungal mutualism (K. M. Lader, T. E. Dawson and C. D. Specht, unpublished data).

BELOW THE GROUND: MYCORRHIZ AE

In addition to free-living microbes inhabiting soils is a very large number of mycorrhizae—​fungi intimately associated with the roots of higher plants and able to form a symbiotic relationship with them. Mycorrhizal hyphae (very thin strands of the fungi that colonize a large volume of soil) mine and transport mineral nutrients and water from the soil into the plant in exchange for energy in the form of carbon (sugars) that the host plant produces. Three major types of the fungi have this type of relationship with plants: ecotomycorrhizae, arbuscular mycorrhizae, and ericoid mycorrhizae (Molina 1994).

ECTOMYCORRHIZAE (ECM):  These consist of a fungal sheath, or mantle, formed on the short, fine-feeder roots of plants. The mantle serves as a storage body for nutrients transported from the soil. These fungi, generally basidiomycetes and occasionally ascomycetes, penetrate plant roots between the cortical cells and have hyphae that connect to the soil and extend outward from the root into the soil. These fungi often produce fruiting bodies, mushrooms—​many of which are prized as edible. Ectomycorrhizae are found in Douglas-fir as well as tanoak. Bergemann and Garbelotto (2006) estimated that eighty-three species of fungi species provided mycorrhizal associations in tanoak. Similarly, Kennedy et al. (2003) found many of species of fungi forming ECM connections with 540  Ecosystems

Douglas-fir and further showed that some of these species are the same as found on tanoak.

ARBUSCU L AR MYCORRH IZ AE (AM):  These do not form a fungal sheath around the roots but penetrate the roots and move into cortical cells, forming finely branched hyphae that proliferate each single cortical cell. The hyphae also penetrate the soil but do not produce aboveground fruiting bodies; rather, they form numerous spores and some fruiting bodies beneath the ground. They, like the ecotomycorrhizae, transport soil nutrients, especially phosphorus and nitrogen, into the host plant in return for carbon from the host plant. The coast redwood has AM fungi associated with its fine roots.

ER ICOID MYCORRHIZ AE:  These are restricted to species of the family Ericaceae. In the redwood forest they are associated with very abundant understory shrubs Gaultheria, Rhododendron, and Vaccinium. These mycorrhizal fungi are associated with young roots and occupy the epidermal cells of these roots. This mycorrhizal type is unique in that it can mobilize nitrogen directly from organic sources, bypassing the longer path of decomposition to a more reduced form of nitrogen. Pacific madrone, although a member of the Ericaceae, does not have standard ericoid mycorrhizae; rather, it has an intermediate type that shares morphological features of ecotomycorrhiza and ericoid mycorrhizas and has been termed “arbutoid mycorrhizas.” Like tanoak, Pacific madrone may share mycorrhizal networks with Douglas-fir. Kennedy et al. (2012) suggest based on shared species of mycorrhizae with Douglas-fir that Pacific madrone, which vigorously resprouts after disturbance including fire, serves as a reservoir of beneficial fungi to the nonsprouting Douglas-fir as it regenerates from seed.

Wildlife The fauna of the coastal redwood forest shows its strongest alliance to the vast coniferous forests in the Pacific Northwest. At first glance many biologists have incorrectly concluded that redwood and other coastal forests are depauperate of animals. Upon closer inspection, and if both vertebrates and invertebrates are tallied, their wildlife diversity rivals many other ecosystem types in the state. One reason the redwood forest seems like a “zoological desert” is that so much of its diversity is harbored in the forest canopy, well out of sight to most observers. Additionally, a sizable fraction of the wildlife is cryptic, nocturnal, or rare. Some animals are federally listed as threatened or endangered; the most notable are the marbled murrelet (Brachyramphus marmoratus) and the northern spotted owl (Strix occidentalis caurina).

INVERTEBR ATE S

The vast majority of species found in the redwood forest, and particularly in old-growth redwood, are invertebrates. Snails and slugs are common in this forest type and include the second largest slug in the world, the Pacific banana slug (Ariolimax columbianus). In contrast, ants, butterflies, and bees are uncommon. Pollination is commonly performed by flies, moths, and beetles. Ancient redwood, Douglas-fir, and Sitka spruce trees offer a wide variety of habitats for insects. Many canopy invertebrates have very specialized habitats, preferring to live in very old trees with narrowly restricted distributions. Moreover, enough canopy invertebrates occur in the large moss mats found in and on old-growth trees to support the dietary needs of resident clouded salamanders (Aneides ferreus), which may live their entire lives in the tree crowns. For this reason concern exists that there has been considerable loss of species with the loss of the old-growth forests (Cooperrider et al. 2000). Although many insects are found in the redwoods, most do not get their food directly from the tree itself. No insect is apparently capable of killing a redwood tree—​a record not matched by any other tree species in North America. The cooccurring Douglas-fir has thirty or more species of bark beetles that attack it, whereas the redwood has only four. Redwood does not produce much resin, a common insect deterrent, so protection against pests or pathogens likely comes from other chemical defense compounds. Thus, remarkably for such a long-lived tree inhabited by short life-span insects, evolution by insects in breaching defensive compounds does not appear very successful (Snyder 1992).

VERTEBR ATE S

The redwood forest is comparatively depauperate in birds, with only six nesting species. However, approximately one hundred species of birds can be found in the redwood forest at one time or another. The remainder of the vertebrate fauna inhabiting redwood stands is not generally distinctive from that found in the coastal forests along the entire western edge of North America and has no apparent endemic full species (but see below). As might be expected for a moist, cool, and high-humidity environment, redwood and Douglas-fir forests are rich in amphibian species numbers and overall abundance. During the breeding season, as many as ten dif-

ferent species co-occur in the same general habitat. The most abundant amphibians are salamanders. Reptiles, due to their relatively high heat requirements, are uncommon and found mostly in open habitats such as forest edges and sunny river courses. Although the redwood forests have been in existence for a very long time, their fauna has changed significantly through the ages except for their herpetofauna, which has relatively ancient members (or taxa that have evolved from them). The mammals of the coastal coniferous forests have had a complex history—​w ith waves of species colonizations, particularly from Eurasia, occurring through time. Many of the mammals seen today have only been residents since the post-Pleistocene. Eighteen mammalian carnivores have historically inhabited the redwood region, most with wide distributions. One of these, the grizzly bear (Ursus arctos), has been extirpated in all of California. Another, the Humboldt marten (Martes americana spp. humboldtensis), is found only in mature redwood forests and was thought to be extinct but is now quite rare and in danger of extinction. This subspecies is genetically distinct from the marten found in the Sierra Nevada. Similarly, the Pacific fisher (Martes pennanti), once with a wide distribution, has one population in coastal redwood forest in addition to the Sierras, where they are thought to be in danger of “imminent extinction.” A number of individual fishers have been transplanted from the coast population to the Sierra (Cooperrider et al. 2000). Mazurek and Zielinski (2004) compared the vertebrate denizens of “legacy” old-growth trees left in harvested stands to young redwoods in these stands that were matched for habitat. The legacy trees were characteristically “battlescarred”—​that is, with at least some of the features of oldgrowth trees such as cavities, deeply furrowed bark, reiterated crowns, and basal hollows (often called “goose pens” for the function they served early foresters as cages for their geese and other livestock inside of the massive, burned-out hollows at the bases of previously burned old-growth redwood trees). Mazurek and Zielinski (2004) found a significantly larger number of bird and bat species associated with the legacy trees. In general, except for bats, only a few small mammals showed a higher association with legacy trees: the wood rat and two insectivores. In total, thirty-eight vertebrate species were associated with legacy trees and twenty-four with young trees. The study concluded that legacy trees represent a “lifeboat” of assemblages for regenerating stands.

Forest Structure In many ways the northernmost temperate redwood forests are comparable to the rain forests of the tropics. While they lack the biotic diversity tropical forests harbor, they share a degree of structural complexity. The redwood forest is layered with plant species of differing heights and light requirements. The composition of these layers varies with site and particularly with latitude. In the well-studied old-growth redwood forest of Muir Woods, McBride and Jacobs (1977) reported that the upper tree layer is composed principally of redwood and some Douglas-fir. A second tree layer consists of shorterstatured trees that also have shorter life spans, including California bay, tanoak, coast live oak (Quercus agrifolia), bigleaf maple (Acer macrophyllum), and red alder (Alnus rubra). Still lower is a rich layer of shrubs consisting of western azalea (Rhododendron occidentale), California hazel (Corylus cornuta Coast Redwood Forests   541

100

Sequoia Pseudotsuga Tsuga Acer Umbellularia Corylus Vaccinium Rubus Polystichum Oxalis

90 80

Height (m)

70 60

50 40 30 20 10 0 0

10000

20000

30000

40000

Crown volume (m3 ha-1) FIGURE 2 6.4. Vertical distribution of crown volume (5-meter-deep slices for volume) in an old-growth redwood forest on a welldeveloped, low-elevation site in Humboldt County. Small and lowdensity species not shown. Source: Sillett and Van Pelt 2007.

californica, thimbleberry (Rubus parviflorus), sword fern (Polystichum munitum), and California huckleberry. Below that, a relatively species-rich herb layer is composed of perennial species such as redwood sorrel, redwood violet (Viola sempervirens), wild ginger (Asarum caudatum), fairy lantern (Disporum smithii), western trillium (Trillium ovatum), California fetid adder’s tongue (Scoliopus bigelovii), miner’s lettuce (Montia perfoliata), and Andrew’s clintonia (Clintonia andrewsiana). Finally there is a moss layer characterized by the Oregon eurhynchium moss (Eurhynchium oreganum). Although there can be considerable structural complexity from the top to the bottom of the old-growth forest, the vast majority of the forest biomass is redwood (Figure 26.4). On the forest floor of unmanaged forests is a considerable amount of fallen dead wood of varying ages. This common feature is quite different from that observed in secondary forests and even some of California’s state parks (Figure 26.5). The explorations of Jedediah Smith in 1828 indicate the difficulty of traversing mature redwood forests on deep soils: when the massive trees fall, “they often take other trees with them, leaving jackstraws of massive logs over older piles of rot-resistant logs. Often the forest floor was such a tangled mass that the men could not see the ground. They scrambled over the logs, but horses had to be led around them.” They estimated that it would take a month to traverse 65 kilometers through such country (Sawyer 2006).

Forest Dynamics Redwood trees of the coastal forests commonly grow taller (over 100 meters) and live longer (more than two thousand years) than any other tree species in North America. In a word, redwood forests are truly unique, so it is not surprising that they violate common rules of forest succession. For example, in many forests one expects the seeds of a late successional species to be relatively large and able to establish in the shady understory that they have created. However, redwood seeds are very small and rarely establish in nature with542  Ecosystems

FIGURE 26.5. Forest floor at the Humboldt Redwood State Park, with researchers at the base of the tree preparing instruments for measuring canopy light climate. Note the more typical forest floor of this forest with considerable dead wood on the floor of differing ages. Photo: Anthony Ambrose.

out the aid of disturbances caused by flooding, fire, or tree falls that create bare mineral soil able to support new seedling germination. Lorimer et al. (2009) reconstructed the history of fire, flooding, and landslides in redwood forests and their role in forest regeneration. They concluded that episodic establishment within the forest on micro-disturbances, such tree-fall root mounds or on decaying logs was sufficient to maintain the redwood dominants, particularly coupled with their vegetative resprouting characteristics and their great longevity. In light of these features, yearly or even decadal seedling input is not required for maintenance. An ongoing and novel perturbation in the structure of redwood forests is occurring due to the invasion of the pathogen sudden oak death (SOD) since the mid-1990s. This pathogen has resulted in death or severe dieback of tanoak in many regions (see Chapter 13, “Biological Invasions”). The cause of this disease, a relative of the one that caused the Irish potato famine, is a new species of the fungus-like protist microorganism Phytophthora ramorum. It was simultaneously found infecting Rhododendron shrubs in the Netherlands and most likely was exchanged between continents by the nursery trade (Garbelotto and Rizzo 2005). SOD has been found to infect a wide number of plant species in the redwoods. An associate of tanoak in redwood forests, California bay, is in part responsible for the epidemic movement of this pathogen because it can become infected and lead to high spore production and relatively long-distance spore dispersal without itself dying. The loss or impairment of tanoak by SOD most likely will have ecosystem-level impacts due to tanoaks’ extensive mycorrhizal networks (Bergemann et al. 2013). The rapidity of change will be determined in part by the abundance of California bay in individual stands (Cobb et al. 2012). Projections suggest that the redwood tree will benefit from loss of the tanoak competitor (Waring and O’Hara 2008). Another dimension of this marked change in the forest is the impact it can have on the amount of standing dead biomass of tanoak in many forests; these dead trees represent an additional fuel source for wildfires (Kuljian and Varner 2010) that could increase the vulnerability of redwood forest to damage. Changes in the fire regime along with other critical global environmental changes (warmer and also drier

future climates) are likely to alter fire regimes (intensity and duration) of and in redwood forests in the future. In a recent study of an unusual episode of fires throughout the redwood region due to dry lightning events in June 2008, although both redwood and tanoak resprout after fire damage, the bole survival of redwood was greater than tanoak (Ramage et al. 2010). So, while redwood clearly has survived under many different climatic and site conditions over its very long history, it might also benefit under some new and increasingly widespread conditions. This depends, however, on a continued supply of water through fog or rainfall.

REGENER ATION

Old-growth redwoods can produce millions of seeds ha-1 of forested land (Boe 1961). However, these seeds are quite small (averaging about 2 milligrams per individual seed), and generally fewer than 10% of those dispersed are viable. Of those viable seeds produced, a high percentage will actually germinate. Seed disperses in late fall and can peak in early winter. Seed output varies though time, with periods (years) of high seed production every five to seven years in some localities, mostly in the central and southerly part of the redwood range (Sloan and Boe 2008). In northern stands little to no interannual variation in seed production occurs, and many years can pass without a single viable seed being produced. The small seeds do not disperse very far, in contrast to wind-dispersed seeds. Viable seed that hit the forest floor are generally shortlived; redwood forests therefore have little to no seed bank. If seeds land in a moist area, they will germinate, whether on a log, on litter, or on the rare patch of bare soil. Seed will also very likely be consumed by numerous seed predators before germination—​birds (juncos, towhees, and song sparrows) and mammals (deer mice [Peromyscus maniculatis] and brush rabbits [Sylvilagus bachmani]) in particular. A host of fungal pathogens can also attack both seeds and seedlings, and if they do, they can largely eliminate regeneration. Those seeds that fall on mineral soils free of pathogenic fungi have the best chance of success. Such sites are found where fire has consumed the litter layer and where bare soils are produced by tree fall and flooding that brings silt over the site (McBride and Jacobs 1977). If and when seeds do germinate, the resulting seedlings have another barrier to overcome before transitioning into saplings: herbivory. Pacific banana slugs, pack rats (Neotoma fuscipes), and brush rabbits are among the animals that can do the most notable damage (McBride and Jacobs 1977). Despite the low odds of success that a redwood seed and seedling have for completing successful germination and development into a healthy sapling, survival still appears adequate to maintain viable populations and persistent forests. Redwoods are not completely dependent on sexual reproduction for long-term survival and have a remarkable capacity for stand maintenance through vegetative reproduction (i.e., resprouting from basal burls or even from the main trunk) when injured by disturbances like fire. The long lifespan of redwoods and their resistance to insects and pathogens results in forests with the potential for persistence well into the future as long as further harvesting of old growth or rapid environmental changes do not challenge them anew. If the remaining redwood forests are subject to science-based conservation and management, their future could be bright. Although an individual redwood tree can live for millen-

nia, its actual lifetime can be longer because it can readily respond to catastrophes by resprouting. Even in the seedling stage, it is already “preparing” for the possibility of the loss of the shoot by the formation of a lignotuber, or a swelling of the stem containing growth buds, produced at the site of the cotyledons (Del Tredici 1999). With age the lignotubers can become enormous. The rings of saplings that surround harvested mature redwood trees, so commonly seen in young redwood forests and many old-growth forests as well, are the result of resprouting from the lignotuber. On the stem, after injury, lignotuber-like structures are formed and are generally called “burls.” These, like the ground lignotuber, contain suppressed buds and can sprout under certain conditions such as injury. In addition are dormant epicormic buds beneath the bark that will sprout following fire injury. The remarkable regenerative power of redwood even extends to establishment of new “daughter” plants by layering (a partially buried stem can generate roots and new stems, and large branches that fall from old trees can regenerate an entirely new tree). When we think of regeneration, we generally look to the forest floor. In the redwood forest we also need to look up. A lot of regeneration of the canopy takes place due to storm events. One of the redwood giant trees, known as the Arco Giant, illustrates this phenomenon: a storm in January 1998 broke off the top of this tree, and a new crown regenerated at the point of breakage (Van Pelt 2001).

FIRE FREQUENCY AND RE SPONSE

Fire return intervals in redwood forests predictably decrease from very wet northern coastal stands (125 to 500 years) to those in the drier, southern, and interior localities of its distribution (about 50 years). However, considerable variance has occurred around these means through history. Frequent burning of prairies adjacent to the redwoods by indigenous Californians caused some ignitions of the forests at the transition zone that did not spread to a great degree due to the cool, moist forest interiors. The thick, insulating bark of older redwood trees can protect them from ignition to some degree if it retains moisture. However, if dry, it will burn, resulting in partially hollowed trunk sections called “goose pens” at the bases of trees—​hollows that can become larger through recurring, low-level fires (Stuart and Stephens 2006). If trees do burn, entry points can emerge for fungal infection into otherwise healthy sapwood. Brown and Baxter (2003) postulate that prior to human occupation of the redwoods, fires in this mesic ecosystem were not common. Fires can stimulate reproduction but are certainly not a prerequisite for the establishment of new seedlings. Disturbance gaps within the forest can also provide sites for reproduction. The uneven age structure in mature, natural forest stands is a testimony to the patchy nature of recruitment opportunities. The long life­ span and unique life history characteristics of coast redwood therefore mean that in a sense this species is a “pioneer” as well as a “climax” species in the forest that it largely defines (Lorimer et al. 2009).

Redwood Forest Energy and Carbon Balance Redwood represents an unusual keystone species in that it is the primary controller of the microclimate essential for a host of dependent species of this forest type. Light, temperature, Coast Redwood Forests   543

Height within tree (m)

83 53 43 33 23 83 53 43 33 23 83 53 43 33 23

0:00

Temperature (ºC) 24 20 16 12

Relative humidity (% RH)

80 60 40 20

Incident light (µmol m-2 s-1) 1500 1000 500 0

6:00

12:00

18:00

Time FIGURE 2 6.6. Temperature, relative humidity, and incident photo­ synthetically active radiation measured throughout the canopy of a redwood forest. Source: Tolle et al. 2005.

humidity, and the overall moisture regime are all influenced by the presence of these massive forest trees. Light levels at the forest floor are very, very low (1–​3% of full sun), and sun flecks are of a short duration (Figure 26.6). Plants at the bottom of the forest floor are thought to be specialized for efficient use of the light they do receive (Santiago and Dawson 2014). One of the most common redwood understory herbs is redwood sorrel, which survives in an environment that receives only about 0.5% of full sunlight (Bjorkman and Powles 1980). This truly shade-tolerant plant reaches 90% of its saturated photosynthetic rate at only 100 quanta m-2 s-1 of incoming radiation (1,600 quanta m-2 s-1 is full sunlight) (Powles and Bjorkman 1980). Their biochemical adaptation to these very low light levels conversely puts them under stress if full light flecks impinge upon them. When this happens, the trifoliate leaves fold up to reduce the amount of full radiation received to prevent damage to their shade-adapted photosynthetic machinery. This response happens in a matter of minutes but can fully reverse after the passing of the sun fleck, over a period of an hour or more. The very low energy received at the forest floor throughout the year puts constraints on adaptive possibilities. Even though specialized plant species can thrive in low-light environments, once established the likelihood of receiving enough energy to go from seed to seed in one year is low. No annuals grow in the full-shade microsites. This is comparable to alpine environments, where, due not to light limitations but to temperature constraints, few annual plants occur. (In the circumpolar arctic tundra is only one annual plant species). Redwood forests are world-class in terms of carbon accumulation and storage. Their remarkable storage is the result not of exceptional rates of carbon accumulation in any given year but rather of the longevity of the tree. Busing and Fujimori (2005) reported on the carbon stores in an old-growth stand of redwoods in Bull Creek, in Humboldt Redwoods State Park, occurring on an alluvial flat. The stand they investigated was dominated (99% by basal area) by redwoods, some exceeding 90 meters in height. California bay, tanoak, and Pacific yew were also in the stand but as subordinates. Their estimates of stem biomass ranged from 3,000 to 5,200 Mg ha-1 544  Ecosystems

depending on how stem volume was calculated and which wood-specific gravity was used. Using stem diameter increment (growth) during the interval from a previous study of the site in 1972 (Fujimori 1977), their most accurate estimate of annual stem net primary productivity was between 4 and 5 Mg ha-1 yr-1 with total tree, aboveground net productivity of 7 to 10 Mg ha-1 yr-1. The forest floor contained 262 Mg ha-1 of coarse woody debris and 5 Mg ha-1 of fine woody material. These values indicate a uniquely high value of standing biomass and moderate-to-high annual productivity compared to other forests of the world. In a different location, Sillett and Van Pelt (2007) measured total aboveground biomass distribution in a 1 hectare Prairie Creek stand. They calculated a static biomass value of 4,283 Mg ha-1 exclusive of snags and logs. Forest stands such as these represent the highest biomasses per unit ground area found anywhere in the world. The leaf area for this forest was 14.2 m2 m-2 of forest floor. As of 2014, newly obtained data point to even higher biomass and carbon stores than these previously published results (Robert Van Pelt, pers. comm.). Annual litterfall in old-growth redwood has been measured at a range of sites from 3,120 to 4,690 kg ha-1 yr-1 with decomposition rate constants of 0.177 to 0.238 yr-1 (decomposition rate constant = annual litterfall/total forest floor litter mass). The total litter mass in one forest ranged from 15,700 to 30,000 kg ha-1 (Pillers and Stuart 1993). Decomposition rates were more related to moisture availability than to temperature, which is probably representative of most California ecosystems where the warmest temperatures coincide with a lack of moisture.

Fog and Coast Redwood Forest Water Balance Redwood trees evolved and expanded their range across many temperate zones over their 100+ million-year history (Mao et al. 2012) well before the appearance and strengthening of the Mediterranean-type climate with its long summer drought. Given the shrinkage of the past distribution of this species to the narrow (40–​70 kilometers wide) and long (approximately 750 kilometers) strip along coastal California where fog is frequent, it seems reasonable to conjecture that summertime fog and the water subsidies it can provide have replaced at least some fraction of the missing summer rain redwoods once enjoyed during warmer and wetter times in the past. The several million years that the Mediterranean-type climate has been in place in California have been marked by cool, wet winters, with generally ample rainfall and rainless summer months. During June through October each year, when the coastal upwelling of very cold, deep ocean waters along the Pacific edge strengthens, fog is most abundant (Figure 26.7). Fog banks form most nights and move onshore as the warmer air inland raises. When the fog is intercepted by vegetation, and especially the towering and deep crowns of the coast redwood trees, the water that drips from tree crowns can account for over 33% of the total annual hydrological input (precipitation + fog). Using the hydrogen-stable isotope composition of fog water, it was determined that fog was taken up by the plants and that 35% to 80% of all the water used by plants within the understory community came from fog (Dawson 1998, Limm et al. 2009). Quantification of fog water inputs to coast redwood forests in Humboldt County, near Orick and Redwood National and State Parks, has continued since 1991 in order to quan-

Rainfall (mm)

300 250

A

200 150 100 50 0

Fog days (#)

15

B

10 5 0

Jan

Feb

Mar

Apr

May

June

July

Aug

Sep Oct

Nov

Dec

Month FIGURE 26.7. (A) Rainfall amount and (B) fog frequency by month in the North Coast redwood country. Source: After Dawson 1998.

tify long-term trends and variation in fog water input relative to winter rainfall and to relate trends to broad-scale climatic factors. In parallel, research focused on looking a broad-scale trends in fog along the entire California coast has shown that overall fog frequency has declined by more than 30% since 1951 and that fog inputs vary quite dramatically from year to year (Johnstone and Dawson 2010). Despite the overall decline in fog, twenty-three years of collector data show that 33% ± 7% of all water inputs to these forest ecosystems come as fog in the summer (Dawson, unpublished data). Interestingly, it is clear that not only are these summertime water inputs important but that the tree crowns themselves are providing the impaction surfaces for collections and drip to occur. This is evidenced by the fact that when trees were removed, fog inputs declined significantly (Dawson 1998). In these logged sites, overall site humidity decreased 17–​36% and temperatures increased 3–​5ºC due to greater surface heating. These changes in local microclimate appear to affect overall rainfall inputs slightly but to reduce fog water inputs by more than 65% (Dawson, unpublished data). Redwood needles can also directly take up a small fraction of their summer daily water requirement from fog. Using both field observations and isotope labeling of water, Burgess and Dawson (2004) showed that 6–​8% of the water that condenses onto foliage each foggy night enters the leaf directly. While this input is low in comparison to fog drip into the soil from tree crowns, it is being absorbed directly into the sites where is can best be utilized. As such, its physiological impact is disproportionally high compared with fog drip that must be taken up by roots and elevated through these towering giants to leaves 50–​110+ meters above the ground before it can be used. Beyond contributing to water supplies, fog also reduces the evaporative demand from leaf surfaces, both day and night, and hence reduces overall transpiration and increased tree and forest water-use efficiency in the high, diffuse light

environment that can prevail on foggy days. Foliar uptake of water has been shown to enhance photosynthesis in coast redwood up to threefold compared to plants that do not receive such wetting events (see Simonin et al. 2009). Finally, although redwood trees in any particular stand will receive the same amount of precipitation input (from vertically delivered rainfall), they differ individually in how much fog they trap as a function of location. Since fog moves horizontally from the ocean onto land, trees on the edge of a forest stand trap more fog than those in the interior (Ewing et al. 2009). The capacity of leaves to absorb fog water, as well as rain and dew, directly has also recently been found in a majority of the redwood understory species studied to date (ten to twelve species). This direct foliar uptake improves plant water balance as well as reducing nighttime transpiration (Figure 26.8).

Nutrient Dynamics Redwood forests, not unlike other large-statured forests, contain the largest fraction of their nutrients within the longlived, massive trees. For example, approximately 80% of the carbon and phosphorus at the old-growth site in Humboldt State Park resides in the trees (Zinke et al. 1979) (Table 26.2). This means that when the trees are harvested, a large amount of the nutrient capital is removed from the site, particularly phosphorus. Dahlgren (1998) studied the nitrogen dynamics of an eighty-year-old, secondary-growth, redwood–​Douglasfir stand at the Caspar Creek experimental watershed in the Jackson Demonstration State Forest of Mendocino County. During the eighty-year period, the forest had accumulated 1,480 kg ha-1 of nitrogen (N). Harvest of this forest would remove 950 kg ha-1 of these stores (wood plus bark) from the site. Additional site losses of nitrogen would occur by streamwater flux, particularly suspended sediment flux, due mainly to harvesting activities (80–​160 kg ha-1). Inputs are meager, Coast Redwood Forests   545

Polystichum munitum (POMU)

Pseudotsuga menziesii (PSME)

u

gn

u

gn

1 cm

1 cm

SESE gn

u

Sequoia sempervirens (SESE)

gn

u

1 cm

u

1 cm

Polypodium californicum (POCA)

gn

Lithocarpus densiflora (LIDE)

gn

u

PSME 1 cm Gautheria shallon (GASH)

1 cm Umbellularia californica (UMCA)

ARME

Vaccinium ovatum (VAOV)

gn

u

UMCA

1 cm

u

gn

LIDE

u

gn

Increasing benefit from leaf wetness

Arbutus menziesii (ARME)

1 cm Oxalis oregana (OXOR)

u

gn

POCA POMU OXOR GASH

VAOV

1 cm

1 cm

FIGURE 26.8. Proportional foliar water uptake capacity (down arrows, U) and proportional suppression of nighttime transpiration loss through stomata (up arrows, g n) for ten redwood forest species. During leaf wetting, increased foliar hydration in many of the dominant broadleaf, coniferous, and fern species of the redwood forest ecosystem occurs. Down-arrow thickness represents the foliar uptake capacity of each species relative to the maximum capacity measured and up arrows illustrate water conservation when leaf wetting stops nocturnal water loss through stomata, with arrow thickness representing the nocturnal conductance rate of each species relative to the maximum rate measured. Species are ranked in order of how influential foliar uptake may be for leaf hydration relative to the suppression of nocturnal conductance when leaves are wet. P. munitum is ranked first because it demonstrated the highest ratio of foliar uptake capacity to nocturnal conductance and U. californica and O. oregana are ranked last because no foliar uptake capacity was measured. All illustrated species experience leaf wetting either in the canopy of redwood forest, where fog impaction and interception occurs first during fog exposure, or on the forest floor, where occult precipitation delivers fog water after canopy foliage saturates. Crown silhouettes on the left indicate the relative position of each species within the redwood forest profile. Source: Limm et al. 2009.

with atmospheric deposition adding 20 kg ha-1 during the eighty-year period by rough calculations. Using a detailed sampling array, higher deposition values were measured at a more southerly site in Sonoma County somewhat nearer to developed areas. The latter study measured a mean value over a three-year sampling period of 0.78 kg N ha-1 yr-1 during the fog season (June through October) and 1.62 kg N ha-1 yr-1 during the rainy months (November through May; total of 3.3 kg N ha-1yr-1). Dissolved nitrogen, ammonium, and nitrate were depleted as fog and rain moved through the canopy to the forest floor (Ewing et al. 2009). Some controversy occurred regarding nitrate availability for tree growth in mature redwood forests (Bradbury and Firestone 2007). Early studies implied a great lack of nitrate in old-growth redwood forest soils, which would have severe implications for reproduction (Florence 1965; Bollen and Wright 1961). These studies were based on the lack of nitrate accumulation when the soils were incubated. Given the earlier results, Stone and Vasey (1968) proposed that active management is needed in old-growth stands to maintain them. They specifically recommended not controlling major floods by damming, since this would restrict inputs of fresh silt to the habitat and provide more available nutrients on a fresh surface for establishment. They further noted that old redwood trees are particularly tolerant to flooding events and can generate a new root system (Figure 26.9).

Aquatic Systems: Land-Water Connections Virtually all vertebrates of redwood forests have ties to estuarine habitats—​most stands contain surface streams with connections to the sea. Historically, marine nutrients have also been delivered to coastal and inland forests by migrating salmon and seabirds (Merz and Moyle 2006). In turn, forests provided habitat and resources to organisms inhabiting the streams (see Chapter 33, “Rivers”). Streamside trees interact strongly with the streambed either by contributing deadwood or as rooted trees in the streamside itself. These features alter streamflow and can form a partial dam in the stream, serving as a sediment trap, or can block upward migration of fish. Large, downed redwood material can remain in place for over a century. The resulting streamflow alterations, sediment movement, and entrapment by the woody material provides habitats for diverse organisms. Logging has also had profound and long-term residual impacts on stream dynamics and on the organisms that reside there. For example, Ashton et al. (2006) found that headwater streams located in old-growth redwood forests maintained higher numbers and densities of amphibian species than did those in successional forests thirty-seven to sixty years postharvest. They attributed the biotic dissimilarities not to differences in thermal environments provided by the vegetation cover but rather to differences still evident in sediment loads of the streams even after a long period after perturbation.

Ecosystem Services Redwood forests have provided benefits to societies since humans first occupied California. California’s northern coastal indigenous peoples used redwood for building canoes and as timber for houses. They harvested the fruits of tanoak and hunted deer and elk common in the redwood region.

TA B L E 2 6 . 2 Nutrient content in trees in an old-growth redwood stand at the Humboldt State Park About 80% of the carbon and phosphorus reside in the trees (values are in g m-2)

Carbon Foliage

Nitrogen

Phosphorus

1,234

10.4

1.09

Wood

62,000

97.6

6.2

Bark

19,000

69.6

0.1

Total tree

82,192

177.6

7.4

Stem

(% of site total) Litter Soil (to 1m) Forest site total

84 1,660

16 19.8

77 1.8

14,361

907

0.3

98,213

1,104

9.5

Source: Zinke et al. 1979.

However, the main food for indigenous peoples of the redwood region was obtained from the coast, salmon-filled rivers, and more open vegetation adjacent to deeply wooded areas (Lightfoot and Parrish 2009). The redwood forest has also been considered the central resource of California’s “second Gold Rush,” providing lumber for the early development of the city of San Francisco. The “gold mine” of favored heartwood of these trees is no longer available. However, regrowth of these forests still provides a substantial amount of highly favored building material. Moreover, these forests sequester large amounts of carbon and might be the greatest forest biomass accumulators on Earth (Busing and Fujimori, 2005). In an era when carbon sequestration in forests can serve as a potential mitigation strategy against steady increases in anthropogenic CO2 emissions, this service is increasingly valuable. Though the forests redwood trees inhabit are limited in geographical extent and land cover, they provide a promising sequestration vehicle, particularly if integrated with sustainable timber harvest practices where a sizable fraction of land is being actively used to grow redwood. Redwood forests also provide a number of regulating services, including erosion control. This service is particularly critical in the redwood forest region because it is often marked by steep topography with highly erodible substrates. The visible effects of erosion caused by logging these forests in an unsustainable manner in the past motivated the establishment of Redwood National Park. The redwood canopy overstory provides other unique regulating services related to microclimate control, such as fog-water harvesting, that enhance habitats for other service-providing organisms of the forest. Finally, the cultural services provided by these unique and awesome forests inspire not only the residents of the region but also visitors from around the world. It was the threats to these forests that played an important role in building the U.S. conservation movement starting with the work of the Save the Redwoods League (see Chapter 1, “Introduction”). Coast Redwood Forests   547

Douglas-fir

Ocean fog

Coast live oak So u

pe

g cin -fa

Madrone

th slo -fa pe cin g

slo

th

r No Resprouting redwood stump

Bay Ferns Redwood burl

Seedling in humus-free soil

New roots after burial by flood Soil moisture 18% in August

FIGURE 26.9. Regeneration of a new root system by a redwood tree subsequent to burial by silt from flooding. Source: Bakker 1985.

History of Redwood Exploitation The remaining vast, old-growth redwood forest is only a very small fraction (4–​5%) of what it once was. The first use of redwoods in construction after settlement appears to have been for construction of the Missions Santa Clara, Dolores, and Presidio of San Francisco. Subsequently, the Russian settlement at Fort Ross, starting in 1812, used local redwood for the construction of the fort and for several boats (which were unsuccessful due to lack of proper seasoning). San Francisco utilized redwood timber for construction, at first from San Mateo County but by 1820 mainly from Marin County. The first waterpower mills appeared in 1834 in Sonoma County; however, it was the advent of the steam engine–​powered sawmill at Bodega, also in 1834, that changed the industry. Sudden demand for lumber after the Gold Rush fueled a surge in the development of the forest industry. Lumbering and milling accelerated, and the opening of Humboldt Bay as a lumber source resulted in nine sawmills in operation there by 1854. As demand for lumber grew, lumber was loaded onto ships, often precariously by cable suspensions, from sixty “landings” along the north coast from Bodega to Humboldt (Figure 26.10). This era left behind stump-filled landscapes throughout the region (Figure 26.11). By 1860, Humboldt County, the second-ranking county in timber production in the state, was producing 71,000 m3 (30 million board feet) yr-1, mostly redwood. Caterpillar tractors first appeared at Klamath Bluff in 1925 and power saws first made their appearance in the 1930s. After World War II the lumber industry accelerated; tractors blazed roads into logging areas, and big trucks with trailers carried logs to the mills. These practices caused considerable physical damage to watersheds. Most loss of soil was not due 548  Ecosystems

to harvesting the trees per se but to the accompanying road building. Harvesting rates continued to climb. In Del Norte County annual timber harvest grew from 125,000 m3 (53 million board feet) in 1946 to 710,000 m3 (300 million board feet) in 1953 and included Douglas-fir. The latter figure was beyond replacement capacity and raised considerable alarm by conservationists (information in this section from Bearss 1965).

The Fight to Save the Redwoods Early Conservation The history of the efforts to conserve the coastal redwood forests is a dramatic one, and as it extends back over a century it reflects the development of conservation efforts in the country as a whole. This history also indicates that some of the strong, early conceptual battles are still with us and that a strong philosophical position can bring strength but also the danger that rigidity can work against achieving goals. The first redwood park established was Big Basin State Redwoods Park in the Santa Cruz Mountains in 1902, culminating the lobbying efforts of a group of twenty-six citizens of the local area who founded the Sempervirens Fund to accomplish this goal (Yaryan et al. 2000). The park has grown from 1,540 hectares (3,800 acres) when established (also as the first California State Park) to approximately 7,300 hectares (18,000 acres) today. In early state history the Santa Cruz Mountains were heavily logged because they were so close to the building boom occurring in San Francisco following the Gold Rush. Between 1850 to 1890 most of the area’s redwood trees were harvested. An irony is that many of these old-growth trees, as lumber in buildings, went up in smoke in the fires that

FIGURE 2 6.10. Loading schooners with redwood logs and lumber for transport to San Francisco from Westport Landing, Mendocino County. Estimated date of picture is 1882. Photo from the Kelley House Archives, courtesy of Carolyn Zeitler and Nancy Freeze.

raged periodically in that city and culminated with the massive fires associated with the 1906 earthquake. A state bond issue in 1909 provided resources to open up the northern riches of the redwood forests to public—​a nd to private—​interests with the construction of the Redwood Highway. Timber harvested before the construction of this overland pathway was transported overland to seagoing vessels. Actual construction of the new highway was not completed until 1923, when through traffic in Del Norte and Humboldt Counties became a reality. The state required the counties to turn over the land needed for the highway without logging it first. Thus only that logging necessary for the roadway construction took place; this was also the case in subsequent highway route relocations (Bearss 1969). Opening the prime redwood forests to easy transportation sent a warning signal to the growing conservation community of the increasing threats to these forests. Three distinguished conservationists—​Madison Grant, Henry Fairfield Osborn, and John Campbell Merriam—​ i n 1917 traveled north to investigate the new country opening up. Grant was a founder of the New York Zoological Society and Osborn was on the staff of the New York Museum of Natural History. Merriam was a professor of paleontology at the University of California in Berkeley, later to become the director of the Carnegie Institution of Washington (Schrepfer 1983). The excursion prompted the three to establish the Save the Redwoods League in 1918. Near this time, William Kent from Marin County, a member of Congress and leader in the establishment of the U.S. National Park Service (1916), purchased the last old-growth stand of redwoods in Marin County and donated it to the U.S. government. Kent called for the formation of a Redwood National Park in 1913. Unfortunately his vision was not to become a reality until a half a century later, in 1968 (Schrepfer 1983). A philosophical schism between the leaders of the Save

FIGURE 26.11. A stump-filled, deforested landscape in the 1890s. Mad River area, Humboldt County. Note resprouting. Courtesy of Humboldt State University Library.

the Redwoods League and those of the National Park Service and the Forest Service in the early twentieth century mirrored broader debates in conservation at the time. Both of the agencies took a more utilitarian view of wildlands than was acceptable to the leaders of the Save the Redwoods League. Gifford Pinchot, the second director of the U.S. Forest Service, promoted “wise use of natural resources based on science;” the Park Service policy of 1924 stated that National Parks must “be conserved in their natural state so that coming generations, as well as the people of our own time, may be Coast Redwood Forests   549

assured their use for the purpose of recreation, education, and scientific research.” In 1938 the Save the Redwoods League rejected a National Park Service proposal for a Redwood National Park on grounds that the Park Service was pandering to public desires, detracting from the inspirational and educational value of the parks, and destroying their primitive nature (Schrepfer 1983). The efforts of the Save the Redwoods League have since gone into redwood forestland acquisition in support of the development of sixty-two parks and preserves encompassing nearly 73,000 hectares. The dramatic escalation of redwood harvesting rates in the 1950s made clear that the philosophy of solely saving patches of the biggest, old-growth trees on lower alluvial sections would lead to jeopardy. Record winter rains in 1954–​ 1955 caused the loss of hundreds of giant redwoods at Bull Creek Flat in Humboldt Redwoods State Park. This, along with restructuring of the Redwood Highway cutting through this park that caused more loss of trees, highlighted the fragile nature of old-growth forest protection. By the end of the 1950s only 10% of the original, old-growth redwood belt remained (Noss 2000). These events led to involvement by another conservation group, the Sierra Club, in the redwood protection movement—​w ith a new approach. Its leadership concluded that federal involvement was needed to meet the crisis, and they revived the movement to form a Redwood National Park. They started an aggressive campaign to bring this about, which included the publication of two books. The first documented the enormous losses that were occuring (Hyde and Leydet 1963), while the second advanced a specific site proposal for the formation of a National Park (Leydet 1969). This effort introduced a new, more aggressive approach to conservation—​education and publicity followed by lobbying and litigation—​u nder the leadership of the late David Brower, whose work contributed to the enormous accomplishments in environmental legislation in the 1960s and 1970s. The renewed battle to form Redwood National Park initially pitted the Sierra Club (which proposed a landscape-selection approach for the park—​Redwood Creek, Prairie Creek Redwoods State Park) against the Save the Redwood League (which advocated saving the oldest trees at Bull Creek, Humboldt Redwoods State Park). Most of all, it was a battle between the forest industry and conservationists waged slowly in Congress, during which a great deal of additional old-growth forest was lost. In 1968 the park was finally approved at Redwood Creek (23,000 hectares, but only 4,400 hectares of old growth). A battle ensued to enlarge the park to protect the forests there from continued harvesting in the surrounding area. In 1978, Congress approved a second National Redwood Park Act that added 19,500 hectares to the park as well as an upstream protection zone of 12,000 hectares. Susan Schrepfer (1983) documents the long and dramatic saga of the formation of Redwood National Park.

What Will the Future Bring? If the climate of today were to persist, changes to the dynamics of California’s redwood forests would nevertheless ensue from pertubations such as invasive pathogens. These forests are, however, remarkably resilient ecosystems because of the capacity of many of the dominant plants to respond to damage by resprouting. Redwoods themselves are remarkably resistant to insects, disease, and even fire. While it is unfor550  Ecosystems

tunate that we have lost such a vast extent of old-growth redwood forests, we have protected, thanks to the intensive efforts of many individuals and groups, representive oldgrowth stands of trees of extrordinary heights and biomass. The bulk of redwood forests are in private hands and are managed as production forests. Because these holdings are larger than in the past, there is the opportunity for planning at the landscape level. Further, the public owns three times as much area in young-growth redwood forests than it does in old growth, presenting opportunities to develop innovative, sustainable practices for the whole forest ecosystem. California’s Forest Practice Act of 1973 was passed in response to the harvest practices that led to loss of more than 90% of mature redwood forests. The act has changed forestry practices on both private and public lands, with particular attention to protecting riparian areas. Selective harvesting is becoming more common, and some private companies are managing habitat for endangered and critical wildlife (Hartley 2012) (see Chapter 36, “Forestry”). The primary threat to the future of these forests is the fate of the cool, temperate climate, including foggy summers, that favors redwoods. A critical “controller” of fog production is the cold water associated with the California Current. If the current warmed, we would lose one of the critical elements behind fog and low cloud formation along the coast (Johnstone and Dawson 2010, Schwartz et al. 2014). Fog and cloud cover have already declined over the past sixty to one hundred years; it is not clear whether this will continue (Johnstone and Dawson 2010). With global warming, climatevegetation models predict a shift from the prevailing, moist coniferous forest species toward evergreen forest dominants currently on more xeric (e.g., south-facing) slopes (see Chapter 14, “Climate Change Impacts”).

Summary The redwood forest region of California occupies the coastal plains and mountains of northwestern California. The region is rich in its diversity of ecosystem types—​a feature driven mainly by topographic and substrate diversity. North coastal forests, which extend into Oregon and Washington, are remarkable for their diversity of conifers for their longevity. This heavily forested region is relatively sparsely occupied by humans. The high rainfall of the area’s uplands feeds a large number of rivers that flow into the Pacific, including some in a wild condition, a rarity for California. This chapter focuses on the iconic redwood forest ecosystem, located mainly on ocean-facing slopes and plains. This forest, in its primal state, attains world records for tree heights and biomass accumulated during the millenial lifespan of its dominants. This forest type had a vast distributional range in prehistoric times, when a summer rainfall climate existed in California. With the disappearance of summer rainfall, this ecosystem is now restricted to the cool temperatures of the north coast, where fog occurs in the summer. Redwood trees trap fog water to support their own growth and provide moisture from drip to understory plants as well as tree roots. The future of this ecosystem under ongoing climate change will depend largely on the fate of the cold California Current, which has a marked impact on patterns of winter rainfall and generates both the summer drought and the conditions leading to summer fog. Because of the very large leaf area of these forests, and

their predominately evergreen nature, very little light falls on understory plants, particularly at ground level. The inhabitants of these areas have specialized capacity for capture and utilization of light energy. The relatively stable, cool, and moist conditions of the forest support a diversity of fungi that evidently play a key role in ecosystem nutrient balance through symbiotic relationships. Further, high atmospheric moisture and diversity of tree trunk surfaces, and soil development in tree crevices, provide habitat for a large number of epiphytic plants and animals. Unlike in the Sierra Nevada, where much land was put into forest reserves and parklands early in the history of the state, the vast majority of California’s northern coastal forests remained unprotected until relatively recently. Only in the past several decades was Redwood National Park established, although earlier efforts provided some conservation areas elsewhere in the state. Today only 4–​5% of old-growth redwood forests remain, virtually all of it protected. The large area of young forest, under both public and private ownership, is now being managed in a more sustainable manner with increased attention to the dynamics of the entire redwood ecosystem.

Acknowledgments We thank George Koch, Steve Sillett, and Anthony Ambrose for advice and help on this chapter as well as the Save the Redwoods League for the years of support that made many of the redwood investigations cited in this chapter possible.

Recommended Reading Evarts, J., and M. Popper, editors. 2001. Coast redwood: A natural and cultural history. Cachuma Press, Los Olivos, California. Noss, R. F., editor. 2000. The redwood forest: History, ecology, and conservation of the coast redwoods. Island Press, Washington, D.C. Sawyer, J. O. 2006. Northwest California: A natural history. University of California Press, Berkeley, California. Schrepfer, S. R. 1983. The fight to save the redwoods. University of Wisconsin Press, Madison, Wisconsin.

Glossary Bole  The main trunk of a tree. Cyanolichens  Lichens that have a blue-green algal component and can fix atmospheric nitrogen. Ectomycorrhizae  Fungi that form a sheath around the fine roots of woody plants and trade nitrogen and phosphorus from the soil in exchange for carbon of the host plant. Epicormic buds  Latent buds that can sprout from the bark in response to tree damage. Epiphytes  Plants that are not attached to the ground but rather are attached to branches. They acquire their water and nutrients from the atmosphere. Ericoid mycorrhizae  Fungi that associate with the roots of plant members of the Ericacea family in a symbiotic relationship. Some of these fungi have the capacity to obtain nitrogen directly from decaying plant material. Legacy trees  Old-growth trees that remain in a forest that is predominately second-growth. Overstory  The trees that form the upper canopy of a forest.

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eral Technical Report PSW-151. Albany, California: U.S Forest Service. Noss, R. F., editor. 2000. The redwood forest: History, ecology, and conservation of the coast redwoods. Island Press, Washington, D.C. Olson, D. F., D. F. Roy, and G. A. Walters. 1990. Sequoia sempervirens (D. Don) Endl. Pages 541–​551 in R. M. Barnes and B. H. Konkala, editors. Silvics of North America. Volume 1. Conifers. U.S. Department of Agriculture Forest Service, Washington, D.C. Pike, L. H., R. A. Rydell, and W. C. Denison. 1977. 400-year-old Douglas-fir tree and its epiphytes: Biomass, surface-area, and their distributions. Canadian Journal of Forest Research/Revue Canadienne de Recherche Forestiere 7:680–​699. Pillers, M. D., and J. D. Stuart. 1993. Leaf-litter accretion and decomposition in interior and coastal old-growth redwood stands. Canadian Journal of Forest Research 23:552–557. Powles, S. B., and O. Bjorkman. 1980. Leaf movement in the shade species Oxalis oregana. II. Role of protection against injury by intense light. Carnegie Institution Year Book 79:63–​66. Ramage, B. S., K. L. O’Hara, and B. T. Caldwell. 2010. The role of fire in the competitive dynamics of coast redwood forests. Ecosphere 1(6):20. Santiago, L. S. and T. E. Dawson. 2013. Light use efficiency of California redwood forest understory plants along a moisture gradient. Oecologia 174:351–363. Sawyer, J. O. 2006. Northwest California: A natural history. University of California Press, Berkeley, California. Schrepfer, S. R. 1983. The fight to save the redwoods. University of Wisconsin Press, Madison, Wisconsin. Schwartz, R. E., A. Gershunov, S. F. Iacobellis, and D. R. Cayan. 2014. North American West Coast summer low cloudiness: Broadscale variability associated with sea surface temperature. Geophysical Research Letters 41:3307–​3314. Sillett, S. C., and R. Van Pelt. 2007. Trunk reiteration promotes epiphytes and water storage in an old-growth redwood forest canopy. Ecological Monographs 77:335–​359. Simonin, K. S., L. S. Santiago, and T. E. Dawson. 2009. Fog interception by Sequoia sempervirens (D. Don) crowns decouples physiology from soil water deficit. Plant, Cell, and Environment 32:882–​892. Sloan, J. P., and K. N. Boe. 2008. Sequoia sempervirens (Lamb. ex D. Don) Endl. Pages 1034–​1036 in F. T. Bonner and R. P. Karrfalt, editors. Woody Plant Seed Manual, Agriculture Handbook 727. U.S. Department of Agriculture Forest Service, Washington, D.C. Snyder, J. A. 1992. The ecology of Sequoia sempervirens. San Jose State University, San Jose, California. Stone, E. C. and R. B. Vasey. 1968. Preservation of coast redwood on alluvial flats: Because man has altered environment, active management is now required. Science 159:157–161. Stuart, J. D., and S. L. Stephens. 2006. North coast bioregion. Pages 147–​169 in N. Sugihara, J. van Wagtendonk, J. Fites-Kaufmann, K. Shaffer, and A. Thode, editors. Fire in California’s Ecosystems. University of California Press, Berkeley, California. Tolle, G., et al. 2005. A macroscope in the redwoods. Third International Conference on Embedded Networked Sensor Systems. San Diego, California. U.S. Geological Survey (USGS). 1999. Digital representations of tree species range maps from “Atlas of United States Trees” by Elbert L. Little, Jr. (and other publications). Van Pelt, R. 2001. Forest giants of the Pacific Coast. University of Washington Press, Seattle, Washington. Waring, K. M., and K. L. O’Hara. 2008. Redwood/tanoak stand development and response to tanoak mortality caused by Phytophthora ramorum. Forest Ecology and Management 255:2650–​2658. Waring, R. H., and J. F. Franklin. 1979. Evergreen coniferous forests of the Pacific Northwest. Science 204:1380–​1386. Whittaker, R. H. 1960. Vegetation of the Siskiyou Mountains, Oregon and California. Ecological Monographs 30:279–​338. Williams, C. B., and S. C. Sillett. 2007. Epiphyte communities on redwood (Sequoia sempervirens) in northwestern California. Bryologist 110:420–​452. Yaryan, W., D. Verardo, and J. Verardo. 2000. The Sempervirens Story: A Century of Preserving California's Ancient Redwood Forest 1900–2000. Los Altos, California, Sempervirens Fund. Zinke, P. J., A. Stangenberger, and W. Colwell. 1979. Fertility of the forest. California Agriculture 33:10–​11.

T WENT Y-SE VEN

Montane Forests M ALCOLM NORTH, BR ANDON COLLINS, HUGH SAFFORD, and NATHAN L . STEPHENSON

Introduction California’s montane forests include some of the most productive and diverse temperate ecosystems in the world. They contain the largest single-stem tree (the 1,487 cubic meter General Sherman giant sequoia [Sequoiadendron giganteum]) (Van Pelt 2001) and the highest conifer diversity (thirty-plus species in the Klamath-Siskiyou mountain range) (Sawyer 2006) in the world (Table 27.1). Although these forests share some attributes with the Pacific Northwest (i.e., long-lived large trees and some common wildlife species) (North et al. 2004) and Southwest (i.e., historical forests dominated by pine and shaped by frequent fire), their combination of high productivity, strong seasonal drought, and fire dependence distinguish them ecologically from montane forests in these adjacent areas. The distribution of different forest types is strongly influenced by temperature and precipitation gradients associated with elevation and inland distance from the Pacific Ocean. Historically the forests have been logged, but interestingly for the nation’s most populous state, some large areas of

montane forest—​especially in the central and southern Sierra Nevada—​remain only lightly affected by human resource use. Management practices in these montane forests have often been controversial, and fire suppression has significantly altered forest conditions such that fires escaping containment are often large and produce extensive areas burned at high severity. Climate change, California’s increasing population, and projected increases in wildfire pose challenges that will require collaborative and inventive future management. In this chapter we focus on montane forest ecosystems in California’s Sierra Nevada, Klamath, Cascade, Coastal, Traverse, and Peninsular Ranges. These forests often border, at lower elevations, warmer, drier ecosystems that include oaksavanna and chaparral (see Chapters 24, “Chaparral,” and 25, “Oak Woodlands”). At upper elevations, montane forests often border colder red fir and subalpine forests characterized by deeper, more persistent snowpacks (see Chapter 28, “Subalpine Forests”). 553

TA B L E 27.1 Characteristics of mature trees for major tree species in California’s montane forests at the turn of nineteenth century

Species

Elevation (m)

Height (m)

Diameter at breast height (DBH) (cm)

Mature age (years)

White fir

800–2,300 N: 1,500–2,400 S

53–58 (66)

110–170 (223)

200–300 (372)

Red fir

1,400–2,400 N; 1,700–2,700 S

51–56 (77)

100–180 (295)

250–400 (665)

Incense cedar

600–1,600 N; 800–2,100 S

24–31 (70)

150–210 (456)

250–600 (unk)

Jeffrey pine

1,500–2,400 N; 1,700–2,800 S

34–49 (61)

90–150 (250)

120–350 (813)

Sugar pine

1,000–2,000 N; 1,400–2,700 S

55–61 (82)

120–180 (352)

100–400 (unk)

Ponderosa pine

300–1,800 N; 1,200–2,100 S

45–55 (70)B

90–120 (277)C

100–350 (907)A

Douglas-fir

300–2,100 N; 600–2,100 S

46–53 (92)A

120–210 (485)A

120–400 (1,350)

Black oak

900–1,500 N; 1,400–2,100 S

28–34 (38)C

130–160 (273)C

120–400 (unk)

Giant sequoia

900–2,000 N; 1,600–2,700 S

50–75 (94)

400–600 (1,228)

400–1,100 (3,266)

Source: According to Sudworth 1900. Note: In height and DBH columns, values in parentheses are for the largest dimension on record (in van Pelt 2001) and the oldest on record (in The Gymnosperm Database), www.conifers.org. A. Record holder in Washington State B. Record holder in Idaho C. Record holder in Oregon Unk = unknown

Physiographic Setting Climate California’s Mediterranean climate, in which 85% of annual precipitation occurs between November and May, significantly influences montane forest composition, distribution, and ecosystem functions (Minnich 2007). Unlike montane weather in much of the Rockies and the Southwest, summer thunderstorms and significant rain events are infrequent and often highly localized in California. During the growing season (generally March through July), plants rely on soil moisture stocks from winter precipitation and melting snow (Major 1977). The distribution of montane forest is influenced by site water availability and evaporative demand and roughly corresponds to an elevation zone with a mean annual temperature of 7°C to 12°C and total precipitation of 800 to 1,800 mm yr-1 (Stephenson 1998, Goulden et al. 2012). In general, the elevation of this zone shifts upward from north to south and coast to inland. For example, mixed-conifer forests in the Klamath Mountains occur between about 1,000 to

Photo on previous page: Mixed-conifer forest with a restored fire regime, containing horizontal and vertical structural diversity, in the Illilouette Basin of Yosemite National Park. Photo: Marc Meyer. 554  Ecosystems

2,100 meters in elevation, compared to 1,400 to 2,400 meters in the southern Sierra Nevada (Fites-Kaufman et al. 2007). Temperatures decrease by approximately 4oC to 6oC km-1 of elevation gain (Major 1977). Precipitation rises with elevation up to a point then gradually decreases. For example, maximum precipitation in the central Sierra Nevada occurs around 1,900 meters (Armstrong and Stidd 1967, Major 1977). The percentage of annual precipitation that occurs as snow ranges from approximately 25% to 65% in lower to upper montane forests. Montane forests typically transition to red fir (Abies magnifica)–​dominated forest types at elevations where winter and spring temperatures ensure that most of the precipitation occurs as snow (Barbour et al. 1991). A large increase in snowpack accompanies the transition from ponderosa pine (Pinus ponderosa)–​dominated forest through mixed conifer to red fir. For example, in the Pit River basin in northern California, April 1 snow depths below 1,500 meters are minimal (13 centimeters at Burney Springs in ponderosa pine) but reach over 200 centimeters at elevations over 1,950 meters in the red fir zone. In the American River basin of the central Sierra Nevada, similar patterns occur, with 37 centimeters average April 1 snow depth at 1,600 meters eleva-

tion (pine-dominated, mixed conifer) and 215 centimeters at 2,270 meters (red fir) (WRCC). In red fir forests the deeper snowpack and cooler temperatures substantially reduce the length and magnitude of the summer drought (Royce and Barbour 2001). On an annual basis, California experiences one of the most variable precipitation regimes in the United States. In montane forests this means snowpack depth (and available water during the growing season) can vary by more than an order of magnitude between years (Mote 2006). The El Niño/Southern Oscillation (ENSO) often drives this interannual variability, but the area affected in any annual event varies across the range of montane forests (Minnich 2007). An El Niño year may produce a deep snowpack in the southern California mountains while northern California experiences normal or below normal precipitation. Patterns vary year-to-year depending on the strength of the ENSO event. California Department of Water Resources splits its Sierra Nevada snowpack assessment into northern, central, and southern zones because winter storms tracking from the Pacific Northwest and southern California might only impact the northern and southern half, respectively, of the state (see Chapter 2, “Climate”). Long-term averages suggest Yosemite National Park might roughly be a transition zone, as montane forest precipitation significantly decreases south of the park coincident with the southern range limit of Douglas-fir (Pseudotsuga menziesii) in the Sierra Nevada (Burns and Honkala 1990). Precipitation patterns within the wet season are also highly variable. A few large winter storms usually account for a third to a half of annual precipitation and can occur over a seasonal total of only five to ten wet days per year. So-called atmospheric rivers (ARs) generate 20% to 50% of the state’s precipitation totals (Dettinger et al. 2011). ARs are narrow bands (typically less than 200 kilometers wide) of concentrated water vapor that develop over the oceans and direct large amounts of moisture at continental areas (see Chapter 2, “Climate”). Due to the influence of periodic ARs developing over the tropical Pacific, California experiences more extreme precipitation events than any other part of the U.S., including the hurricane-affected Gulf Coast (Dettinger et al. 2011). ARs in montane forests, such as the New Year’s Day storm of 1997, can cause widespread flooding and landslides, reshaping riparian corridors and forest stands on unstable slopes.

Soils Montane forest soils are generally relatively young and weakly developed because of the recent glaciation of most California mountain ranges (generally higher than 1,400 meters in elevation) during the Late Wisconsin glacial episode (thirty thousand to ten thousand years ago) (Atwater et al. 1986; see Chapter 4,“Geomorphology and Soils”). Soils over granitic parent material, which predominate in montane ecosystems south of the Merced River (O’Green et al. 2007), often are highly porous and well-drained because they contain a substantial fraction of decomposed granite. California’s mountains also contain older, metamorphosed rocks invaded by batholithic magma and young volcanic and sedimentary postbatholithic rocks (Harden 2004). Most lower montane soils are Alfisols and more highly leached Ultisols and are moderate to strongly acidic. Organic horizons may be deep (often depending on time since last fire) due to low decomposition rates resulting from conifer’s poor litter quality and

from minimal topsoil mixing by soil fauna (O’Green et al. 2007). Microbial activity and organic matter decomposition are also limited by dry conditions during California’s annual summer drought. For example, the thick litter layers often found around the base of mid-elevation trees result in part from reduced decomposition rates from rapid drying produced by accelerated snowmelt from solar heating of the tree bole (Johnson et al. 2009). Soil depth varies considerably, strongly affecting water-holding capacity and in turn forest productivity. Maximum clay content may range from 30% to 55% in the lower montane zone, but a significant reduction occurs in soil development above 1,500 to 1,800 meters in elevation, with clay content often dropping to less than 15% (O’Green et al. 2007). Inceptisols and Andisols dominate these higher-elevation soils. Generally, such soils are shallow and rich in organic matter with large, unconsolidated fragments and limited soil moisture storage. One study examining soil change along an elevation gradient (200–​2 ,900 meters) found decreasing pH (about two units) and base saturation (90% to 10%) and increasing organic carbon with increasing elevation (Dahlgren et al. 1997). They also found maximum chemical weathering at mid-elevations (ponderosa and mixed-conifer forest types). Aspect and slope position also influence soil development and processes, with more weathering and deeper, richer soils on mesic, northerly aspects and lower-slope positions compared to xeric, southfacing, and upper-slope conditions. After water, nitrogen is typically the most limiting plant growth resource in temperate zones (Vitousek and Howarth 1991). In California’s montane forests mineral soil holds most (65–​9 0%) of the nitrogen (Johnson et al. 2008). Nitrogen loss from fire volatilization increases with fire intensity but is often rapidly replenished by common actinorhizal shrubs such as Ceanothus spp. and bear clover (Chamaebatia foliolosa) that are associated with nitrogen-fixing bacteria (Frankia spp.) (Oakley et al. 2003, 2004; Stein et al. 2010). Ultramafic soils are not widespread in Sierra Nevada montane conifer ecosystems except in the Feather River drainages, but where present they significantly limit plant productivity and species composition as they lack most macronutrients and contain various heavy metals. The Klamath-Siskiyou Mountains have extensive ultramafic soil areas that support many rare plants and plant communities unique to this region of high diversity (Alexander et al. 2007).

Montane Forest Types Several classification schemes exist for montane forest types in California (Critchfield 1971, Davis et al. 1995, Holland and Keil 1995, Barbour et al. 2007, Sawyer et al. 2009). Although each slightly differs in the number and types of forest ecosystems, they generally concur, particularly for the most widely distributed forest types discussed in this chapter. More detailed vegetation information using finer forest-type classifications are available in Barbour et al. (2007) and Sawyer et al. (2009). We generally have used the California Wildlife Habitat Relationships System (Meyer and Laudenslayer 1988) (Figure 27.1), which builds upon the U.S. Forest Service California Vegetation (CALVEG 2013) classification based on existing vegetation (rather than potential natural vegetation). All of the forest types discussed in this chapter could be viewed as related to the mixed-conifer group because they are delineated by changes in Mon tane Forests   555

FIGURE 2 7.1 Distribution of montane forest types in California based on California wildlife habitat and CALVEG classifications. Giant sequoia groves not shown. Data from Cal Fire, Fire Resource and Assessment Program (FRAP). Map: Ross Gerrard, USFS PSW Research Station, and Parker Welch, Center for Integrated Spatial Research (CISR).

environmental or edaphic conditions that allow one or several typical mixed-conifer species to become dominant.

Mixed Conifer Mixed conifer (2,484,012 hectares) is one of the most common montane forest ecosystems in California and has the highest diversity among them of vertebrate species (Meyer and Laudenslayer 1988) (Figure 27.2a). Major tree species include ponderosa pine, Jeffrey pine (Pinus jeffreyii), sugar pine (P. lambertiana), white fir (Abies concolor), incense-cedar (Calocedrus decurrens), Douglas-fir, and black oak (Quercus kelloggii). Red fir, lodgepole pine (Pinus contorta murrayana), and western white pine (Pinus monticola)—​all upper montane species—​intermix with mixed conifer at higher elevations and in cold air drainages. Within mixed conifer, species composition can change over small distances, often in response to water availability (i.e., generally transitioning from fir to pine with increasing dryness) and microclimate (i.e., cold air drainages that retain snowpack are dominated by lodgepole pine, and white and red fir). Differences with aspect in temperature and insolation influence the elevations at which different forest types occur (Figure 27.3). Considering the overwhelming importance of water availability and fire—​historic mean fire return intervals (HFRI) averaged eleven to sixteen years (Van de Water and Safford 2011)—​ to mixed-conifer ecosystems, perhaps the most important distinction within the mixture of species concerns those that are highly tolerant of fire and drought but intolerant of shade (black oak and the yellow pines) and those that are less tolerant of fire and drought but grow relatively well in low-light conditions (white fir, incense-cedar, Douglas-fir) (Table 27.2). Historically, frequent fire kept forests generally open and exposed bare mineral soil, although conditions varied with topography and fire history (Collins, Lydersen et al. 2015, Stephens, Lydersen et al. 2015). These conditions favored pines, which could comprise up to 40–​65% of the trees (McKelvey and Johnston 1992, North et al. 2007, Lydersen and North 2012). Fire suppression has significantly increased stem densities and canopy cover and reduced understory light, resulting in heavy dominance by fir and incense-cedar and little shrub cover in many mixed-conifer forests today (North, Oakley et al. 2005; Collins et al. 2011; Dolanc et al. 2013; Knapp et al. 2013).

Klamath Mixed Conifer Mixed-conifer forests in the Klamath Mountains (461,666 hectares) contain many species found in montane ranges in both California and the Pacific Northwest, contributing to high diversity and unique community assemblages (Sawyer 2006) (see Figure 27.2b). The diverse flora has developed over a long period from many biogeographic sources (Whittaker 1960, Briles et al. 2005) including migration from other regions (Stebbins and Major 1965), relictual species such as the recently identified Shasta snow-wreath (Neviusia cliftonii) (Lindstrand and Nelson 2006), and newly evolved taxa (Smith and Sawyer 1985). This diversity does not appear to result from the Klamath’s fire regimes (HFRI seven to thirteen years) (Fry and Stephens 2006), which are similar to those in the Sierra Nevada, but instead is probably due to its ecotonal location straddling an area where different climate, geologic, and edaphic zones collide. Ultramafic, mafic, granitic, sedi-

mentary, and metamorphic substrates are common (Waring 1969, Kruckeberg 1984), and temperature and precipitation gradients are very steep between the wet, cool western slopes and dry, hot interior areas bordering the Central Valley. Pleistocene glaciation in the Klamath Mountains was primarily confined to higher elevations (Sawyer 2007), and unlike most of the Sierra Nevada, modern streamflow is more driven by rainfall than by snowpack (Miller et al. 2003). Populations of some conifers in this range are geographically quite distant from their distributions in other mountain ranges, and some endemic conifers occur here as well (Sawyer 2007). Examples of the former include foxtail pine (Pinus balfouriana; shared with the southern Sierra Nevada), Engelmann spruce (Picea engelmannii), and subalpine fir (Abies lasiocarpa—​both at their southern range limit in the Klamaths); the latter is exemplified by Brewer spruce (Picea breweriana). In addition to thirty conifer species, Sawyer (2007) lists nineteen common hardwood tree species and more than seventy species of shrubs. The rugged topography in these mountains contributes to different fire regimes and forest types associated with aspect, slope position, and elevation (Taylor and Skinner 2003). Because the mountains are discontinuous and highly dissected, forest types tend to occur in patches rather than continuous belts. The rugged topography and lack of a dominant orientation provides many edaphic, microclimate, and disturbance regime differences over fine scales (Skinner et al. 2006). Whereas mixed conifer in the Sierra Nevada may have three to five tree species at any one site, Klamath mixed conifer typically has five to seven or more species growing together. A few sites in the Klamath Mountains apparently support the highest local diversity of conifers in the world. For example, the Sugar Creek drainage in the headwaters of the Scott River contains seventeen conifer species within an area of less than 2.6 square kilometers (Cheng 2004).

White Fir Forests in which white fir (379,372 hectares) makes up more 60% of the relative canopy cover are often typed white fir (Sawyer et al. 2009), although other species such as incensecedar, dogwood (Cornus nuttallii), ponderosa and Jeffrey pine, and Douglas-fir may also be present (see Figure 27.2c). Although white fir is more widely distributed than any other fir in California, intermixing with many species, it becomes dominant in mesic areas that have a longer fire return interval (greater than forty-five years) (Van de Water and Safford 2011), where regenerating trees can grow large enough to survive low- to moderate-intensity fire. White fir forests may be intermixed with or adjacent to mixed conifer and usually indicate an abiotic shift toward cooler, wetter conditions. However, white fir is also found in drier interior ranges including the Warner Mountains (1,550–​1,850 meters) in northeastern California and the Clark, Kingston, and New York Ranges (2,300–​2,900 meters) (Paysen et al. 1980) in the Mojave area. The dominance of a shade-tolerant species creates high canopy cover and multilayer stands with less fine-scale variability in microclimate, habitat, and understory conditions. Few openings and low understory light conditions reduce shrub and herbaceous cover; reduced fire frequency can produce thick litter and duff layers, higher fuel loads, and more snags and coarse woody debris than are common in mixed conifer. Heart rots are common, producing snags for cavity-

Mon tane Forests   557

A

B

C

D

E

F FIGURE 2 7.2 Forest types. A Mixed conifer (Photo: Malcolm North) B Klamath mixed conifer (Photo: Carl Skinner) C White fir (Photo: Malcolm North) D Montane hardwood-conifer (Photo: Carl Skinner) E Giant sequoia (Photo: Nate Stephenson) F Eastside pine (Photo: Malcolm North)

High

Water supply

(Deep soils, no rain shadow)

Low

(Shallow soils or in rain shadow) Alpine

Elevation (meters)

3500

3000

1500

t Wes

ine (

lp Suba

)

pine

e Subalpin

ine

gepole p

Jeffrey

pine

nifer

ixed co

ine-m rosa p

oak Canyon

ck oak

Bla

Low (north-facing)

pine)

Red fir

er d conif - mixe ia r o fi u e q it e Wh nt s ing Gia (includ ed conifer) mix

Ponde

(Foxtail

and Lod

Red fir

2500

2000

hite ern w

High (south-facing)

Evaporative demand

Low (north-facing)

High (south-facing)

Evaporative demand

FIGURE 2 7.3 The approximate distribution of forest types in the southern Sierra Nevada relative to elevation (y-axis), evaporative demand (x-axis), and water supply (compare left and right panels). Source: Fites-Kaufmann et al. 2007.

dependent wildlife and large logs that can be used as runway structures for small mammals. White fir is the preferred tree species for several insect-gleaning birds, including yellow-rumped warblers (Setophaga coronate), western tanagers (Piranga ludoviciana), mountain chickadee (Poecile gambeli), chestnut-backed chickadee (Parus rufescens), golden-crowned kinglet (Regulus satrapa), and black-headed grosbeak (Pheucticus melanocephalus) (Airola and Barrett 1985).

Montane Hardwood-Conifer There are many variations of montane hardwood-conifer (2,484,012 hectares), sometimes called mixed evergreen, which is characterized by composition including at least one-third hardwoods and one-third conifer (Anderson et al. 1976) (see Figure 27.2d). Hardwoods are usually dominated by species from the oak family, such as interior live oak (Quercus wislizeni), canyon live oak (Q. chrysolepis), tanoak (Notholithocarpus densiflorus), California black oak, Oregon white oak (Q. garryana), and golden chinquapin (Chrysolepis chrysophylla). Other hardwood associates include Pacific madrone (Arbutus menziesii) and big-leaf maple (Acer macrophyllum), while common conifers are Douglas-fir, ponderosa pine, white fir, incense-cedar and sugar pine, and in central coastal and southern California, bigcone Douglas-fir (Pseudotsuga macrocarpa) and coulter pine (Pinus coulteri) (Meyer and Laudenslayer 1988). The overstory canopy (30–​6 0 meters) is often dominated by conifers, especially where fire has been lacking for many years. Hardwoods comprise a lower layer (10–​ 3 0 meters) above a shrub layer that can be sparse (in stands with high total canopy cover) to impenetrable (following fire). Although

present in all of California’s mountain ranges, the montane hardwood-conifer type is most extensive in the Klamath and North Coast Ranges and the northern Sierra Nevada and is often transitional between dense coniferous forests and mixed chaparral, open woodlands, or savannas. The mix of species can be highly variable, but some general patterns are apparent. Tanoak is a significant component in more mesic areas in the northern and western montane areas of California, and Douglas-fir is a local dominant in the same areas. Both species are absent in the southern Sierra Nevada and south Coast Ranges. Drier inland regions tend to have more black and canyon live oak, with the latter particularly found on steep slopes with thinner soils where sparse fuel accumulation reduces fire frequency and intensity. A recently introduced pathogen, sudden oak death (Phytophthora ramorum), has spread throughout many coastal montane hardwoodconifer forests within the moist fog belt, whose conditions favor the disease’s dispersal. Tanoak experiences almost 100% mortality to sudden oak death. Although the pathogen rarely kills California bay (Umbellularia californica), research has identified it as the best predictor of the disease’s presence (Rizzo and Garbelotto 2003). Many of these hardwood species benefit from severe fire because they resprout. Warming temperatures, increasing precipitation, and increasing nutrient inputs from air pollution enhance their competitiveness with conifers. Lenihan et al. (2008) projected a notable increase in the area of hardwood forest by the end of the twenty-first century as a result of climate change. A recent study comparing forest structure between the 1930s and 2000s found that stem densities in Sierran montane hardwood forest have increased more than in any other forest type and that the proportion of plots dominated by montane hardwood species increased Mon tane Forests   559

TA B L E 27. 2 Comparative ecological tolerances of common tree species in California’s montane forests Species arranged from low tolerance (top) to high tolerance (bottom)

Shade

Frost

TemperatureA

Drought

FireB

Black oak/western juniper

Madrone

Lodgepole pine

Red fir

Lodgepole pine

Ponderosa pine/giant sequoia

Douglas-fir

Red fir

White fir

Sugar pine/white fir

Lodgepole pine

White fir

Jeffrey pine

Sugar pine/ giant sequoia

Incense cedar/ Douglas-fir

Sugar pine

Sugar pine/giant sequoia

White fir/giant sequoia

Douglas-fir

Jeffrey pine/ponderosa pine/giant sequoia

Incense cedar

Incense cedar

Douglas-fir/sugar pine/ incense cedar

Lodgepole pine/ incense cedar/ madrone

Douglas-fir

Ponderosa pine/ Jeffrey pine/red fir

Ponderosa pine/black oak/ madrone

Ponderosa pine

Red fir

Lodgepole pine

White fir

Jeffrey pine Black oak

Source: Data from Minore 1979, Burns and Honkala 1990, and the Fire Effects Information System (FEIS). A. Least heat tolerant/most cold tolerant on top. B. Fire tolerance of mature trees. Fir and Douglas-fir seedlings and saplings are less tolerant of fire than yellow pine and sugar pine.

by almost 100% between the two time periods (Dolanc et al. 2013).

regeneration has been restored (Stephenson 1996, Stephenson 1999, York et al. 2013).

Giant Sequoia

Ponderosa, Jeffrey, and “Eastside” Pine

Giant sequoia occurs naturally in roughly seventy relatively small, scattered groves (total area of 14,600 hectares) along the western slope of the Sierra Nevada, mostly south of the Kings River (Stephenson 1996, Fites-Kaufman et al. 2007) (see Figure 27.2e). The eight groves north of the Kings River span elevations of 1,370 to 2,000 meters, while those to the south are mostly found between 1,700 and 2,250 meters in elevation. Within these groves, giant sequoia usually dominates basal area, but by stem frequency and composition the groves would be considered mixed conifer. Grove locations are usually characterized by deep, well-drained soils with relatively high water availability. Although the trees can grow rapidly, typically to 400–​8 00 centimeters in diameter and 65–​8 0 meters tall, they are also long-lived, reaching ages of one thousand years or older. More than any other tree species in the Sierra Nevada, giant sequoia is a pioneer species requiring disturbance for successful regeneration (Stephenson 1994). In the past, frequent, moderate-intensity fires burned through sequoia groves, creating occasional gaps in the forest canopy in locations where these fires burned at high severity. These canopy gaps, with their greatly reduced competition for light and water, are the sites of virtually all successful sequoia regeneration (Stephenson 1994, York et al. 2011, Meyer and Safford 2011). Following Euro-American settlement, more than a century of fire exclusion led to a nearly complete failure of sequoia regeneration. Where fire has subsequently been reintroduced, sequoia

Sometimes collectively called the “yellow pines,” ponderosa and Jeffrey pine are closely related (both in the subgenus Pinus, section Pinus, subsection Ponderosae) and occasionally hybridize (Baldwin et al. 2012). Ponderosa pine, one of the most widely distributed pine species in North America (327,778 hectares), is found throughout the mountainous regions of the western U.S., whereas Jeffrey pine is primarily a California tree, with a few occurrences in westernmost Nevada, southwestern Oregon, and northern Baja California, Mexico (see Figure 27.2f). Of the two species, Jeffrey pine is more stress-tolerant and replaces ponderosa pine at higher elevations, on poorer soils, and in colder and/or drier climates (Haller 1959, Stephenson 1998, Barbour and Minnich 2000). Ponderosa pine–​dominated forests can occur from approximately 300 to 1,800 meters and 1,200 to 2,100 meters in northern and southern California, respectively (FitesKaufman et al. 2007). Jeffrey pine–​dominated forests occur mostly between 1,500 and 2,400 meters and 1,700 and 2,800 meters in northern and southern California (with the highest elevations usually on the east side of the Sierra Nevada), respectively (Fites-Kaufman et al. 2007, Barbour and Minnich 2000). Both yellow pine species also occur in other forest types including mixed conifer, where they were dominant in many places before logging and fire suppression. A large area of the northern Sierra Nevada east of the crest supports a mixed yellow pine forest, sometimes called “eastside pine,” with co-dominance by ponderosa and Jeffrey pine

560  Ecosystems

(782,526 hectares of Jeffrey and eastside pine area combined). Like white fir forests, forests dominated by ponderosa and Jeffrey pine are closely intermingled with mixed conifer, in this case indicating a shift toward drier, warmer (ponderosa) or drier, colder (Jeffrey) site conditions. Historically these forests had very frequent fires that supported low-density openstand conditions characterized by shrub patches, sparse litter cover, and relatively high diversity of herbs and grasses.

Forest Structure and Function In the nineteenth century John Muir wrote, “These  forests were so open, early travelers could ride a horse or even pull wagons through [them].” An early timber survey in the northern Sierra Nevada noted the same conditions and lamented that fire kept the forest at only 30% of its potential lumber stocking (Lieberg 1902). Historically these low-density, large tree–​dominated forests had an almost flat diameter distribution (an equal abundance of all tree sizes) in contrast to the reverse J-shaped distribution (tree abundance rapidly decreases in larger-size classes) many early foresters were familiar with from forests with more infrequent disturbance regimes (North et al. 2007). Based on detailed timber surveys conducted by the Forest Service in the central Sierra Nevada in 1911, tree densities ranged from 40 to 80 trees ha-1 and estimated canopy cover was 17–​24% (Collins et al. 2011). These open conditions were maintained by frequent surface fires that consumed surface fuels and small-diameter trees, providing a pulse of nutrients to the soil, creating patches of bare mineral soil for seed establishment, and reducing competition for soil moisture (Gray et al. 2005, Zald et al. 2008). Without fire, forest structure becomes more homogeneous, and some ecosystem functions “stall” (Ma et al. 2004, North and Rosenthal 2006) (Figure 27.4). Frequent fire creates structural diversity at fine (stand) and coarse (landscape) scales associated with several ecosystem processes. The within-stand structure has been characterized as containing three main conditions: individual trees, clumps of trees, and openings or gaps (ICO) (Larson and Churchill 2012, Lydersen et al. 2013, Fry et al. 2014). Stand-level average canopy cover under frequent-fire conditions is typically low (20–​45%) compared to modern, fire-suppressed conditions (typically 55–​85%). However, within a stand, ICO conditions produce heterogeneity such that canopy closure (a pointlevel measure) (Jennings et al. 1999, North and Stine 2012) is highly variable, providing a scattering of dense areas for wildlife cover. Several studies suggest this fine-scale heterogeneity affects ecosystem conditions and functions, producing a wide range of microclimates (Rambo and North 2009, Ma et al. 2010), a diversity of understory plants (Wayman and North 2007) and soil invertebrates (Marra and Edmonds 2005), variation in soil respiration (Concilio et al. 2005, Ryu et al. 2009), and limits to pest and pathogen spread (Maloney et al. 2008). In addition, modeling efforts that have compared fire-suppressed forest conditions with two different fuel reduction treatments found higher avian richness in forests treated to create variable canopy closure and increase structural heterogeneity (White et al. 2013a, b). At a larger scale, models suggest that fire created a mosaic of different forest seral conditions that diversified landscape structure (Kane et al. 2014). For example, modeling by the LANDFIRE program (Rollins and Frame 2006, Rollins 2009) predicts that under an active fire regime, 10–​2 0% of yellow

and mixed-conifer forests in California would have been in early seral stages (herbs, shrubs, seedlings/saplings) with approximately 30–​40% in areas dominated by trees between 10 and 53 centimeters dbh (diameter at breast height) (5–​21”), and 40–​60% in areas dominated by larger trees (higher than 53 centimeters dbh). Furthermore, the models indicate that most of the landscape was under open forests of less than 50% canopy cover (“open” stages), especially in the yellow pine and drier mixed-conifer types (Rollins 2009). This is quite different than modern forest conditions where 85% of montane forests are dominated by 10–​53 centimeter trees and canopy cover averages greater than 65%. Water availability appears to be one of the strongest influences on ecosystem function. At fine scales, functions such as decomposition, nutrient cycling, and soil respiration vary strongly within forest stands by patch type (North and Chen 2005) and their different levels of available soil moisture (Erickson et al. 2005, North, Oakley et al. 2005, Concilio et al. 2006). At larger scales, forest greenness (Trujillo et al. 2012), CO2 uptake, and evapotranspiration (Goulden et al. 2012) are correlated with elevational differences in snowpack depth and total precipitation.

Fauna Montane forests in California support at least 355 vertebrate species (Verner and Boss 1980). The high species richness of montane forests probably stems in part from changing habitat conditions created by frequent fire and seral development. For example, a study of the avian community between a burned area and neighboring unburned forest in the Sierra Nevada found that over a third of species occurred only in the burned area (Bock and Lynch 1970). In addition, a study of breeding birds observed over a twenty-five-year period found that bird community guild structure shifted among species with different foraging strategies (i.e., foliage searching to bark gleaning) as forest succession progressed (Raphael et al. 1987). Concern has often focused on species that might be affected by modern changes in forest conditions that differ from their historical analogs or that have become increasingly rare (North and Manley 2012, Stephens et al. 2014). For instance, some songbirds are strongly associated with shrub patches (Burnett et al. 2009) now uncommon in the low-light understory of fire-suppressed forests (Knapp et al. 2013). The most widely known sensitive species, however, are associated with old forest conditions such as the northern (Strix occidentalis caurina) (Moen and Gutierrez 1997, North et al. 1999) and California spotted owls (S. o. occidentalis) (North et al. 2000, Lee and Irwin 2005), northern goshawk (Accipiter gentilis) (Morrison et al. 2009), fisher and marten (Martes pennanti and M. martes) (Zielinski et al. 2004a, b), southern red-backed vole (Clethrionomys gapperi) (Sullivan and Sullivan 2001), and northern flying squirrel (Glaucomys sabrinus) (Meyer et al. 2005, 2007; Meyer, North, and Kelt 2007) (Figure 27.5). The California spotted owl and fisher have been studied more extensively than other species because both are considered threatened. Guidelines for maintaining and improving their habitats strongly affect forest management on public lands (North, Stine et al. 2009; Roberts and North 2012). Both species are associated with large, old structures that contain high levels of canopy closure to use for nesting and resting. This has often resulted in minimal or no fuels removal in these areas, which in turn makes these sites prone to burning at Mon tane Forests   561

Historic pine-dominated mixed conifer

Low-intensity fires kill most fir

1

Pine-dominated forest type persists

2 Pine-dominated fire active + unthinned

1

Selective logging of mature pines

2

Fire-suppression allows fir to establish

3

Shift to fir-dominated forest type

Fir-dominated fire suppressed + selective harvest

FIGURE 2 7.4 Two generalized successional pathways for historic mixed-conifer forests (top). The left side shows how an active-fire regime maintains a resilient composition and structure dominated by a low density of large pine. The right side shows how past selective logging and fire suppression can lead to a high-density, white fir–​dominated forest stand. This right side is a common condition in many montane forests and can be very susceptible to high-intensity fire and drought mortality. Source: Earles et al. 2014.

high severity in the event of a wildfire and to subsequent loss of nesting and resting habitat (North et al. 2010). For foraging, however, both species use a variety of habitat conditions, possibly because they have broad prey bases that include several small mammal species associated with a range of forest and shrub conditions (Innes et al. 2007, Meyer et al. 2007). Some controversy has focused on the black-backed woodpecker (Picoides arcticus), a species associated with large, recently dead (four to eight years old) trees and often found foraging in “snag forests” produced by stand-replacing fires (Saab et al. 2007, Hanson and North 2008). The black-backed 562  Ecosystems

woodpecker might seem like an unlikely candidate for sensitive species status. With fire suppression, although the extent of wildfire has decreased, increased fire severity has kept the area of snag forests at levels consistent with estimates of historical conditions (though patch size has significantly increased) (Miller et al. 2009, Mallek et al. 2013). The concern with black-backed woodpecker habitat is not an areal decrease but a reduction in habitat suitability if many snags are removed by postfire salvage logging. In addition to reduced old forest conditions, some special habitat elements (e.g., “defect” trees) may have declined

A

C

B

D

E

FIGURE 2 7.5 Sensitive species that affect land management in California’s montane forests: (A) fisher resting on a large black oak limb, (B) a northern flying squirrel holding a truffle (its primary food source), (C) northern goshawk, (D) a California spotted owl, and (E) a black-backed woodpecker on a snag.

in abundance (Bouldin 1999). Large defect trees and snags are often rare in managed montane forests because they are removed for worker safety, and past stand “improvement” practices removed “defect” structures that did not contribute to wood production. Trees and snags selected by primary cavity nesters, woodpeckers, and nuthatches (Sitta spp.) could be particularly important because the cavities, once vacated, are used by other birds and mammals (Bull et al. 1997). Several studies have found that cavity availability can limit abundances of some of these species in managed forests (Carey et al. 1997, Carey 2002, Cockle et al. 2011, Wiebe 2011).

Ecosystem Characteristics Drought, Pests, and Pathogens Although montane forests are adapted to annual drought stress characteristic of Mediterranean climates, periods of multiple, consecutive dry years can have large impacts (e.g., see Guarin and Taylor 2005). For example, a massive die-off of conifer trees took place in the San Bernardino Mountains after the drought of the late 1990s and early 2000s. In the absence of frequent fire, increases in forest density result in Mon tane Forests   563

greater competition for scarce water (Innes 1992, Dolph et al. 1995). Potential increases in older tree mortality are a major concern because large trees are often more prone to droughtinduced mortality (Allen et al. 2010). Some studies have found higher than expected mortality rates in large trees (Smith et al. 2005, Lutz et al. 2009), suggesting that a “leave it alone” forest management approach that does not reduce stand density might actually contribute to the loss of old-growth trees. Drought itself is usually not the proximal cause of tree mortality, however, as drought-induced stress also leads to greater insect and disease susceptibility (Savage 1994, Logan et al. 2003, Fettig et al. 2007, Allen et al. 2010). In general, open stands with a mix of species have had more localized damage and mortality, while the scale and extent of mortality have been greater in dense, single-species stands and plantations (Stephens et al. 2012). Beetles are probably the greatest source of stress and mortality. Some beetle species are specialists focused primarily on one or two species, such as western (Dendroctonus brevicomis) and Jeffrey pine (D. jeffreyi) beetles primarily affecting ponderosa and Jeffrey pines, respectively. Mountain pine beetle (D. ponderosae) and California fivespined ips (Ips paraconfusus), however, are more generalist and affect most of the conifers in California’s montane ecosystems (Fettig 2012). Several pathogens also notably influence montane tree mortality. White pine blister rust (Cronartium ribicola) has been devastating to sugar pine since the disease entered northern California around 1930, and impacts to western white pine are also locally severe (Maloney et al. 2011). Pathologists and foresters have widely collected sugar pine seeds and conducted nursery experiments to identify blister rust–​resistant individuals to help regenerate the species (Kinloch 1992). Root rot (Heterobasidion spp.) disease is also widespread, particularly in fir-dominated forests. Some evidence suggests that forest thinning can accelerate spread of root rot because the disease’s windblown spores can colonize tree stumps (Rizzo and Slaughter 2001).

Fire Under presettlement conditions, most of California’s montane forests supported fire regimes characterized by frequent, predominantly low- to moderate-severity fires (Agee 1993, Sugihara et al. 2006, Barbour et al. 2007) (see Chapter 3, “Fire as an Ecosystem Process”). Historically these fire regimes were limited principally by the amount of available fuels rather than by fuel moisture during the summer drought. As elevation increases, the role of fuel moisture becomes more important, gradually supplanting fuel availability in red fir and higher-elevation forest types (Agee 1993, Miller and Urban 1999a, Sugihara et al. 2006, Van de Water and Safford 2011). Historically ignitions originated with Native American burning or lightning. Oral history suggests that many groups used fire to produce more open forest conditions favorable for foraging and hunting (Anderson et al. 1997). Whether these ignitions were concentrated in a few favored places or broadly used has been debated (Parker 2002; see Chapter 10, “Indigenous California”). In contrast, areas of heavy lightning activity are more easily identified. In a statewide analysis (van Wagtendonk and Cayan 2008), strikes increased with elevation up to 2,400 meters, had the highest monthly totals from June through the end of September, and occurred most between the hours of 1400 and 1900. Fire ignitions from lightning likely varied substantially from year to year. For the 564  Ecosystems

period between 1985 and 2000, van Wagtendonk and Cayan (2008) found a fivefold difference between the years with the highest and lowest number of strikes. Across the state, fire return intervals (FRIs) averaged eleven to sixteen years in yellow pine and mixed-conifer forests, with a mean minimum and maximum FRI of five and forty to eighty years, respectively (Van de Water and Safford 2011; North, Van de Water et al. 2009). Presettlement fire frequencies were higher in the drier, lower-elevation forest types (yellow pine and dry mixed conifer) and lower in moister and higher-elevation montane forests (Caprio and Swetnam 1995, Sugihara et al. 2006, Fites-Kaufman et al. 2007). Fire frequencies and patterns of fire severity were also influenced by local topographic variables. Several studies have documented longer FRIs and greater proportions of high severity on cooler, more mesic slopes (mostly north-facing), with the opposite pattern on warmer, more xeric slopes (mostly south-facing) (Kilgore and Taylor 1979, Fites-Kaufman et al. 1997, Taylor 2000, Beaty and Taylor 2001). Riparian areas also followed this pattern, with forests around smaller, headwater streams having a similar fire regime to adjacent uplands while larger streams (generally third order or greater) had longer fire return intervals (Van de Water and North 2010, 2011). In the absence of fire, many modern forests have unusually high fuel loads with much greater potential for high-severity, crown fires (Brown et al. 2008, Taylor et al. 2013). These conditions have shifted the fire regime from “fuel-limited” to “climate-limited” or “weather-limited” (Miller and Urban 1999b, Running 2006, Morgan et al. 2008, Collins et al. 2009, Miller et al. 2009, Steel, Safford et al. 2015). Adding to this trend is a policy of fire suppression on many forested lands, causing most wildfires to occur when they escape containment during extreme weather conditions (i.e., low humidity and high temperatures and wind speeds). In most montane forests the proportion and area of stand-replacing fire area and the sizes of stand-replacing patches are increasing (Miller et al. 2009, Miller and Safford 2012; however, see Miller, Skinner et al. 2012). These increases may be problematic because most of California’s montane trees species do not have direct mechanisms to regenerate following stand-replacing fire (e.g., serotiny, vegetative sprouting) (Goforth and Minnich 2008, Keeley 2012). This is particularly a concern in large stand-replacing patches, where the likelihood of wind-blown seed establishing is low (McDonald 1980). Conifer regeneration in stand-replacing patches can be highly variable. However, a recent study found that it was completely absent in nearly three-quarters of sampled high-severity patches, at least in the short term after fire (less than ten years) (Collins and Roller 2013).

Topography’s Influence Slope aspect, through its effects on insolation and hence the evaporative demand experienced by plants, has a relatively modest influence on montane forests and mostly affects the elevation at which particular forest types are found (Stephenson 1998, Fites-Kaufman et al. 2007, Lydersen and North 2012). For example, in the Sierra Nevada a given montane forest type can generally be found a few hundred meters higher on south-facing (sunward) slopes than on north-facing (shaded) slopes (see Figure 27.3). Water availability has more dramatic effects (Tague et al. 2009). For instance, firs are usually most abundant where water availability is high (such as on deep soils, with their high water-holding capaci-

N

E

W

Mature fir

Mature sugar pine

Mature Jeffery pine

Intermediate fir

Intermediate pine

Black oak

Small pine

Hardwood

Small fir S

Dry Relative soil moisture

Wet

FIGURE 2 7.6 Landscape schematic of variable mixed-conifer conditions produced by an active fire regime. Forest density and composition vary with topographic features such as slope, aspect, and slope position. Ridgetops, with drier soils and higher fire intensity, have lower stem density and a higher percentage of pine than more mesic riparian areas with lower-intensity fire. Midslope forest density and composition vary with aspect: density and fir abundance increase on more northern aspects (right side) and flatter slope angles. Illustration by Steve Oerding.

ties), whereas pines are most abundant where water availability is low (such as on shallow soils or in rain shadows) (Stephenson 1998; Fites-Kaufman et al. 2007; Meyer, North, Gray et al. 2007). Slope steepness and slope position (e.g., ridgetop, midslope, valley bottom) also affect the reception and retention of both meteoric waters and water flowing above, within, and beneath the soil. The influence of topography can be twofold, affecting both productivity and fire intensity (Figure 27.6) (Kane et al. 2015). Topographic locations that contain more mesic, productive sites (i.e., lower slope and riparian areas) were associated with greater densities of large, overstory trees, high total basal area and canopy cover, and an abundance of large snags and logs. This high-biomass forest structure existed in these topographic positions regardless of recent fire history. Outside of mesic sites and in forests that still have an active fire regime (i.e., no suppression), recent fire history was found to have the strongest influence on understory conditions (Lydersen and North 2012). Small tree density decreased and shrub cover increased with increased fire intensity and frequency, which in turn tended to occur on upper slope and ridgetop

locations. These findings suggest that topography, fire history, and their interaction produce the heterogeneity characteristic of montane forest landscapes (Taylor and Skinner 2003, Lydersen and North 2012).

Wind Overall, few historical accounts exist of large wind events in montane forests. In at least one study, the random direction of downed trees in old mixed conifer suggested that big wind events were not a significant driver of mortality (Innes et al. 2006). According to maps (Peterson 2000), California and neighboring states are subject to fewer major wind events like tornados and convective events (“downbursts”) than any other part of the contiguous United States. However, winds can have strong local effects. Very high winds can be common when winter storms arrive at the Sierra Nevada crest, but these elevations generally support subalpine forests. One recent event in fall 2011 in Devil’s Postpile National Monument in the upper San Joaquin River basin had winds exceedMon tane Forests   565

ing 145 kilometers per hour (Hilimire et al. 2012). Thousands of mature trees were downed—​mostly red fir, white fir, and lodgepole pine—​but areas of Jeffrey pine were also impacted. In some areas more than 70% of live trees were downed. Large trees and snags were more susceptible to uprooting than smaller ones, and effects were distributed fairly evenly across species. This size-dependent response to wind has a very different impact on forest structure than does fire, which preferentially kills smaller trees.

Forest Turnover Montane forests are more dynamic than forests found at higher elevations. For example, tree turnover rates (the average of tree recruitment and mortality rates) for old-growth Sierra Nevada forests are roughly three times greater in montane forests than in subalpine forests at treeline (Stephenson and van Mantgem 2005). The strong decline of forest turnover rates with increasing elevation may be related to parallel declines in forest productivity (Stephenson and van Mantgem 2005). Background tree mortality rates in montane forests generally can be higher in fire-suppressed forests than in contemporary forests with a more intact fire regime, possibly due to reduced competition among trees in burned stands (Ansley and Battles 1998, Maloney and Rizzo 2002, Stephens and Gill 2005). Similarly, modern plantation studies show much higher annual mortality in high-density than in low-density stands. In ponderosa pine, one study found annual mortality rates of between 0% and 0.8% in thinned stands of less than 332 trees per hectare, versus rates of 0.6% to 2.3% in stands of more than 2,450 trees per hectare (Zhang et al. 2006). Recent studies (van Mantgem and Stephenson 2007, van Mantgem et al. 2009) found that tree mortality rates in western U.S. forests have roughly doubled over the past few decades—​a n apparent consequence of warming temperatures.

Ecosystem Services The Millennium Ecosystem Assessment defines ecosystem services as the direct and indirect benefits people obtain from ecological systems (MEA 2005). California’s montane forests contribute to quality of life for millions of people, many living at some distance from the state’s mountain ranges. Ecosystem services are broadly categorized as provisioning (e.g., water, timber, fuels, food); regulating (e.g., carbon sequestration, erosion control, water quality); cultural (e.g., recreation, spiritual enrichment, educational opportunities); or supporting (biological diversity, nutrient cycling, etc.). All of these services are important, but we focus here on water, recreation, and carbon because of their particular relevance to California policies and economic development.

WATER

By one estimate, about 246,700,000,000 cubic meters (200 million acre feet [maf]) of precipitation falls annually on California, of which about 92,500,000,000 cubic meters (75 maf) is unimpaired runoff available for management and use (Energy Almanac 2014). About two-thirds of this annual runoff comes from one-fifth of California’s land area—​the mountains in the northern half of the state. A substantial portion 566  Ecosystems

of this water originates from precipitation in forested watersheds within the montane forest zone. Most of this water is eventually used by agriculture (41,900,000,000 cubic meters, or 34 maf). Furthermore, most montane rivers are highly engineered with multiple dams and impoundments that contribute to California’s greater than 13,725 gigawatt hours of hydroelectric power capacity (meeting about 8% of California’s electricity demand). A recent assessment of forested watersheds found the greatest threats to water quality and fisheries were concentrated in north coast watersheds. These threats stemmed from erosion following forest management activities, development, mass wasting, and high-severity wildfire (California Department of Forest and Fire Protection 2010). The high canopy cover in these forests caused by fire suppression might reduce water runoff because less snow reaches the ground and more is caught in the canopy, where it can sublimate directly back into the atmosphere (Golding and Swanson 1986, Essery et al. 2003). Climate change is expected to increase the percentage of precipitation that occurs as rain rather than snow (Hunsaker et al. 2012). This is expected to accelerate snowmelt and to challenge current reservoir capacity.

RECRE ATION

Montane forests are used heavily for a wide range of recreational activities. In 2010, Yosemite National Park alone drew four million visitors and provided more than $350 million in tourism revenue. Surveys on National Forest land in California found that the most popular activities were relaxing (52%), viewing natural features (52%), hiking and walking (47%), viewing wildlife (38%), and downhill skiing (36%), with an average of one to five trips per visitor annually (U.S. Department of Agriculture, Forest Service Region 2012). For the Sierra Nevada one study estimated an average rate of fifty million to sixty million annual visitor days for public forestlands alone (Duane 1996). Giant sequoia groves are probably among the most visited forest ecosystems in the world, and their appeal was instrumental in halting logging of the groves and establishing federal protection in the 1890s.

CARBON STOR AGE

Through the long-lived nature of many trees, global forests store twice as much carbon as Earth’s atmosphere. Global forest growth is a significant net carbon sink, adding 2.4±1.0 Pg C year-1 to biomass storage (Pan et al. 2011) and helping to offset anthropogenic emissions of CO2. Although developing countries often reduce their carbon stores when forestland is converted to other uses, California’s forest acreage has not changed appreciably over the past fifty years. Most montane forests in the state have been net carbon sinks in the last century due to regrowth from past harvesting and ingrowth from fire suppression (Hurteau and North 2009). California forests (all types and ownerships) are estimated to store 2.3 Pg of carbon (Fried and Zhou 2008). However, loss due to fire and conversion of forests due to development could offset, or even exceed, carbon stored from tree growth (Battles et al. 2013, Gonzalez, Battles et al. 2015). Thus the long-term stability of these carbon stores in forest is a key concern. There has been substantial debate about whether carbon loss through fuels treatment (mechanical thinning and/or

prescribed fire) in fire-prone forests is offset by a reduction in later carbon emissions if the treated stand is burned by wildfire (Hurteau et al. 2008, Hurteau and North 2009, Mitchell et al. 2009, Hurteau and North 2010, North and Hurteau 2011, Campbell et al. 2012, Carlson et al. 2012). In general, treating forests leads to net carbon loss because of the low current probability of wildfire burning the treated area, the modest reduction in wildfire combustion and carbon emissions, and the need to maintain fuels reduction through periodic, additional carbon removal (Campbell et al. 2012). The concept of carbon carrying capacity (Keith et al. 2009) could be particularly relevant to California’s montane forests. Carbon carrying capacity emphasizes the level of stable carbon storage that a forest can maintain over the long term. In the absence of disturbance, a forest can “pack on” more carbon as tree density and size increase (Hurteau and North 2009, Hurteau et al. 2013). Many montane forests are in this state today. This additional biomass, however, makes the forest prone to disturbances—​such as drought stress, pests, pathogens, and higher-severity wildfire—​that increase tree mortality. Mortality reduces carbon stocks as dead trees decompose and much of the carbon returns to the atmosphere through efflux. Carbon carrying capacity, therefore, is lower than the maximum storage potential of a forest but represents the biomass that can be maintained in the context of disturbance and mortality agents characteristic of a particular ecosystem. In California’s forests with historically frequent fire and drought events, carbon carrying capacity is the amount that a forest can store while maintaining low levels of mortality in response to periodic disturbances. In general, forests managed so that growth and carbon accumulation are concentrated in large trees will provide longer, more secure carbon storage than forests where growth is concentrated in a high density of small trees prone to pest, pathogen, and fire mortality (North, Hurteau et al. 2009; Earles et al. 2014). Recent research shows that large trees have remarkably high growth rates, giving them a more dynamic role in forest carbon storage than had been previously appreciated (Stephenson et al. 2014).

Human Impacts Although American Indians used trees for a variety of purposes, large-scale timber harvest did not begin until after widespread Euro-American settlement (circa 1850). Most logging before the 1900s was done to support mining operations. Timber was cut to build homes and commercial buildings, tunnels, mine and ore processing infrastructure, and railroad lines. It was also the fuel for heating, railroad engines, and other machines, and the various types of mills used for processing ore. In some areas a very valuable market in sugar pine shakes (for roofing or siding) also arose (McKelvey and Johnston 1992). The majority of timber harvests before and after 1900 occurred in yellow pine and mixed-conifer forest and often selected the largest, most valuable pine trees (Sudworth 1900, Leiberg 1902). Between the 1890s and 1920s, railroad lines were extended throughout the state’s lower- and middle-elevation forests to access timber resources beyond the reach of animal-drawn transport. After the Second World War, dramatically increasing wood demand from federal lands led the Forest Service to greatly expand their sale of timber. For example, harvest on the Eldorado National Forest averaged approximately 3.8 million board feet per year between 1902 and 1940 but increased

to 35.1 million board feet during the war, and to over 56 million board feet per year between the end of the war and 1959 (Beesley 1996). Harvest techniques were more industrial than before the war, and large areas of forest were clearcut. Since the 1960s, national legislation, regulations, changing economics, and environmental concerns have acted in concert to greatly reduce the amount of logging occurring on California public land, although private lands have made up some of the difference (see Chapter 36, “Forestry”). In the end, Barbour et al. (1993) estimated that half the original area of California’s mixed-conifer forest had been cut at least once in the past 150 years. The Sierra Nevada Ecosystem Project executive summary (SNEP 1996) has a succinct summary of the impacts of European settlement on montane ecosystems (Figure 27.4): “The primary impact of 150 years of forestry on middle-elevation conifer forests has been to simplify structure (including large trees, snags, woody debris of large diameter, canopies of multiple heights and closures, and complex spatial mosaics of vegetation), and presumably function, of these forests. By reducing the structural complexity of forests, by homogenizing landscape mosaics of woody debris, snags, canopy layers, tree age and size diversity, and forest gaps, species diversity has also been reduced and simplified.” Livestock grazing did not widely affect montane forests because, with the exception of scattered meadows, forage in these ecosystems is scarce, and most sheep and cattle concentrate their summer grazing in alpine meadows. Some areas of montane forests, however, are heavily affected by air pollution (see Chapter 7, “Atmospheric Chemistry”).

Current Management Strategies Fuels Treatment Fuels treatment is becoming the dominant forest management activity on public lands throughout the montane forest region of California. Mechanical thinning, prescribed fire, or combinations of both are most often used to reduce fuels (Safford et al. 2009; Safford, Stevens et al. 2012). Although controversy has persisted over the ecological effects of these treatments, a recent article synthesizing published studies found “few unintended consequences, since most ecosystem components (vegetation, soils, wildlife, bark beetles, carbon sequestration) exhibit very subtle effects or no measurable effects at all” to treatments (Stephens et al. 2012). Aside from controversy, limited budgets and other regulatory constraints have significantly reduced the pace and scale of fuels treatments (North et al. 2015). First-priority actions usually treat areas near homes in the wildland-urban interface (or WUI). With increasing home construction in these areas, more fuel treatment effort has been concentrated in these areas and correspondingly less in the larger forest matrix (Theobold and Romme 2007). One study of federal forestlands in the Sierra Nevada compared current levels of all fuels reduction treatments (including wildfire) to historical levels of fuel reduction from frequent fire. The study’s authors found that fewer than 20% of forests needing treatment were actually treated each year (North, Collins et al. 2012). They also calculated that at current rates more than 60% of the forest would never get treated, as maintenance of existing treatments would eventually subsume all of the fuels reduction effort. In mechanical fuels reduction, two measures are commonly used for implementing treatments: maximum tree diameter Mon tane Forests   567

FIGURE 2 7.7 A pine plantation forest managed to maximize tree growth rates. A single tree species is planted at regular spacing, producing a simplified stand structure. Photo: Malcolm North.

FIGURE 2 7.8 A prescribed burn at Blodgett Experimental Forest burning at low intensity and effectively reducing fuel loads. Photo: Kevin Krasnow.

removed (“diameter limits”) and minimum residual canopy cover. These metrics are set by the standards and guidelines in planning documents (e.g., SNFPA 2004). Diameter limits and canopy cover requirements are intended to ensure that treatments will move forest structure toward an “old forest” condition. If the diameter limit is set too high, large trees that do not substantially affect fuel conditions might be removed (Bigelow and North 2012). If the diameter limit is set too low, treatment might not produce the open conditions described by studies of historical forest structure (Beaty and Taylor 2007, 2008, Collins et al. 2011, Taylor et al., 2013) or create enough openings to regenerate shade-intolerant, fire-resistant species such as pines (Bigelow et al. 2011). Fire, as both prescribed burning and managed wildfire, is generally underused for fuels treatment (Figure 27.8). Although ecological restoration of these forests requires fire, numerous constraints limit its use (Collins et al. 2010; North, Collins et al. 2012). These include impacts to local communities from smoke production, reduced recreation opportu568  Ecosystems

nities, inadequate personnel to conduct and monitor fires, liability for fire escapes, and risk-adverse policies and institutions. Many concerns about fuel treatment intensity and fire use are inherently social in nature (McCaffrey and Olsen 2012). Addressing these issues will require more focused engagement and education of local communities and the general public to balance shorter-term impacts with the potential for longer-term benefits.

Increasing Forest Heterogeneity and Resilience Efforts to increase forest resilience have emphasized management strategies that work with and adapt to dynamic ecological processes at different scales (North et al. 2014). Management is now often focused on restoring heterogeneous forest conditions consistent with how productivity and historical fire intensity affected stand- and landscape-level forest conditions (North and Keeton 2008, Lydersen and North

2012) (see Figure 27.6). Forest managers use existing stand conditions and topography as a template to vary treatments in order to simultaneously achieve objectives such as fire hazard reduction, provision of wildlife habitat, and forest restoration (Knapp et al. 2012; North, Boynton et al. 2012). Within stands, thinning treatments attempt to create the ICO (individual tree, clumps of trees, and openings) structure that would have been created by frequent fire. Managers vary the proportions and sizes of these three structural conditions with small-scale changes in soil moisture and microclimate conditions (North, Stine et al. 2009, North 2012). At a larger scale, managers try to produce the forest density and composition associated with different slope positions and aspects that affect productivity and would have influenced fire severity (Underwood et al. 2010; North, Boynton et al. 2012).

Restoration Successes Several national parks in California have recognized the importance of fire in montane forests and been able to overcome the many challenges associated with managing fire. Two of the most notable examples are the Illilouette basin in Yosemite National Park and the Sugarloaf basin in Sequoia–​K ings Canyon National Park. In both areas, lightning-ignited fires have been allowed to burn relatively unimpeded since the early 1970s (see Lead Photo for this chapter). Although both areas experienced several decades of fire suppression, fire occurrence since the onset of natural fire programs in the 1970s is similar to that in the historical period (1700–​1900, prior to fire suppression) (Collins and Stephens 2007). In addition, fire effects and interactions among fires in the program are consistent with our understanding of how historical fires burned in these landscapes (Collins et al. 2007, Collins et al. 2009, Collins and Stephens 2010, van Wagtendonk et al. 2012). This suggests that fire in both areas might approximate a restored regime. This type of restoration cannot likely take place across much of the montane forest region with managed fire alone. However, the examples of Sugarloaf and Illilouette basins, as well as other areas with successful managed fire programs such as Lassen Volcanic National Park, illustrate the potential to expand fire use to meet restoration objectives. An important objective of these programs is to allow fires to burn under a range of fuel moisture and weather conditions rather than only under the fairly extreme conditions associated with “escaped” wildfires common on Forest Service land managed for fire suppression (Miller, Collins et al. 2012; North, Collins et al. 2012, Lydersen, North et al. 2014). The Forest Service and other landowners sometimes object to wider use of fire. Their reasons can include pursuit of multiple objectives, air quality restrictions, and lack of budget and personnel.

Future Scenarios Drought and Bark Beetles Warming temperatures will probably reduce the depth and duration of montane snowpacks, lengthening and deepening the summer drought. This will likely increase moisture stress for many forests (Safford, North et al. 2012, McDowell and Allen 2015) (Figure 27.9). Climate models currently do not agree on future precipitation patterns in California, but they all predict temperature increases and greater year-to-year

variability. This will likely mean more pronounced El Niño/ La Niña cycles that drive cycling between moderate snowpacks and potentially none at all for montane forests. These drought cycles could become bottlenecks for forest regeneration, killing most seedlings and saplings in dry conditions that are pronounced and/or occur in sequential years (Gray et al. 2005; North, Hurteau et al. 2005). Climate change could also increase bark beetle populations because warming can allow extra generations to complete their life cycles each year and adult beetle emergence and flight to occur early in the season and to continue further into the fall (Fettig 2012). Mountain pine beetles will likely become especially damaging to higher-elevation conifer forests (Bentz et al. 2010). Large, warming- and drought-driven beetle outbreaks have recently occurred in the U.S. and Canadian Rockies (Kurz et al. 2008) and might occur in California’s montane forests in the future. Bark beetle populations currently restricted to the southwestern U.S. and Mexico will also likely move northward as climates warm.

Fire The combination of warmer climate and possibly increased fuel production (due to lengthened growing seasons) will likely cause more frequent and extensive fires throughout western North America (Price and Rind 1994, Flannigan et al. 2000, Committee on Stabilization Targets for Atmospheric Greenhouse Gas Concentrations et al. 2011, Yue et al. 2013). A recent study from the northern Sierra Nevada indicates noticeable increases in the occurrence of high-to-extreme fire weather since the mid-1990s (Collins 2014). These increases, which are expected to continue at least into the near future, are likely contributing to the rising incidence of large fires in the region (Collins 2014, Lydersen, North et al. 2014). Fire responds rapidly to changes in climate and could overshadow the direct effects of climate change on tree species distributions and migrations (Flannigan et al. 2000, Dale et al. 2001). Under most climate change projections, fire will increase in frequency, size, and severity (Flannigan et al. 2009). The human population of California is expected to increase to more than fifty million by 2050 with a large increase in wildland/urban interface settlements. While educational efforts can help to reduce fire ignitions and improve public safety, more people usually leads to more fire (Syphard et al. 2009). Increased frequencies and intensities of fire in coniferous forest in California will almost certainly drive abrupt changes in tree species compositions and will likely reduce the size and extent of old-growth forest conditions (McKenzie et al. 2004, Stephens et al. 2013).

Species Distribution Projected changes in California’s terrestrial avifauna and flora are likely over the next century. Stralberg et al. (2009) developed current and future species distribution models for sixty focal bird species and found that novel avian assemblages with no modern analogy could occupy over half of California. This implies a dramatic reshuffling of avian communities and altered pattern of species interactions, even in the upper elevations of the Sierra Nevada, where only a modest proportion of novel avian communities were projected. A similar study projected that 66% of California’s native flora will experience greater than 80% reduction in range size within a cenMon tane Forests   569

mm 0–200 201–400 401–600 601–800 801–1,000 1,001–1,200 1201–1,400

A

0

B

C

1,401–1,473

100 200 km

FIGURE 2 7.9 Current (A) and future (B, C) projections of climatic water deficit for California’s montane forests. Climatic water deficit is the amount of water (scaled in millimeters [mm]) by which potential evapotranspiration exceeds actual evapotranspiration indicating relative drought stress. Projections based on (B) the Parallel Climate Model (PCM) and (C) the Geophysical Fluid Dynamics Laboratory CM 2.1 model (GFDL), using the A2 (medium-high) CO2 emissions scenario. Illustration by Jim Thorne.

tury (Loarie et al. 2008). Their study identified the southern Sierra Nevada and the coastal mountains of northwest California as climate change refugia, defined as areas projected to harbor species with shrinking ranges (presumably retaining subsets of regional species assemblages over time). Loarie et al. (2008) recommended novel adaptive management approaches and large-scale planning efforts that promote landscape and regional habitat connectivity. They also recommended serious consideration of human-assisted dispersal of California’s flora and prioritization of climate change refugia for conservation and restoration. California’s montane forests have withstood the pressures of the state’s burgeoning human population, frequent droughts and the long-term, general absence of its keystone process—​frequent, low-intensity fire. Yet the future promises that these stressors will persist and possibly be amplified by climatic change. The challenge to conserving California montane forests into the future is to increase their resilience while sustaining the old growth, wildlife, and ecosystem services that make them so unique among the world’s temperate forests.

Summary The strong, seasonal drought and historically frequent fire associated with a Mediterranean-type climate shape the composition and distribution of California’s montane forests. Differences in fire intensity and soil moisture availability associated with small- and large-scale topographic features such as drainages, aspect, and slope position affect ecosystem pro570  Ecosystems

ductivity and processes as well as ecosystem resilience to the most common stressors: fire, drought, and bark beetles. The resulting forest is highly heterogeneous, and the range of habitats—​f rom dry, open woodlands with understory shrubs to dense, mesic, multistory stands—​supports the highest vertebrate diversity of California’s forest types. Sensitive and threatened species are most associated with forest structures and habitat that have become increasingly rare after a century of logging and fire suppression. Management of these forests on public lands tends to focus on reducing densities of trees and fuels accumulated from fire suppression and increasing frequency and extent of low-intensity burns. This type of burning has demonstrated potential to restore many ecosystem processes that have stalled in the long absence of fire and to increase forest resilience to stresses likely to increase under climate change, such as drought and pests. Montane forests provide important ecosystem services to the state’s large and growing population, including much of its water, hydroelectric power, and substantial carbon storage, which can help offset human CO2 emissions. Although many challenges confront montane forests as human population and rural home construction increase, lessons learned from past forest management and progressive use of fire by the National Parks provide future pathways for sustaining and improving the ecological resilience of these forests.

Acknowledgments We would like to thank Ross Gerrard, who provided the montane ecosystem map, and Carl Skinner, who provide feed-

back and photographs—​both of the U.S. Forest Service Pacific Southwest Research Station.

Recommended Reading Fites-Kaufman, J., P. Rundel, N. L. Stephenson, and D. A. Weixelman 2007. Montane and subalpine vegetation of the Sierra Nevada and Cascade Ranges. Pages 456–​501 in M. Barbour, T. Keeler–​Wolf, and A. A. Schoenherr, editors. Terrestrial Vegetation of California. University of California Press, Berkeley, California. North, M., P. Stine, K. O’Hara, W. Zielinski, and S. Stephens. 2009. An ecosystem management strategy for Sierran mixed–​conifer forests. Pacific Southwest General Technical Report. PSW–​GTR-220. U.S. Department of Agriculture Forest Service, Albany, California. Safford, H. D., M. North, and M. D. Meyer. 2012. Climate change and the relevance of historical forest conditions. Pages 23–​45 in M. North, editor. Managing Sierra Nevada Forests. General Technical Report PSW-GTR-237. U.S. Department of Agriculture Forest Service, Pacific Southwest Research Station, Albany, California. Stephenson, N. L. 1998. Actual evapotranspiration and deficit: Biologically meaningful correlates of vegetation distribution across spatial scales. Journal of Biogeography 25(5):855–​870.

Glossary Abiotic  Not associated with or derived from living organisms. Abiotic factors in an environment include factors such as sunlight, temperature, wind patterns, and precipitation. Alfisols  One of twelve soil orders in the U.S. Soil Taxonomy, Alfisols make up 9.6% of global soils. They are primarily in cool, moisture regions of the Northern Hemisphere and have sufficient water to support at least three consecutive months of plant growth. They have high-to-medium base saturation, are moderately weathered, and are rich in iron and aluminum. Andisols  One of twelve soil orders in the U.S. Soil Taxonomy, Andisols account for only 0.7% of soils globally. They are formed from volcanic parent material, are high in organic matter content and phosphorous, and have a low bulk density. Basal area  A sum of the cross-sectional area of trees stems, measured by the diameter at breast height (dbh) (1.3 meters above the ground) and standardized to a hectare or acre area. It is a commonly used forestry measure that indicates the relative amount of biomass (and by implication resource use) of different sizes and species of trees within a stand. Canopy closure  This is a point measure of how much of the sky hemisphere is obscured by vegetation. Canopy cover  This is a stand-level average of how vertically porous a forest canopy is. Diameter distribution  The number of trees in different diameter-size classes. It is a widely used measure in forestry that provides insight into a stand’s structure and disturbance history. Edaphic  Produced or influenced by the soil. Forest resilience  The capacity of a forest to absorb disturbance and reorganize while still retaining its essential structure, composition, and ecological functions. Inceptisols  One of twelve soil orders in the U.S. Soil Taxonomy, Inceptisols (9.9% globally) often lack distinctive subsurface horizons. Inceptisols are generally found in landscapes with continuously eroded conditions or areas with young deposits. Seral  A phase in the sequential development of a community. Ultisols  One of twelve soil orders in the U.S. Soil Taxonomy, Ultisols (8.5% globally) have low base saturation at depth.

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Mon tane Forests   577

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T WENT Y-EIGHT

Subalpine Forests CONSTANCE I. M I LL AR and PH I LIP W. RUNDEL

Introduction Subalpine forests in California, bounded by the treeline at their upper margin, are the forest zone influenced primarily by abiotic controls, including persistent snowpack, desiccating winds, acute and chronic extreme temperatures, soil moisture and evapotranspirative stresses in both summer and winter, and short growing seasons (Fites-Kaufman et al. 2007). Subalpine forest species derive their annual precipitation primarily in the form of snow. Disturbances such as fire, and biotic interactions including competition, are less important than in montane forests. Although some subalpine forests are dense and have closed canopies, most are more accurately considered woodlands, with short-statured individuals and wide spacing of young as well as old trees. Subalpine forest stands are commonly interrupted by areas of exposed bedrock, snowfields, and upland herbaceous and shrub types—​ the latter comprising important components of broader subalpine ecosystems (Figure 28.1; Rundel et al. 1990, Sawyer et al. 2009). Subalpine forests comprise the highest-elevation ecosystems in California dominated by trees. Although scattered upright trees and wind-swept, shrubby individuals (krummholz) grow sparsely in the alpine zone, subalpine forests have

their upper limit at the alpine-treeline ecotone. Treeline has long fascinated ecologists for its predominance worldwide, from equatorial tropical forests to polar zones. While many environmental factors mediate the exact location of regional treelines—​a “devil-is-in-the-details” that also delights ecologists—​a robust unifying theory has been developed to explain the treeline ecotone as the thermal contour (isotherm) on the landscape where average growing-season temperature is 6.4°C (Körner and Paulsen 2004, Körner 2012). In this context “trees” are defined as plants having upright stems that attain height ≥3 meters regardless of taxonomy, and “forest” is characterized as more-or-less continuous patches of trees whose crowns form at least a loose canopy (Körner 2007). Although not without some controversy, the hypothesized mechanism behind the global treeline isotherm relates to the fact that upright trees are more closely coupled with the atmosphere than shorter-statured vegetation types such as those found in the alpine zone. This coupling is tightly interrelated with rooting zone temperatures, tissue thermal capacities, primary production (photosynthesis and carbon allocation), water transport, canopy shade, snowfall filtering, and relationships of incoming solar radiation. 579

FIGURE 28.1 Typical woodland structure of

California’s subalpine forest ecosystems, characterized by scattered trees and abundant rocky ground. Pinus albicaulis forest type, Humphreys Basin, Sierra Nevada. Photo: Constance Millar.

Two corollaries follow from this treeline mechanism: that mean growing-season temperature mechanistically translates into a life-form boundary (the alpine-forest ecotone), and that treeline should not be strictly related to elevation. Nonetheless, for a particular region, elevation provides a rough proxy for the thermal treeline. The treeline isotherm logically rises where local conditions are warmer (e.g., south slopes), depresses where cooler (north slopes), and varies by latitude as well as regional climate regimes. California traverses more than nine degrees of latitude, and thermal treeline elevations also vary among the mountain regions of the state. They are lowest in the north, where they range from about 2,700 meters near Mount Shasta to 2,800 meters on Mount Lassen. At similar respective latitudes, treeline elevation is slightly lower in the Klamath Mountains to the west and slightly higher in the Warner Mountains to the east due to differing climate regimes and species compositions. In the Sierra Nevada thermal treeline ranges from 2,800 meters in the northern forests; to 3,000 meters near Donner Pass; to 3,200 meters in the Yosemite region; and to 3,500 meters in the southern Sierra Nevada (Rundel 2011). Thermal treeline in the Great Basin ranges to the east of the Sierra Nevada are slightly higher than corresponding Sierran latitudinal positions. Treeline isotherm is the background regulator for the highest (coolest) occurrence of subalpine forests; however, local environmental factors control the specific position (including elevation) of upper subalpine forests. These include slope and aspect, substrate type and geomorphology, avalanche occurrence, and other disturbance history. This “ecological noise” can be critically important for ecosystem function and diversity and reminds us that changes in treeline position over time (or lack of change) are not necessarily indicators of climate change. Subalpine forests in California include communities dominated by whitebark pine (Pinus albicaulis), foxtail pine (P. balfouriana), limber pine (P. flexilis), western white pine (P. monticola), mountain hemlock (Tsuga mertensiana), or Sierra juniper (Juniperus grandis, formerly J. occidentalis var. australis). In addiPhoto on previous page: Long-lived bristlecone pines of the White Mountains are emblematic of subalpine forest ecosystems in California. Photo: Constance Millar. 580  Ecosystems

tion, lodgepole pine (Pinus contorta) also commonly occurs in subalpine forests in California, either as the dominant species or intermixed with others. Because it extends across many more environments than subalpine, including elevations down to sea level, lodgepole pine alone is not an indicator of subalpine forests. In addition to these conifers, several very small stands of otherwise wide-ranging subalpine fir (Abies lasiocarpa) grow in the Trinity Alps and Marble Mountains of northwest California, and several tiny stands of Alaska yellow-cedar (Callitropsis nootkatensis, formerly Chamaecyparis nootkatensis) occur in the Siskiyou Mountains; these species are indicators of the subalpine zone at these rare locations. The hardwoods quaking aspen (Populus tremuloides) and curl-leaf mountain mahogany (Cercocarpus ledifolius) also grow commonly in subalpine environments, but because they extend abundantly to lower montane zones, they are not indicator species. Whereas the upper bounds of subalpine forests have a robust, thermal delineation and form a visible transition from forest to alpine vegetation, the lower limits of the subalpine zone are less distinct. These generally follow the elevation of snowpack dominance, which strongly influences tree species diversity. The subalpine/montane forest ecotone is also controlled by shifts in fire regimes (Caprio and Graber 2000, Minnich 2007). On the one hand, while the dense canopies and surface fires of lower-elevation red and white fir (Abies magnifica, A. concolor, respectively) limit establishment of subalpine species, high-intensity fires burning downslope from lodgepole pine or hemlock forests can create openings in the fir forests and expose mineral soils. In these cases, subalpine species can advance downslope until succession of fir regains dominance uphill. As with upper treeline, elevation only roughly defines lower limits of the subalpine zone, and these vary with latitude across the state. Lower boundaries extend to 2,200 meters in the Klamath Mountains; 2,300 meters at Mount Shasta; 2,400 meters in the northern Sierra Nevada; 2,750 meters in the southern Sierra Nevada; 2,900 meters in the southern California mountains; and 3,000 meters in the Great Basin ranges (Griffin and Critchfield 1976, Elliott-Fisk and Peterson 1991, Holland and Keil 1995). In California today, subalpine forest ecosystems conservatively extend over 390,270 hectares of California (Figure 28.2,

FIGURE 28.2 Distribution of subalpine forest ecosystems in California. Source: Data from U.S. Geological Survey, Gap Analysis Program (GAP). Map: P. Welch, Center for Integrated Spatial Research (CISR).

Environmental Controls

TA B L E 2 8 .1 Area and percentage of total subalpine forests in California by mountain region

Area (hectares)

Mountain region South CascadesA Great Basin North

B

Klamath Mountains

C

Sierra Nevada Great Basin, Central

D

Great Basin, Southern Southern California Total

F

E

Percentage of total

284

0.1

3,494

0.1

77,920

20.1

290,830

75.2

10,590

2.7

376

0.1

6,777

1.7

390,271

100

Data from U.S. Geological Survey, Gap Analysis Program (GAP). A. Mounts Shasta and Lassen B. Warner Mountains C. Marble Mountains, Trinity Alps, Salmon Mountains, Yolla Bolly Mountains D. Sweetwater Mountains, White-Inyo Range E. Panamint Range F. Tehacapi Mountains, San Gabriel Mountains, San Bernar­ dino Mountains, San Jacinto Mountains

Table 28.1; Davis et al. 1998) and occur in the Klamath Mountains, including the Marble Mountains, Trinity Alps, Mount Eddy, Salmon Mountains, and North and South Yolla Bolly Mountains; southern Cascade Range, including Mounts Shasta and Lassen; Sierra Nevada; Great Basin ranges, including the Warner Mountains, Carson Range, Zunamed Mountains (Charlet 2014), and Sweetwater Mountains; Glass Mountains, Mono Craters, White-Inyo Range, and Panamint Range; and southern California ranges, including the Tehachapi Mountains, San Gabriel Mountains, San Bernardino Mountains, and San Jacinto Mountains (Griffin and Critchfield 1976). Forests types differ across mountain regions of the state in overall tree species diversity as well as species dominance, diversity of affiliated nonarboreal vegetation, faunal relations, climate interactions, productivity, and biogeochemistry. Although subalpine forests commonly occur on all slope aspects of California mountain ranges, they are limited to wetter aspects in more arid regions such as southern California and the Great Basin. These usually include western slopes in southern California (Holland and Keil 1995) and northern slopes in the Great Basin (Elliott-Fisk and Peterson 1991). The large majority of subalpine forest ecosystem in California occurs in the Sierra Nevada, with more than 75% of the total (see Table 28.1). Subalpine forests dominate in a broad band on the gradual west slope of the Sierras and in a narrow, less diverse and more scattered band on the steep eastern escarpment. The areas high enough to support subalpine forest in the jumbled Klamath Mountains of northwest California collectively amount to the second largest region, with 20% of the state’s total. The southern Cascades support a deceptively small amount of subalpine forest (less than 1%), which results from the narrow perimeter area around Mounts Shasta and Lassen. The remaining mountain regions of the Great Basin and southern California each also contain less than 1% of the total subalpine forest in the state (see Table 28.1). 582  Ecosystems

Geology, Geomorphology, and Soils The environmental context for California’s subalpine ecosystems derives from the unique sequence of historical geologic processes that gave rise to its upland regions (see Chapter 8, “Ecosystems Past: Vegetation Prehistory”). Subalpine forests have shifted greatly in diversity and geography over the past thirty million years as topography changed in the California region. Prior to that time, California was mostly under water and/or characterized by lowlands with subtropical climates. Mountain ranges of the pre-Sierra/Cascade cordillera first emerged as eruptive centers along the subduction plate boundary that defined the Pacific margin of North America more than seventy-five million years ago (Millar 2012). Tectonic action related to plate boundaries led to emplacement of magmatic batholiths (subsequently granitic rocks) deep below the continent. Plate-boundary tectonics also catalyzed extensive aboveground volcanoes that defined the Nevadan and Sevier orogenies and led to development of the extensive Nevadaplano, with high-elevation summits that extended across present-day eastern California and Nevada. This early volcanism largely defined the stage for subsequent bedrock exposures, soil development, and geomorphic conditions supporting subalpine forests in California today. On the highest ranges and especially in the arid ranges where erosion has been minimal (e.g., the White-Inyo Range and parts of the southern Sierra Nevada), highly metamorphosed rocks called roof pendants occur and date to times when California was submerged under sea. These rocks are often characterized by complex, colorful, and tortuously folded strata, including formations of limestone, marble, and other carbonate substrates. Where they are exposed, unusual chemical compositions and pH levels constrain plant growth to species able to tolerate these conditions, with bristlecone pine on dolomite substrate as an example. Also dating to these eras are exposures of ultramafic and serpentine rocks, with patchwork soils of complex origin primarily derived from former oceanic terranes subsequently accreted into California. Soils derived from these rocks also present nutritional limitations for plant growth and exclude many taxa. Tolerant subalpine species such as foxtail pine and western white pine can be found on these soils in the few locations where they are exposed at high elevations, primarily in northwest California. In eastern California hydrothermal alteration of volcanic rocks created substrates with another type of unique chemistry limiting plant growth. Subalpine conifers such as lodgepole pine and limber pine are able to grow on these soils, and are often found on these substrates in very disjunct locations and at much lower elevations than usual, in zones otherwise dominated by montane or woodland conifers. Far more extensive substrates underlying California subalpine forests are granitic rocks and associated soils that derive from the early magmatic plutons of subduction plate dynamics. These were exposed over subsequent eras during the processes of mountain-building and erosion by glaciers, water, and wind. Granitic rocks create soils that favor growth of many subalpine conifer species, with characteristics such as coarse grain that enable drainage yet adequate water-holding capacity, intermediate to moderate acidity, and a sufficient balance of vital plant nutrients. In some regions, such as the Great Basin ranges, southern Cascades, and eastern Sierra Nevada, geologic hot spots occur where range-front faulting

is active or magmatic centers are shallow. In these locations volcanism has continued from the late Tertiary into present times. Soils that develop in these regions, especially from Quaternary eruptions such as Mounts Shasta and Lassen and the Glass Mountains in eastern California, are poorly developed and challenge plant growth.

Current Climate and Climate Variability Although subalpine forest ecosystems in California lie within the general Mediterranean-climate regime of the state, high elevations modify its influence. For instance, as elevation increases, temperatures and evaporative demand decrease, reducing the stress of the otherwise long summer drought. The subalpine forest zone in California is characterized by short growing seasons (six to nine weeks), prolonged winter snowpack (usually deeper than 2 meters except in the Great Basin ranges), and cool summer and winter temperatures with frost possible any month (Agee 1993, Fites-Kaufmann et al. 2007). Proximity to the Pacific Ocean and dominance of prevailing storms from the west protect these high-elevation ecosystems from extreme cold, although the Great Basin ranges of eastern California experience more continental climates. These include greater extremes, especially of cold temperatures in winter, than other mountain regions in the state experience. Annual and monthly temperatures tend to be cooler as the subalpine zone rises in elevation (i.e., with decreasing latitude) and in interior ranges, regardless of latitude (Table 28.2; PRISM climate model, Daly et al. 1994). Precipitation falls on subalpine forests mostly as winter snow. Summer precipitation derives from local convectional storms, which vary in intensity and abundance across the range of subalpine forests as well as by topographic position within ranges (Fites-Kaufman et al. 2007). Gradients of precipitation occur in both latitude and longitude. Annual precipitation, including winter snowfall, is generally highest in the northern mountains, including the Klamath Mountains and southern Cascades, which can approach conditions of the Pacific Northwest, and lowest in the semiarid regimes of the southeastern Great Basin ranges (see Table 28.2). Despite their southerly latitude, precipitation in subalpine regions of southern California is similar to locations in the central Sierra Nevada, though far less than in the southern Cascades and Klamath Mountains. Precipitation also varies strongly across heterogeneous environments within mountain ranges, so some subalpine sites receive high precipitation despite their location in a generally dry region and vice versa. More precipitation falls in summer in California’s southern subalpine forests than in northern forests due to the Gulf of California monsoon influence (see Table 28.2). In the southeastern Great Basin ranges, for instance, summer monthly precipitation is about equal to the winter amount, although annual averages are an order of magnitude lower than in northern mountains. July tends to be the driest month in the subalpine zone, with increasing precipitation in August and September. This trend reflects the various influences of summer convective activity and monsoon, especially in the southern regions; and early snowfalls, especially in northern regions. Longitudinal trends also occur in precipitation across the subalpine regions of California, with mountains nearer the Pacific Ocean (e.g., Marble Mountains, Yolla Bolly Mountains) generally receiving more annual precipitation (including winter snowpack) than progressively inland ranges

at the same latitudes. This results from California’s regional rain shadow (see Chapter 2, “Climate”). Rainshadow effects are also common within mountain ranges and shape subalpine forest composition and structure at local and regional scales. These result from local orographic effects, where moisture-laden clouds condense as rain when clouds rise on west slopes of the mountains and evaporate on the east slopes. Orographic processes, even over short distances across range crests, can translate to large differences in annual precipitation for local subalpine forests. Snowfall and snowpack data and models are lacking for most of the state’s subalpine regions. In California, SNOTEL sites (automated snow-measuring stations run by the U.S. Department of Agriculture Natural Resources Conservation Service) exist only in the Warner Mountains, central-eastern Sierra Nevada, the Carson Range, and the Sweetwater Mountains, and most stations are located in the upper montane forest zone rather than in the subalpine. The sufficiently high stations, however, provide a window into snowfall depth and interannual variation in subalpine forests across regions (Figure 28.3a). The trend of snowfall follows an expected geographic pattern, with latitude trumping orographic effects. One of the northernmost sites (Dismal Swamp), in the interior Warner Mountains, has the highest April 1 snow depth over the years of all sites. Snowpack generally decreases from north to south among the Sierra Nevada stations with the lowest depths at the southernmost station, Virginia Ridge just north of the Mono Basin. Snow depths also decrease eastward in the Great Basin, including the Carson Range (the Heavenly Valley station) and the Sweetwater Mountains (the Lobdell Lake station). While it is common to define ecosystem envelopes by their temperature and precipitation parameters and to compare differences in these variables among regions, factors related to water availability and timing—​not too much, not too little, when needed—​are often more important in this region (Stephenson 1998). Temperature is important in controlling upper treeline, but evaporative stress, often measured through soil moisture interactions and climatic water deficit (CWD), strongly influences subalpine distribution of species at lower elevations and interior dry margins. Intrinsic differences in evaporative demand and water supply regulate the ability of trees to survive and grow. Local topographic and substrate effects, interacting with rainfall and snowfall, determine the amount and retention of soil moisture and lead to differences in plant growth on soils of differing water-holding capacities, such as granitic versus metamorphic substrates. Similarly, differences in elevation of forests on different aspects reflect available growing-season soil moisture, which drives the presence of subalpine forests about 200 meters higher on steep, south-facing slopes than on steep, north-facing slopes (FitesKaufmann et al. 2007). In subalpine forests, CWD values are generally greater than 200 millimeters and are important in distinguishing the niche for this forest type from lower montane forests (Stephenson 1990, 1998). Variation in interannual CWD can be a key trigger, especially when combined with chronic warm summer temperatures, for subalpine forest insect outbreaks and forest mortality (Millar, Westfall et al. 2007; Millar et al. 2012). While these summary patterns of temperature, precipitation, and available soil moisture define general boundary conditions, high interannual and interdecadal variation in California’s weather exerts important controls on vegetation distribution and structure. The primary drivers of this variSubalpine Forests   583

TABLE 28.2 Climate data for subalpine forest zones in California by mountain range

Temperature °C Jan Latitude (° W)

Longitude Elevation (° N) (M)

Location

Mountain range

Kings Castle Pk

Marble Mtns

41.616

123.222

Warren Peak

Warner Mtns

41.377

Mt Shasta, Panther Mdws

S Cascades

High Lake, Russian Wild.

July

Annual

Precipi­tation (mm)

max

min

max

min

max

min

mean

Jan

July

Aug

Sept

Annual

2212

1.5

-4.7

20.7

8.9

9.8

0.7

5.3

528

11

17

53

2902

120.219

2820

-2.1

-10.5

18.7

3.0

6.9

-4.7

1.1

162

14

21

39

1163

41.357

122.195

2454

1.0

-6.5

19.9

6.5

8.8

-1.1

3.9

332

10

16

38

2030

Salmon Mtns

41.298

122.956

2209

2.2

-4.8

20.6

9.5

9.8

1.0

5.4

249

16

23

43

1445

Caribou Lake

Trinity Alps

41.032

122.971

2133

1.9

-5.2

20.2

8.6

9.6

0.4

5.0

510

4

51

30

2134

Shadow Lake, Mt Lassen

S Cascades

40.480

121.472

2332

3.1

-6.3

21.9

6.9

11.2

5.1

488

12

36

54

2838

Lake Helen, Mt Lassen

S Cascades

40.475

121.503

2608

1.8

-7.2

19.9

5.9

9.4

-2.1

3.7

519

12

39

58

2994

S Yolly Bolly Peak

Yolla Bolly Mtns

40.037

122.863

2320

3.5

-4.6

22.1

10.0

11.5

1.5

6.5

364

6

10

25

1898

Mt Rose Saddle

Sierra Nevada

39.340

119.931

2970

0.4

-10.0

18.7

4.6

7.8

-4.3

1.8

229

21

24

44

1521

Ebbetts Pass, Highland Pk

Sierra Nevada

38.489

119.798

2754

2.5

-8.3

19.4

6.3

9.8

-2.3

3.8

234

27

40

38

1391

Ellery Lake, Tioga Pass

Sierra Nevada

37.934

119.238

2930

1.8

-9.9

19.4

5.6

9.3

-3.4

3.0

135

21

16

27

752

Arrowhead Lk, Mammoth

Sierra Nevada

37.581

118.981

3007

3.6

-8.2

20.3

6.4

10.6

-2.4

4.1

275

8

4

18

1348

Crooked Creek

White Mtns

37.506

118.170

3140

0.2

-11.8

17.8

3.8

7.8

-5.3

1.3

33

28

35

21

391

First Lake, N Palisades

Sierra Nevada

37.129

118.487

3073

1.7

-10.3

18.4

4.4

8.8

-4.0

2.4

132

10

13

39

836

Wacoba Mtn

Inyo Mtns

37.025

118.002

3105

2.4

-8.3

19.0

6.5

9.4

-2.4

3.5

33

25

29

19

346

Cottonwood Basin

Sierra Nevada

36.489

118.200

3262

1.0

-10.7

16.8

3.4

7.7

-5.0

1.4

87

7

8

18

511

Telescope Peak

Panamint Range

36.175

117.092

3150

1.9

-8.9

21.0

6.0

10.8

-2.5

4.2

39

20

39

34

362

Cucamonga Peak

San Gabriel Mtns

34.222

117.586

2700

6.7

-2.9

23.3

10.3

12.8

1.8

7.3

231

6

15

31

1142

San Gorgonio Mtn

San Bernardino Mtns

34.103

116.822

3201

1.7

-10.5

17.8

2.6

8.5

-5.3

1.6

261

17

25

34

1306

2757

1.9

-7.9

19.8

6.3

9.5

-2.1

3.7

255

15

24

Average

35

1437

Source: Data excerpted from the PRISM climate model (Daly et al. 1994) for point locations selected as representative of the mid-upper subalpine zone for the region. PRISM data represent 1971-2000 normals, with 800 m grid.

FIGURE 28.3 Snow depths from SNOTEL snow-monitoring stations in California subalpine zones. Source: Data extracted from NRCS SNOTEL station data, 2013. A Snow-depth variation (April 1) across diverse subalpine locations in the Warner Mountains (Dismal Swamp, Cedar Pass), Sierra Nevada

(Independence Lake, Squaw Valley, Echo Peak, Burnside Lake, Summit Meadow, Virginia Ridge), Carson Range (Heavenly Valley), and Sweetwater Mountains (Lobdell Lake), 1999–​2 013. B Monthly snow depth and temperature variation at the Virginia Ridge SNOTEL site in the subalpine forest of the Sierra Nevada (2879

meters), 2001–​2 013.

ability are forcing mechanisms related to ocean circulation. Precipitation especially varies in episodic, often quasi-cyclic manners, and in patterns that vary across California (Redmond and Koch 1991, Abatzoglou et al. 2009). In particular, the El Niño–​Southern Oscillation (ENSO) (Cayan et al. 1999) and multidecadal modes, including the Pacific Decadal Oscillation (PDO) and Atlantic Multidecadal Oscillation (Mantua et al. 1997, Cayan et al. 1998, McCabe et al. 2004), drive large differences in precipitation delivered to California among years and among decades (see Chapter 2, “Climate”). An example of vegetation response to these variations is the episodic response of lodgepole pine recruitment into subalpine meadows during the twentieth century, which occurred during negative phases of the PDO (Millar et al. 2004). Particularly important for biota are recurring multiyear droughts 586  Ecosystems

that characterize the instrumental record (Cayan et al. 1999) as well as historical reconstructions (Biondi et al. 2001, Cook et al. 2007); these droughts often trigger forest insect and pathogen infestation (e.g., limber and whitebark pine; see Millar, Westfall et al. 2007; Millar et al. 2012). Recent research is also elucidating the importance of atmospheric rivers (Dettinger 2013) as a significant determinant of interannual variability in precipitation that, at elevations of subalpine forests, translates to large differences in snowpack (McCabe and Dettinger 1999) (see Figure 28.3b). California’s high wind speeds may occur anywhere within the state, with the greatest velocities at high elevations (WRCC 2013). An important controlling factor exerted by wind on subalpine forests is in combination with snowfall and topography, which together influence snowpack drifting

A

B

B

FIGURE 28.4 Wind-sculpted and wind-thrown whitebark pine forests near treeline. Photos: Constance Millar. A Stunted trees and krummholz matts, Mount

Dunderberg, Sierra Nevada. B Windthrow in whitebark forests as a result of the

autumn 2011 extreme downslope wind event, Tioga Crest, Sierra Nevada.

and variability in snow depth. Patches and zones of deeper, wind-influenced snowdrifts define areas that retain moisture late into the growing season (persistent snowfields) and have higher CWD. Many of these locations support subalpine forest but are surrounded by dry upland slopes with herbaceous or shrub cover. Scattered across the landscape, these support small to large “snowpocket forests,” which tend to occur on north aspects and in slumps, along stepped terrain, or in depressions. In interaction with forest density, wind also influences the distribution of snow accumulation under the forest canopy. Low-moderate forest densities typical of many subalpine forests maintain the highest amount of snowpack relative to either higher or lower densities (Raleigh et al. 2013, Lundquist et al. 2014). Chronic winds in exposed areas, especially in winter when tree crowns are not protected by snow, affect crown growth and shape (e.g., krummholz and branch flagging; Figure 28.4a) and limit tree regeneration to windsheltered sites. Windthrow is relatively rare in subalpine forests, given the inherent mechanical capacity of the species to accommodate

and resist wind. Occasionally, however, the “perfect storm” of atmospheric conditions coincides to produce monstrous wind events. The most recent and potentially largest recorded of these was an extreme downslope wind event in the central Sierra Nevada on November 30 and December 1, 2011. This extreme event was unusual for its wind direction (north), duration (over twelve hours), and sustained high velocities, which exceeded 145 kilometers per hour for the duration of the event with gusts over 240 kilometers per hour (Hilimire et al. 2013). Montane forests on the high west slopes of Yosemite National Park and Devil’s Postpile National Monument sustained massive, although localized, forest downfall, and subalpine whitebark pine forests of this region also experienced dramatic local areas of windthrow (see Figure 28.4b). Avalanches occur throughout the snow zone of California’s mountains but become more common with increasing elevation and steeper slopes. Avalanches exert locally important controls on subalpine forest ecosystems through effects on tree size, form, persistence, and species diversity. Severe avalanches uproot both mature and most young trees, Subalpine Forests   587

and recurring avalanches maintain slopes in treeless conditions, favoring sprouting shrubs such as alder (Alnus spp.). Avalanches also produce a variety of geomorphic effects on subalpine environments. These include scouring soils from hillslopes, maintenance of vertical troughs, accumulation of debris in the runout zone, and creation of impact and scour pits (Davis 1962). In some canyons avalanches are common enough (≥ one per decade) to give the slopes a striped appearance, where the tracks are treeless or with young tree cohorts and are separated by protected zones where mature forests can develop (Martinelli 1974, Mears 1992). Where avalanches are separated by intervals of several decades, conifers or aspen often regrow. A thick jumble of debris can remain in avalanche runout zones for decades if undisturbed and potentially influences other disturbances such as fire, insects, and disease. Along the edges of avalanche tracks, surviving trees are often broken and twisted with bark broken off—​conditions that stress trees and favor entry of insects. During heavy snowpack years, such as the record wet winter of 1985–​1986, avalanches occurred in unusually high numbers in the Sierra Nevada and toppled hundreds of hectares of subalpine forest (Wilson 1986, Kattleman 1996). A large proportion of trees were 125–​150 years old. Some trees destroyed near Sonora Pass were 350 years old. The effect of avalanches in that season on forest throwdown can still be seen from many transmountain passes, such as along Tioga Pass in Yosemite National Park.

Ecosystems of the Subalpine Forest Zone Subalpine ecosystems in California are commonly dominated by open stands of conifer forest. Local areas of deciduous broadleaf trees can also occur along riparian corridors or other areas with available water. Scattered areas of wet and dry meadows, often with associated shrublands, are present; extensive montane chaparral communities dominate some regions. Through all of the subalpine forest communities in California is a general pattern of decreasing stand densities and basal areas with increasing elevation (Pinder et al. 1997). These declines are associated with a complex mix of environmental and climatic factors, including decreasing soil depth and development, lower temperature, shortening of the growing season, increased wind, and increased effects of snowmelt depth and topography on water availability. Models of site moisture availability and irradiance coupled with field measurements of stand characteristics and tree-ring records suggest strong correlations of microsite conditions with age class (Bunn et al. 2005). Finally, these declines have also been associated with lower nutrient inputs from aboveground litter (Fites-Kaufman et al. 2007).

Subalpine Adaptations to Extreme Physical Conditions Many species of subalpine ecosystems, like those in the alpine zone, have evolved specialized adaptations to endure extreme climates and environments including rocky substrates with thin, poorly developed and often nutritionally impoverished soils; steep, unstable slopes that experience avalanches and landslides; and subfreezing temperatures, high and desiccating winds, and intense solar radiation. A case in point is Great Basin bristlecone pine, which grows in the White-Inyo and Panamint Ranges of California and many more ranges 588  Ecosystems

in Nevada and Utah. Throughout its distribution, bristlecone pine forests occur at the highest elevations and extreme exposures under cold, arid climates. Evolved adaptations to these conditions are many. One that contributes to the species capacity to persist in these environments is its needle retention, which is longer than other conifer species, reaching over fifty years (Barber 2013, Ewers and Schmid 1981). This unusual capacity enables trees to retain foliage and to photosynthesize (i.e., to survive) even during multiyear periods when weather conditions in the growing season are severe enough that new needles cannot develop. Waxiness and resin buildup on needles add to their durability as well. Another attribute contributing to the species’ persistence and great longevity is the capacity to form stripbark growth. This occurs when portions of the main stem (secondary cambium) die back as the tree ages. This leaves increasingly smaller strips of live stem (cambium) and bark on one side of the tree. Such stripbark trees can continue to grow for centuries and millennia—​a capacity shared with only a few other, and mostly subalpine, conifers. The stripbark habit is assumed to be an adaptation to the extreme climate conditions of the species’ range, enabling trees to “cast away” branches and stem as stress increases and remain alive with only part of the vasculature and crown functional. Many bristlecone pines, especially those that have developed stripbark, also have intensely spiral grain, known to be a highly heritable trait. This leads to a corkscrew form of the main stem, which has the effect of exposing more of the crown—​especially when a narrow strip as a result of stripbark—​to sunlight. High fecundity of bristlecone pine is known to persist throughout the life of individuals, and even trees more than three millennia in age produce many cones with fertile seed. Because replacement is very low for longlived individuals, this high fecundity of stands of mixed ages provides high genetic diversity for seedling generations that can be important for natural selection as climates change over the course of time (decades to millennia). Another example of adaptation to extreme conditions of the subalpine zone is the crown plasticity of several conifer species, especially the capacity to form krummholz. This ability to tolerate nonapical-dominance and to spread laterally allows species such as whitebark pine, limber pine, and mountain hemlock to remain below the sheltering influence of winter snowpacks, where temperatures are stable at freezing temperature and protected from desiccating winds.

Whitebark Pine Forests Whitebark pine is a wide-ranging treeline species that extends from central British Columbia east to Wyoming and south to the central Sierra Nevada (Weaver 2001). It forms the dominant treeline species in the southern Cascade Range and on the higher slopes in the Warner Mountains. Whitebark pine forests are scattered in the Klamath Mountains with populations on Mount Eddy, Thompson Peak, Russian Peak, and the Marble Mountains (Griffin and Critchfield 1976). At several locations in the Klamath Mountains, such as Crater Creek and Sugar Creek Research Natural Areas (RNAs), stands of exceptional subalpine diversity exist with high density, productivity, and basal area (Cheng 2004). In these areas, whitebark pine is one prominent subalpine forest type out of seven that commonly occur. Whitebark pine is common in the bands of subalpine eco-

systems that ring the southern Cascade volcanoes, especially Mounts Shasta and Lassen. Whitebark pine forests occur mixed with mountain hemlock as low as 2,103 meters along ridgetops of the Antelope Creek RNA, forming one of the lowest subalpine whitebark pine occurrences in this region (Cheng 2004). In the Sierra Nevada, whitebark pine ecosystems occur abundantly from the Lake Tahoe Basin south to Mount Whitney. In the central Sierra, whitebark pine typically is present in mixed stands with lodgepole pine, mountain hemlock, and Sierra juniper; while in the southern Sierra it grows with limber pine and slightly overlaps in range with foxtail pine (see Figure 28.1). A watershed study in Eastern Brook Lakes on the eastern slope of the Sierra Nevada at 3,170 to 3,780 meters found mixed dominance of lodgepole pine and whitebark pine. The mean leaf area index for canopies of whitebark pine was 4.6 m 2m2, compared to 4.1 m2m-2 for lodgepole pine (Peterson et al. 1989). Whitebark pine forests are considered keystone ecosystems for the subalpine zone throughout the cordillera of western North America (Tomback and Achuff 2010). Whitebark pine is highly plastic in crown and growth form and varies readily in response to severity of growing conditions. On favorable sites it can form upright, small trees 10 to 15 meters in height that live to 350 years. At higher elevations above treeline or exposed slopes below, its crown becomes stunted, often exhibiting gnarled and twisted branches in response to desiccating winds. A lower ground layer of prostrate crown is often present in these stands. In the treeline ecotone and up to 500 meters above treeline, whitebark pines readily take on a multistemmed krummholz form of growth, and finally a low mat of growth less than 1 meter in height (Fites-Kaufman et al. 2007). At these locations whitebark pine stands commonly form monotypic communities that dominate the upper treeline ecotone and play important roles in snowpack retention. Krummholz plants often root as the crowns spread across the ground, and individuals can live up to 1,700 years old (King and Graumlich 1998). Krummholz mats are often thought to be clonal , deriving from a single seed, but genetic studies show this is not the case, at least for krummholz trees with crowns larger than about 3 meters in diameter (Rogers et al. 1999). A single krummholz mat can comprise 2 to 12 genets, with genetic variation and genetic distance among individuals within the krummholz increasing in the downwind direction. Krummholz crowns are very dense and provide important hiding cover for small mammals, especially the white-tailed hare (Lepus townsendii). A remarkable coadaptation exists between whitebark pine cones and seeds and Clark’s nutcracker (Nucifraga columbiana), a midsized bird in the crow family (Tomback 2001). Increasing density in whitebark pine ecosystems in recent decades might relate to changes in behavior of Clark’s nutcracker in response to changing climates as well as to direct response by the pine. Whitebark pine forests exemplify a trend observed for other subalpine forests in California, with the exception of limber pine. Whereas there appears to be little significant advance of whitebark pine seedlings above the twentieth-century upper treeline, density in these zones has been steadily increasing throughout the century, with a net increase in the Sierra Nevada of 30%, including a 44–​91% increase in small tree densities (Dolanc et al. 2013). Correspondingly, the density of large trees has declined. These increases in small tree density are accelerating, especially above 3,000 meter elevation.

Western White Pine Forests Western white pine (Pinus monticola) extends from British Columbia through the Cascade Range and Klamath Mountains, through the northern Great Basin ranges of California, and throughout the Sierra Nevada, where it reaches its limit in southern Tulare County. In the Sierra Nevada it is a minor component of upper montane forests but becomes increasingly important in subalpine habitats, although monotypic stands are rarely more than a few hectares. Most commonly, western white pine mixes with lodgepole pine, Jeffrey pine, mountain hemlock, red fir, and/or whitebark pine (Potter 1998). Although Sierran trees of this species may reach 40 meters in height and 2.5 meters in diameter, larger sizes are attained by the same species in the northern Rocky Mountains and Pacific Northwest (Van Pelt 2001). Western white pine generally maintains an upright tree form of growth nearly to treeline, where it is commonly replaced by whitebark pine or foxtail pine depending on geography. Seedlings are reported to be relatively few compared to other subalpine conifers (Parker 1988).

Foxtail Pine Forests Foxtail pine is the dominant subalpine and treeline pine of the southern Sierra Nevada and is locally important in subalpine forests of the Klamath Mountains. It has highly disjunct populations, with the Sierran and Klamath distributions separated by hundreds of kilometers. These two groups of populations are well differentiated, with the southern Sierra Nevada taxon, subspecies austrina, morphologically distinct in the foliage, bark, cones, and seeds from populations of subspecies balfouriana in the Klamath Mountains (Mastrogiuseppe and Mastrogiuseppe 1980). The disjunction of these two populations is thought to relate to the development of summerdry Mediterranean climates during the late Tertiary (Millar 1996) further modified by effects of glacial/interglacial cycles of the Pleistocene and drought conditions of the mid-Holocene (Eckert et al. 2008). Foxtail pine in the Sierra Nevada is restricted to higher elevations (2,600–​3,660 meters) south of the Middle Fork of the Kings River. At its lower elevational limits it often occurs in open stands with lodgepole pine, Jeffrey pine, western white pine, and red fir. At higher elevations it forms relatively pure but low-density stands, although it often mixes with limber pine. Treeline stands of foxtail pine often show a preference for cooler, north-facing slopes (Rundel and Rabenold 2014), likely related to soil moisture availability (Bunn et al. 2005). Vankat and Major (1978) sampled stands of foxtail pine from elevations of 3,170 to 3,290 meters in Sequoia National Park and reported a relatively high mean density of 418 tree ha-1 and a canopy cover of 26%, with a basal area of 31 m 2ha-1. Tree densities and stand basal areas, however, decline with increasing elevation from foxtail pine woodlands to treeline (Lloyd 1997, 1998, Rundel and Rabenold 2014). Foxtail pines can grow to be several thousand years old. Like bristlecone pine, foxtail pine has highly resinous wood that with the cold, arid climates in the southern Sierra Nevada can persist as remnant dead wood for millennia. Together the live and dead wood are important archives for paleoclimatic and paleoecological study. Foxtail pines have been documented to respond to warm and cold historical climate periods by, respectively, advancing upslope and retractSubalpine Forests   589

FIGURE 28.5 Limber pine forests on the eastern

escarpment of the Sierra Nevada, south of Mammoth Lakes. Photo: Constance Millar.

ing downslope (Scuderi 1993, Lloyd and Graumlich 1997). In the Klamath Mountains, foxtail pine plays a more diverse ecological role than in the Sierra Nevada. Habitat heterogeneity at multiple spatial scales has been found to favor persistence of foxtail pine populations in northwest California (Eckert 2010). At large spatial scales, the presence of ultramafic (low silica content, often basic) soils favors this species relative to other conifers and leads to greater ecological importance.

Limber Pine Forests Limber pine has a wide range extending from central Alberta and South Dakota south to New Mexico in the Rocky Mountains and across the higher ranges of the Great Basin. In California limber pine is most common along the eastern escarpment of the Sierra Nevada, where it extends from scattered and disjunct stands in Buckeye Canyon near Bridgeport, California, then southward with increasing importance (Figure 28.5). The transition at the north between limber pine and whitebark pine forests appears to reflect the latter’s higher tolerance of high snowloads and long, dry summers. In Tulare County of the far southern Sierra, extensive limber pine forests occur on the west slope of the crest as well as on the east slope. In its Sierran belt, limber pine has a niche similar to whitebark pine as the upper-treeline dominant species, even forming ragged krummholz in the treeline ecotone. North of Mammoth Lakes, limber pine becomes restricted to steep, north slopes, usually of decomposed or fractured granitic rocks, whereas to the south and in other mountain ranges (with the exception of the White Mountains) it grows on diverse soil types and all aspects. In the White and Inyo Mountains limber pine is common on granitic and other noncarbonate soils of the subalpine zone. At low and middle elevations limber pine forests are often monotypic, with virtually closed canopy conditions. At higher elevations, sparse stands comprise scattered gnarled giants that can live to two thousand years. Sharp delineations generally occur between limber pine stands on granitic soils and open stands of bristlecone pine on soils of dolomite parent material. While bristlecone pines occasionally mix with limber pine, few mature limber pine stands occur on dolo590  Ecosystems

mite soils in these mountains, and bristlecones have higher upper- and lower-range boundaries. Curiously, however, limber pine seedlings have been recruiting 300 meters upslope in the late twentieth and early twenty-first centuries, above living bristlecone pine forests. Further, these upslope expansions are occurring on dolomite soils, at least in the White Mountains above Patriarch Grove and in the northern Cottonwood Canyon, at elevations and locations where no live bristlecone pine seedlings have yet established. Similar recruitment by limber pine seedlings (only) above current upper treeline is under way in granitic soils in the northern White Mountains. Limber pine forests also occur in Great Basin ranges north of the White Mountains, including the Sweetwater Mountains, Bodie Hills, and Glass Mountains. In the Sweetwaters, limber pine dominates the northern peaks, which have mafic soils of Tertiary volcanic origin, whereas whitebark pine forest is more common on the felsic (more silica-rich) soils of the southern peaks. In the Bodie Hills, limber pine is highly restricted and occurs as scattered individuals on Potato Peak and Bodie Mountain and small stands on the Brawley Peaks and Mount Hicks, just across the Nevada state line. Extensive stands occur in subalpine zones on the barren soils of Quaternary volcanic origin of the Glass Mountains and MonoInyo Craters. Limber pine forests form the treeline community in the higher Transverse and Peninsular Ranges of southern California, with relict populations at relatively low elevations on the crests of Mount Pinos, Brush Mountain, and Frazier Mountain near the junction of the Transverse and Central Coast Ranges. These last populations occur as scattered trees at elevations of approximately 2,600 meters within an open forest dominated by Jeffrey pine. The presence of a relict alpine fellfield (slope area with plant communities influenced by abiotic frost and freeze/thaw dynamics) community on the crest of Mount Pinos suggests that seasonal drought conditions and strong winds may allow the survival of limber pine (Gibson et al. 2008). Although similar in general appearance, limber pine is not closely related to whitebark pine. Like whitebark pine, however, limber pine has convergently evolved large pine nuts that rely heavily on Clark’s nutcrackers for seed dispersal (Tombach and Kramer 1980, Carsey and Tomback 1994).

FIGURE 28.6 Mountain hemlock forests favor cool, moist, often north-facing aspects, such as in Convict Canyon of the Sierra Nevada. Photo: Jeffrey Wyneken.

Unlike whitebark pine, limber pine cones open at maturity, and the seeds have a rudimentary wing. Some seeds are no doubt dispersed by wind and gravity, albeit at relatively short distances from the mother tree.

Bristlecone Pine Forests Great Basin bristlecone pine forms subalpine forests from Utah westward across the higher Great Basin ranges to the White Mountains, with scattered populations in the Inyo and Last Chance Mountains and on Telescope Peak in the Panamint Range. In the White Mountains bristlecone pine occurs largely on dolomite soils, although scattered trees may be present on sandstone and granitic soils with limber pine at elevations of 3,100 to 3,700 meters (Billings and Thompson 1957, Wright and Mooney 1965). It is a medium-size tree, typically 5–​15 meters in height and trunk diameters up to 2.5–​3.6 meters. Cones open at maturity, and the small seeds are winged and aerodynamic, although Clark’s nutcrackers also disperse bristlecone pine. Some large-diameter trees have multiple stems, potentially resulting from seeds cached by Clark’s nutcrackers (Carsey and Tombach 1994). Bristlecone pines are a remarkable species in many respects. Their most well-known feature is the great age reached by individuals, making them the oldest known nonclonal organisms. In 2012 a tree in the White Mountains was found to be 5,062 years old, making it more than two centuries older than the famous Methuselah Tree, the former record holder. Tree ages vary with slope aspect in the White Mountains. Northfacing slopes typically have the oldest trees, with an average of 2,000 years as compared to 1,000 years on south-facing slopes. The dry subalpine climate coupled with the durability of bristlecone wood can preserve them long after death, with dead trunks as old as 7,000 years scattered among living trees. The great longevity of trees and the long persistence of remnant dead wood combine to make bristlecone pine forests one of the most important scientific archives in the world for historical climate. Cross-dated tree-ring series, compiled from live and dead trees in overlapping fashion, have been developed for more than 9,000 continuous years of growth into the past. A short, several-century gap (no wood found) separates

that archive from another well-resolved 2,000-year chronology. Climate reconstructions from this 11,000-year record provide continuous proxies for annual, interannual, decadal, and centennial climate variability over the entire Holocene. Bristlecone pine stands on dolomite in the White Mountains are notable for their almost complete lack of woody understory plants—​a striking contrast to the stands of limber pine, where a number of shrub species are present. Herbaceous perennials growing on dolomite also often differ substantially in community structure from fellfield communities a few meters away on granitic soils. Similarly sharp boundaries exist between dolomite soil communities and nearby sagebrush-dominated communities on shale substrate.

Mountain Hemlock Forests Mountain hemlock forests have a broad distribution that extends from the coastal ranges of Alaska south through British Columbia and the Pacific Northwest into the Sierra Nevada. In the northern Sierra Nevada this species can be found in upper montane forests of red fir and lodgepole pine (Potter 1998) but is more characteristic at higher elevations up to 3,500 meters, where it is frequently the dominant tree species in mixed stands with Sierra juniper and whitebark pine. Mountain hemlock is locally abundant in the Klamath Mountains and the subalpine zones of Mounts Shasta and Lassen in the southern Cascades. Most of the extent of mountain hemlock forests in the Sierra Nevada occurs from Sierra County south through Yosemite National Park, with a few isolated stands reaching Fresno and Tulare County. Mountain hemlock in the Sierra Nevada is most characteristic of moist but well-drained mountain soils, often showing a preference for north-facing slopes (Figure 28.6; Fites-Kaufman et al. 2007). This contrasts with stands in the southern Cascade Range, where greater summer precipitation and warmer temperatures broaden topographic distribution (Parker 1994, 1995). Expansion of mountain hemlock in Lassen Volcanic National Park has been traced to warming temperatures as the Little Ice Age terminated in the early twentieth century, a response that might indicate the species’ behavior to continued warming in the future (Taylor 1995). Subalpine Forests   591

FIGURE 28.7 Old-growth tree on Glass Mountain. Sierra juniper forests contain trees of often massive size and growing on exposed, rocky substrates. Photo: Constance Millar.

In the central Sierra of Yosemite National Park, mountain hemlock forests can be found in extensive groves with virtually closed canopies and individual trees reaching up to 30 meters in height and 2 meters in diameter. At higher elevations mountain hemlock is more scattered and often assumes a lower, shrubby growth form (Fites-Kaufman et al. 2007). Seedlings are relatively shade-tolerant compared to other subalpine conifers and grow well under this type of canopy. South of Yosemite, mountain hemlock becomes increasingly restricted to small stands in cold moist valleys and sheltered ravines, where snowbanks remain late into the summer. Unlike pure stands of the central and northern Sierra Nevada, these scattered trees in the southern portions of the range are commonly mixed with lodgepole pine, foxtail pine, western white pine, and red fir. The southernmost occurrence of mountain hemlock is below Silliman Lake in northern Tulare County, the site of a small grove of about sixty trees with heights up to 24 meters, diameters to nearly 90 centimeters, and healthy reproduction (Parsons 1972).

Sierra Juniper Sierra juniper is one of the most striking trees of subalpine Sierra Nevada ecosystems, with its short but massive trunk appearing to grow out of seemingly solid granite substrate (Figure 28.7). It ranges through the high Sierra Nevada from south of Susanville to Owens Peak in Kern County, with scattered trees in the Inyo, White, and Panamint Mountains (Griffin and Critchfield 1976). Disjunct populations also occur in the San Gabriel and San Bernardino Mountains. Sierra juniper typically grows on shallow soils from 2,100 to 3,000 meters elevation, often with Jeffrey pine, red fir, whitebark pine, mountain hemlock, and/or lodgepole pine. More than any other subalpine tree, Sierra juniper has a remarkable ability to colonize and become established in small fractures of granite domes that would not support other species. Upper montane forests of lodgepole pine forest in the Tahoe Basin support mixed stands of Sierra juniper with red fir and Jeffrey pine, but these associated tree species are replaced by western white pine and mountain hemlock with increasing elevation (Fites-Kaufman et al. 2007). More typically, Sierra 592  Ecosystems

juniper occurs mixed in lodgepole pine stands up to treeline, where it may take on a krummholz growth form. Some Sierra junipers are reported to reach ages of over a thousand years (Graf 1999). The largest Sierra juniper—​a tree 26 meters in height and 4 meters in diameter—​is reported from the Stanislaus National Forest (Lanner 1999).

Lodgepole Pine Forests Open stands of lodgepole pine form a widespread forest belt that covers the upper montane zone and extends into the subalpine over much of California’s high mountains (Figure 28.8). The most common lodgepole pine taxon in subalpine forests is Pinus contorta subsp. murrayana, as distinguished from the Rocky Mountain lodgepole pine (P. contorta subsp. latifolia), the beach pine of the coastal Pacific Northwest (P. contorta subsp. contorta), and the local endemic Bolander pine of the pygmy forest area of Mendocino County (P. contorta subsp. bolanderi). Lodgepole pine forests extend over a very broad geographic and elevational range, including subalpine inclusions in the Klamath Mountains (including an unnamed Del Norte County variant), through the southern Cascades with populations as low as about 1,000 meters on Mount Shasta, in the northern Sierra Nevada to elevations of about 1,830–​2,400 meters, and up to 2,440–​3,350 meters in the southern Sierra Nevada. Topography strongly influences elevational distribution; lodgepole pine forests reach much lower elevations with cold air drainage down glacial canyons (Potter 1998, Fites-Kaufman et al. 2007). Lodgepole pine forests also commonly extend up into the subalpine zone in the northern and central Great Basin ranges, including the Warner Mountains, Carson Range, Sweetwater Mountains, Bodie Hills, and Glass Mountains. Disjunct colonies grow in the White Mountains, the largest of which is a nearly pure stand of approximately 100 hectares near Cabin Creek at 3,200 meters (Critchfield 1957). Larger populations appear on the San Gabriel, San Bernardino, and San Jacinto Mountains in southern California. Lodgepole pine has broad environmental tolerances, colonizing both shallow, rocky soils and semi-saturated meadow edges in an elevational belt from sea level to subalpine habitats. Only

FIGURE 28.8 Lodgepole pine forests often have narrow crowns and relatively closed canopies in the subalpine zone, as in Molybdenite Canyon, Sierra Nevada. Photo: Constance Millar.

rarely does it comprise true treeline forest ecosystems, as it is more typically replaced by whitebark pine, foxtail pine, or limber pine. The generally low stature and open stand structure of lodgepole pine subalpine forests are a function of the short growing season, associated severe climate conditions, and the thin, nutrient-poor soils that characterize the subalpine zone. These stands commonly contain few understory shrubs and little litter accumulation. Mature lodgepole pines in the subalpine zone are generally smaller than mature individuals of the dominant treeline pines and only rarely exceed 50 centimeters in diameter.

Forests of Pacific Northwest Subalpine Tree Species Three conifer species characteristic of subalpine communities of the Pacific Northwest and/or Rocky Mountains barely extend their range into California. Subalpine fir and Engelmann spruce (Picea engelmannii) are widespread in subalpine forests across western North America. The former has six known populations in the Klamath Ranges of western Siskiyou County at1,700 to 2,100 meters, while the latter is known from three populations in the lowest subalpine forests at 1,200 to 2,100 meters in the Klamath and Cascade Ranges. A third, wet-forest species from the Pacific Northwest, Alaska yellow-cedar, extends south from Alaska and barely reaches a few areas of the Klamath Mountains in Siskiyou and Del Norte Counties at elevations to 2,500 meters. While Alaska yellow-cedar is characteristically a species of cool, wet forests, its upper elevational limit extends into subalpine habitats.

Deciduous Subalpine Forests Several deciduous, broad-leaved tree species form dense local stands of subalpine forest in moist environments such as riparian corridors, meadow fringes, and upland slopes with abundant soil moisture. The most common is quaking aspen, which commonly occurs in pure groves fringing wet or moist meadows and on slopes watered by springs or seeps with subsurface water, including talus slopes (Figure 28.9; Potter 1998, Fites-Kaufman et al. 2007). Aspen is wide-

spread in appropriate habitats throughout subalpine areas of the Klamath Mountains, southern Cascade Range, Warner Mountains, Sierra Nevada, and the high mountains of southern California. Aspen is shade-intolerant and requires high light conditions to regenerate. It sprouts vigorously from suckers arising on an extensive lateral root system following fire, which plays an important role in perpetuating aspen stands by reducing competition for light from conifers. This sprouting results in a dense stand of trunks formerly assumed to be wholly clonal. More recent genetic studies reveal that some aspen groves comprise multiple genotypes (Tuskan et al. 1996). Health threats to aspen forests throughout the species range from native insects, pathogens, and incursions from conifer recruitment have heightened attention to this broadleaf ecosystem. Californian populations, however, have so far mostly been unaffected; areas of concern are concentrated in northeastern California and some parts of the northern Sierra Nevada. Notably, rapid mortality caused by sudden aspen decline (Shepperd 2008), first observed and studied in Rocky Mountain and intermountain populations, has not been reported in California (Morelli and Carr 2011). Water birch (Betula occidentalis), a widespread multistemmed small tree 6–​9 meters in height, occurs over a wide range of elevations across the western United States and Canada. In California, water birch ecosystems are common at 1,500 to 2,750 meters and grow mostly along stream corridors draining the east side of the central and southern Sierra Nevada into the Owens Valley, White Mountains, and in disjunct populations in the Klamath ranges. The species is absent from the northern Sierra Nevada and northern California. A third deciduous broadleaf tree that occasionally reaches subalpine habitats is black cottonwood (Populus trichocarpa). This tree is widespread on alluvial flats and streamsides across California up to 3,000 meters. Curl-leaf mountain mahogany, a tall evergreen shrub or small tree in the rose family, extends into the subalpine zone of California, where it can form extensive and dense canopies on dry, rocky, and exposed slopes (Brayton and Mooney 1966). Mountain mahogany has a wide distribution in subalpine zones throughout California, including the Klamath Mountains, southern Cascades, Sierra Nevada, Great Basin ranges, and high ranges of southern California. It has Subalpine Forests   593

FIGURE 28.9 Quaking aspen stands are common along watercourses, such as in Parker Canyon of the eastern Sierra Nevada, along meadow edges, or on slopes with high soil moisture. Photo: Constance Millar.

extremely hard wood and can attain ages of least 1,350 years (Schultz et al. 1990). Mountain mahogany provides browse for deer and bighorn sheep and important hiding cover from predators for these and other midsize to large mammals. Mountain mahogany ecosystems influence subalpine conifers by fixing nitrogen through associated root nodules, thereby increasing available nitrogen in otherwise nutrient-limited high-elevation soils (Lepper and Fleschner 1977).

Subalpine Meadows Meadows are scattered throughout the subalpine and montane forest zones of the Klamath Mountains, Cascade Range, Sierra Nevada, and high southern California mountains. The single most important factor explaining the distribution of meadows is the presence of a shallow water table that provides high soil moisture and excludes establishment by woody plants (Wood 1975). Although the total area of meadows is small, herbaceous plant species in meadows make up a large part of the floral diversity of subalpine zones. Meadow community composition, productivity, and biomass vary widely depending on a suite of factors. Subalpine meadows can be classified into four broad types based on vegetation composition and water table depth. These broad meadow types have been further classified based on vegetation, elevation, water table, landform, hydrology, and soil characteristics (Bennett 1965, Benedict and Major 1982, Ratliffe 1985, Allen-Diaz 1991, Sawyer et al. 2009, Stevenson 2004, Rundel et al. 2009, Weixelman et al. 2001). Wet meadows are composed predominately of perennial sedges, rushes, and grasses. Dominant species generally spread by rhizomes and often form dense sod over large areas. Soils in this type are saturated in the rooting zone for most of the growing season and are generally dark loams due to large amounts of organic material (Weixelman et al. 2001). In contrast, dry meadows are dominated by herbaceous species adapted to drier conditions, including grasses, sedges, and herbaceous dicots. Soils are not saturated within the rooting zone during the growing season, with saturation typically much deeper than the rooting zone (Allen-Diaz 1991, Weix594  Ecosystems

elman et al. 2001). In shrub meadows, open areas are interspersed with clumps of shrubs, often willow (Salix spp.) but sometimes evergreen, ericaceous shrubs such as Rhododendron columbianum (formerly Ledum glandulosum), Kalmia polifolia, and Vaccinium cespitosum (Rundel et al. 2009). Willow stands can include any of a diverse set of Salix species as dominants and occur on sites with periodic flooding during the growing season. Floods allow the ongoing establishment of willow from seed. Drier shrub meadows can have scattered but significant cover of red heather (Phyllodoce breweri), pinemat manzanita (Arctostaphylos nevadensis), the winter deciduous Utah serviceberry (Amelanchier utahensis), bitter cherry (Prunus emarginata), and California mountain ash (Sorbus californica). Woodland meadows are the fourth community type, typified by scattered sedges, grasses, and broadleaf herbs in open stands of lodgepole pine and/or aspen. Great diversity of herbaceous species occurs within this type, varying with elevation, water table, and geographic region, (Fites-Kaufman et al. 2007). The causes and dynamics of lodgepole pine establishment and survival in Sierra meadows appear to be a function of both existing water tables and climate cycles. Fluctuations in water table with interannual and interdecadal climate variability can result in cyclical lodgepole establishment, survival, and mortality (Bartolome et al. 1990, Millar et al. 2004).

Wildlife Diversity of Subalpine Forest Ecosystems Fifty native mammal species commonly use California subalpine forest ecosystems as seasonal or permanent habitat (Table 28.3; Ingles 1965, Jameson and Peeters 2004). These include a range of orders and families, including shrews, bats, rabbits, many rodents, carnivores, and ungulates. Iconic species of the upper subalpine and alpine zones include yellow-bellied marmot (Marmota flaviventris), alpine chipmunk (Neotamias alpinus), Belding’s ground squirrel (Urocitellus beldingi), American pika (Ochotona princeps), and both Sierra Nevada and desert bighorn sheep (Ovis canadensis), each of which depends on specific environments for shelter and forage. Unfortunately, all these species are challenged or thought to be at

TABLE 28.3 Mammal species that use California subalpine ecosystems as habitat (Ingles 1965; Jameson and Peeters 2004)

Order

Family

Insectivora

Soricidae

Species Sorex

palustris lyelli monticolus

Chiroptera

Vespertilionidae

Myotis

lucifugus

Lagomorpha

Ochotonidae

Ochotona

princeps

Leporidae

Lepus

americanus californicus townsendii

Rodentia

Aplodontidae

Aplodontia

rufa

Sciuridae

Marmota

flaviventris

Tamias

alpinus amoenus minimus quadrimaculatus speciosus umbrinus

Geomyidae

Callospermophilus

lateralis

Otospermophilus

beecheyi

Urocitellus

beldingii

Tamiasciurus

douglasii

Thomomys

bottae mazama monticola

Cricetidae subf Cricetinae

subf Microtonae

Reithrodontomys

megalotis

Neotoma

cinerea

Peromyscus

maniculatus

Clethrionomys

californicus

Microtus

longicaudus montanus oregoni

Zapodidae

Phenacomys

intermedius

Zapus

princeps trinotatus

Carnivora

Erethizontidae

Erethizon

dorsatum

Canidae

Canis

latrans

Vulpes

vulpes

Felis

rufus

Puma

concolor

Felidae

(continued)

TA B L E 2 8 . 3 (continued)

Order

Family Mustelidae

Species Gulo

gulo

Martes

americana pennanti

Mephitis

mephitis

Mustela

erminea frenata

Taxidea

taxus

Bassariscus

astutus

Procyon

lotor

Ursidae

Ursus

americanus

Cervidae

Odocoileus

hemionus

Ovis

canadensis

Procyonidae

Artiodactyla

risk from various human stressors. Most common are declines or impacts associated with contemporary climate change (Moritz et al. 2008). Marmot, alpine chipmunk (Rubidge et al. 2012), bushy-tailed packrat (Neotoma cinerea; Moritz et al. 2008), and Belding’s ground squirrel (Morelli et al. 2012) have experienced changes in distribution and population dynamics from warming temperatures and changing snowpacks. American pika has long been considered at risk from changing climate and appears threatened in the central Great Basin. California populations, however, appear to be more buffered against change (CDFW 2013). Sierra Nevada bighorn sheep, a distinct subspecies in the central and southern Sierra Nevada, suffered drastic population declines over the twentieth century. The species was placed on the federal Endangered Species list in 2000 when its numbers declined to near one hundred. Implementation of formal recovery plans has led to recovery toward the goal of five hundred adults distributed throughout historical herd units (Stephenson et al. 2011). Avian species that depend on subalpine forests include mountain bluebird (Sialia currucoides), red crossbill (Loxia curvirostra), pine grosbeak (Pinicola enucleator), Cassin’s finch (Carpodacus cassinii), Williamson’s sapsucker (Sphyrapicus thyroideus), black-backed woodpecker (Picoides arcticus) (Mayer and Laudenslayer 1988), and Clark’s nutcracker (Meyer 2013). The dependence of whitebark pine on Clark’s nutcracker for reproduction is a remarkable example of coadaptation among species in subalpine forests. Seed cones of whitebark pine are unique among the pines in having indehiscent bracts—​t hat is, the cones do not break apart on their own when mature (Figure 28.10). Further, cones remain closed and tightly adhered to the stem even at seed maturity. Cones can only be broken open and seed released by Clark’s nutcrackers, which in turn depend on whitebark pine seeds for food (Tomback 1982, 1986). The birds open the cones while on the tree, carry batches of seeds in specialized pouches under their bills, and plant seeds in caches, usually in protected locations. A single Clark’s Nutcracker caches as many as ninety-eight thousand 596  Ecosystems

seeds per season (Hutchins and Lanner 1982). The birds have uncanny ability to relocate their caches, even under a meter of snow, and they utilize these caches to feed young birds through early growth and development. Sufficient seeds are left unrecovered by Clark’s nutcrackers that whitebark pine seedlings can germinate (Lanner 1996). Clark’s nutcrackers also use limber, foxtail, bristlecone, and western white pine. During migrations to lower altitudes, the birds also extensively harvest the seeds of pinyon pines. Like Clark’s nutcracker, several other bird species and small mammals serve important ecological roles for subalpine tree species. Douglas’s squirrel (Tamiasciurus douglasii), lodgepole chipmunk (Neotamias speciosus), and other seed-caching wildlife species are important seed dispersers and predators of subalpine tree species in subalpine ecosystems (Tomback 1982, Van Der Wall 2008).

Origins of Subalpine Forest Species and Ecosystems The biogeographic origins of California’s present-day subalpine forest species and forest communities are highly complex given the geologic uplift and subsidence history, diversity of historical climates, and influences on vegetation of multiple periods of dramatic climate change that characterize the region (see Chapter 8, “Ecosystems Past: Vegetation Prehistory”; Millar 1996, 2012). The biogeography of most present-day species extends back more than twenty million years. However, species’ locations, environmental contexts, elevations, climates, and vegetation associations have varied drastically over time. Many cool-temperate species, including subalpine conifers and other mountain-adapted plant species still in the region, found refugial habitat in the expansive Nevadaplano uplands during a long period when temperate zone climates elsewhere throughout North America were subtropical. The pre–​Sierra Nevada ranges did not form a hydrologic divide as the Sierra Nevada do now, nor were they the

A

B

FIGURE 28.10 Coadaptations between (A) whitebark pine cone and (B) Clark’s nutcracker ensure that the birds have sustenance and that pines are planted. Photos: Constance Millar.

highest summits of this expansive upland region. The elevations of many ranges within this Nevadaplano, including the pre-Sierran mountains, had summits estimated to extend more than 3,000 meters. Fossil flora that date to the middle Tertiary from this region, now located mostly in Nevada but indicative of California, include a great diversity of gymnosperm and angiosperm taxa, including subalpine species with affiliations to bristlecone pine, foxtail pine, lodgepole pine, western white pine, Alaska yellow-cedar, and mountain hemlock (Millar 1996). These floras, however, do not reflect ecological stratification as do present upland communities, as they included in single associations a mix of diverse montane-adapted species as well as conifer taxa now occurring only in lower montane and coastal types. In addition, many summer/wet-adapted angiosperm species that grow now in southeast North America and even in subtropical climates co-occurred in these upland sites. The diversity and lack of zonation of these Tertiary floras is interpreted to indicate, despite the high elevations, that climates were relatively warm and wet with precipitation distributed year-round. Subsequent changing dynamics of plate boundaries and new tectonic activity associated with the development of the San Andreas, Southern California Shear, and Walker Lane fault zones initiated massive changes to the topography and climate of the region. These in turn triggered drastic changes in the vegetation of the mountains of California and Nevada. By about ten million years ago, extensional forces led to the development of the Great Basin, with its more than three hundred fault-blocked mountains and basins and internal drainage, as well as to development of the present-day Sierra Nevada with its increasing significance as a major hydrologic divide. These tectonic changes catalyzed orographic rainshadow effects leading to vegetation zonation both in elevation, introducing modern subalpine ecosystems, and from the Pacific coast inland (i.e., longitudinally). The California Mediterranean-climate regime grew in dominance, exerting strong selection pressures for traits enhancing survival of increas-

ingly long, dry summers. Truly arid environments (desert) and alpine ecosystems began to emerge for the first this time in the California/Nevada region as well. These changes triggered major extirpations of summer/wet-adapted species. Many montane conifers and cool/mesic-adapted plant species that had lived for millennia on the Nevadaplano persisted in the present-day Sierra Nevada and in northwest California while disappearing from inland Great Basin regions as climates dried. The present-day diversity of subalpine conifers was in place in the California high mountain regions at the onset of the Quaternary, two million years ago (Millar and Woolfenden 1999). The roller-coaster climate changes that ensued brought as many as forty cycles of cold glacial and warm interglacial conditions to North America, with maximum temperature differences in the western mountains of as much as 8°C to 10°C (Millar 2012). Abrupt and gradual changes in climates catalyzed significant movements of subalpine conifers—​ downslope to 2,600 meters in the Sierra Nevada during glacials when ice caps covered the mountains and upslope during interglacials such as the present Holocene. The maximum altitudinal response of subalpine species to glacial-interglacial cycles was approximately 1,000 meters. During the warmest intervals within interglacials, such as the mid-Holocene interval (four thousand to eight thousand years ago), treeline ascended about 100 meters higher than at present in the Sierra Nevada (Anderson 1990, Hallett and Anderson 2010) and as much as 150 meters in the White Mountains (LaMarche 1973). Compositions and associations of subalpine communities also changed between glacials and interglacials with changing taxonomic diversities, species abundances, population expansion and contractions (into important refugial areas), and changes in disturbance regimes (e.g., Mohr et al. 2000, Anderson 1990, Hallett and Anderson 2010) but with no apparent extirpations or extinctions (Millar and Woolfenden 1999). During the latter period of the last glacial period, 13,000 to 11,500 years ago, giant sequoia (Sequiadendron giganteum) moved far Subalpine Forests   597

above its current range in the Sierra Nevada as recorded in a sediment core at 2,863 meters from East Lake. Its floristic associates included subalpine taxa and indicated that giant sequoia was part of the subalpine ecosystem at the time (Power 1998). Changes in the fire regimes of mountain ecosystems across fluctuating glacials and interglacials mirrored changes in mountain climate and vegetation composition and structure (Skinner and Chang 1996). Where fire history studies have been done in the subalpine zones of California, they show increasing fire severity and extent when climate was wet enough to support dense forest growth, and reduced fire effects during dry intervals—​b oth cold and warm—​ when subalpine forests became sparser. In the central Sierra Nevada near Mammoth Lakes during the end of the last glacial period, when climates were cold and dry and subalpine pine and mountain chaparral species remained sparse as they shifted upslope into deglaciated areas, fires were few and low severity (Hallett and Anderson 2010). After eight thousand years ago, as mountain hemlock moved into the region and dense forests formed, fire became more frequent and intense. During the warm and dry mid-Holocene, subalpine forests again became sparser and fire was less important. During the neoglacial dry periods starting approximately four thousand years ago, fire was also of minor importance in the subalpine forests. After about twelve hundred years ago, fires in the subalpine forests became increasingly synchronized with inferred drought, and fire activity in the high Sierra is interpreted to have been highly sensitive to dynamics of the El Niño–​Southern Oscillation (Hallett and Anderson 2010). In the Klamath Mountains fire history reconstructed over the past 15,500 years shows a slightly different pattern of changes in forest composition and density and fire relationships (Mohr et al. 2000). Before the end of the last glacial period, subalpine forests were open and parklike, dominated by scattered pines and firs and marked by low fire frequencies. Forest density increased in the latest Pleistocene and shifted to western white pine, lodgepole, and firs, yet fire frequency remained low during a period interpreted as cold and wet. During the middle Holocene, conditions became warm and dry with increasing fire frequencies. Similar, high frequencies were inferred at the onset of the neoglacial period four thousand years ago as hemlock increased in abundance, displacing pines and oaks. Elsewhere in the Klamath region, fire-scar records of the last four hundred years in the Scott Mountains show that fire return intervals ranged from one to seventy-six years with averages of about seven years (Skinner 2003). Most fires, however, scarred only one sampled tree, suggesting that although fires were frequent in subalpine forests, most were probably small and low-intensity. In the southern Cascades, Bekker and Taylor (2001) found long fire return intervals over the past 350 years in subalpine forests of the Thousand Lake Wilderness. Fire regimes varied with forest composition, elevation, and inferred soil moisture, and fire return intervals ranged from twenty to thirty-seven years for lodgepole pine forests and twenty to forty-seven years for mountain hemlock types. Fires occurred mostly in the late growing season and dormant season (Bekker and Taylor 2001). Similar patterns marked historical western white pine subalpine forests over the last 350 years at Mount Lassen (Taylor 2001). In both regions of the southern Cascades, marked declines in fire frequency took place in the twentieth century. 598  Ecosystems

Ecosystem Dynamics Wildfire By and large, fire is less important in subalpine forests than in forests at lower elevations. Landscape-scale fires are rare because high-elevation landscapes form a mosaic of individual trees; tree stands; upland shrub and herbaceous communities including meadows, wetlands, and riparian corridors; rock outcrops; talus; avalanche tracks; creeks; and lakes. Fuel buildup is usually slight in these extreme conditions with short growing seasons, limiting opportunities for fire to spread. In open whitebark pine woodlands and even krummholz communities, lightning ignitions can cause single trees to explode and burn, but fires that do start from these points tend to smolder at ground level and extend only to the edges of patches of fuel accumulation. Where stand densities increase to the point of canopy closure, however, crown fires and fires of high intensity can occur in subalpine ecosystems. Even in these forests, such as lodgepole pine and mountain hemlock types, deep snowpacks and saturated soils from spring snowmelt usually restrict fires to the late growing season or dormant season (Skinner et al. 2006). On some shallow substrates with poor fertility, as in the Klamath Mountains and southern Cascade Range, fire behavior influences the persistence and dominance of montane shrub communities (e.g., Arctostaphylos nevadensis, Chrysolepis sempervirens, Quercus vaccinifolia, and Ceanothus spp.). Once trees are removed by crown fire and shrub species regenerate into burns, they can claim dominance over time because their resinous, highly flammable canopies increase fire frequencies and they sprout or seed into burned areas more successfully and rapidly than conifers (Pinder et al. 1997). Where investigated, fire regimes in the denser subalpine forest types were characterized by long return intervals in the presettlement period (1700–​1800s) (van de Water and Safford 2011). The longest fire interval reported, 133 years, is for typical subalpine forest types that include whitebark, bristlecone, limber and/or foxtail pine, and mixed stands containing those species plus western white pine, lodgepole pine, and mountain hemlock. Fire return interval for Sierra juniper forests, which almost always occur as sparse stands on rocky substrates, was eighty-three years. Western white pine and curl-leaf mountain mahogany forest fire return intervals were about fifty years, and lodgepole pine forest intervals were thirty-seven years. Somewhat unexpectedly given their high soil moistures, fire return intervals for aspen were shortest of all subalpine types at nineteen years (van de Water and Safford 2011). Presettlement spatial patterns of fires in subalpine environments are difficult to assess and, except at the lower ecosystem border or in special conditions, mostly influence local structure and composition (Caprio and Graber 2000, Meyer 2013). Mean fire size in the southern Cascade subalpine elevations was estimated as 405 hectares for lodgepole pine forests and 140 hectares for red-fir/mountain hemlock, with mean size at Mount Lassen of 176 hectares. Studies in the Tahoe Basin indicate mostly small and patchy presettlement fires, with evidence that some areas burned severely enough to produce even-aged cohorts (Scholl and Taylor 2006). During the later twentieth and twenty-first century, observations of uncontrolled wildfires in subalpine forests of California indicate much smaller spatial extent than these presettlement estimates, with most less than 4 hectares (Meyer 2013). Even with a slight increase in fire size observed during the first

decade of the twenty-first century (Miller et al. 2009), the small sizes and long return intervals of fire in subalpine forests underscore the minor and local effects that fires have on controlling ecosystem structure and function.

Insects and Pathogens Cold temperatures, low humidities, rocky environments, and wide spacing of trees limit insects and disease-causing organisms to minor roles in subalpine forests. Although insects that damage or kill trees (e.g., defoliators and bark beetles) occur in forests at those elevations, they rarely reach outbreak conditions (but see Brunelle et al. 2008 for the Rocky Mountains). Their effects, as observed over the past century, have been to influence background mortality. In the latest twentieth and early twenty-first century, however, tree mortality related to insect and disease outbreaks appears to have vastly increased in subalpine forests of western North America. Investigation of these changes in bark beetle activities point to warming temperatures, which allow the insects to overwinter and in some locations to complete two generations each year (Logan and Powell 2001, 2007). In association with periodic drought, and in conditions where soil moisture stress is high, bark beetle outbreaks, primarily mountain pine beetle (Dendroctonus ponderosae) on whitebark pine, have been at record high levels. California’s subalpine forests have so far resisted landscape-scale mortality from bark beetles (Millar et al. 2012). They have experienced, however, population-scale outbreaks and mortality events in the late twentieth century and early twenty-first century on limber pine (Pinus flexilis) and whitebark pine (Millar, Westfall et al. 2007; Millar et al. 2012). Increasing background temperatures, multiyear drought, and low soil moisture (low CWD) are implicated in both cases. In both outbreaks only stands at low elevations for each species’ range, northerly aspects, and young, dense, fast-growing stands were affected. Further, the degree of forest mortality increased latitudinally, from the southern Sierra Nevada to the Warner Mountains, reflecting improved conditions for insects as precipitation and stand densities increased (Millar et al. 2012). Limitations to spread of beetle outbreak in California are likely related to endogenous factors (insect competitor and prey relations) but also to environmental context of the forests. California’s historically warm and dry Mediterranean climate both preadapts forests to arid conditions and influences bark beetle behavior in ways quite different from situations in the Pacific Northwest and Rocky Mountains (Bentz et al. 2014). In Californian beetle populations, due to historic adaptation to warm, dry climates, the trend toward bivoltinism (two generations per year) is less common, and will require greater temperature increases to evolve, than elsewhere. The major disease-causing species in current subalpine forests of western North America is the invasive white pine blister rust (WPBR), caused by the fungus Cronartium ribicola. This fungus was introduced on nursery stock more than one hundred years ago and has since been spreading on fiveneedled pines (subgenus Strobus) throughout North American forests (Smith 1996). The high-elevation pine species are highly susceptible, but remote locations and unfavorable climate conditions have until recently limited effects on subalpine forests. In recent decades, WPBR has invaded subalpine forests in the Cascade Range, inland mountains, and

Rocky Mountains. The rust has caused extensive mortality on whitebark pine, and in 2011 the U.S. Fish and Wildlife Service granted the species protection under the Endangered Species Act (USFWS 2011). California subalpine forests have so far experienced only localized mortality from WPBR (Maloney and Dunlap 2007, Dunlap 2012) in some parts of the western Sierra Nevada, Mount Rose, and at scattered, low levels elsewhere. WPBR has caused extensive damage to lower-elevation white pine species, especially sugar pine and western white pine. A suite of other insects and pathogens cause minor damage and localized mortality on subalpine forests. Dwarf mistletoes (Arceuthobium spp.), a group of vascular plants that live aerially on conifer branches and stems, cause branch death and brooming on the subalpine pines and many other species (Hawksworth et al. 1996). Dwarf mistletoe is not common in subalpine forests but does occur on limber pine forests of the eastern Sierra Nevada (Millar, Westfall et al. 2007) and whitebark pine forests in northern California, where it appears to exacerbate outbreaks of mountain pine beetle. In the first decade of the 2000s, occasional individual or clumps of bristlecone pine trees in the subalpine forests of the White Mountains were observed dying or experiencing branch dieback. These sporadic effects seem most likely to be caused by black stain fungus (Leptographium wageneri), a native species not known or expected to cause widespread damage to the bristlecone pine forests (B. Bulaon, U.S. Forest Service, pers. comm.). No mountain pine beetle incidences have been known in bristlecone pine forests of California to date. Native red turpentine beetle (Dendroctonus valens) has been found on bristlecone pine in the White Mountains but is mostly a secondary invader that forages on dying trees and dead wood.

Biogeochemical Cycling and Hydrology The biologically rich subalpine and alpine basins of the Sierra Nevada have had a long history of descriptive study, with much of what is known about the hydrology and biogeochemical cycling coming from long-term studies of the subalpine Emerald Lake basin in Sequoia National Park (see Chapter 32, “Lakes”). Many of the hydrologic studies in the Sierra Nevada centered on assessing and monitoring patterns of seasonal and interannual change across large elevation gradients from subalpine and alpine watersheds to the San Joaquin Valley. Recent studies have shown that the fluctuations of selected rivers draining the Sierra Nevada and Rocky Mountains are highly correlated each spring, indicating an organized, regional-scale signal of snowmelt initiation and runoff (Cayan et al. 2001). These basins integrate the effects of broad ranges of aspect and elevation. Hydrologic studies in subalpine watersheds have focused on understanding streamflows, water balance, and the association of snowmelt and runoff with solute chemistry (Kattelmann and Elder 1991, Williams and Melack 1991 Meixner and Bales 2003, Meixner et al. 2004). Other studies have quantified the components of old water stored in the watershed from the previous year (10–​20%) compared to new water from current snowmelt (80–​100%) (Huth et al. 2004). The Emerald Lake watershed provides a case study of biogeochemical cycles of nitrogen in a subalpine watershed (Tonnessen 2001). The soil pool of organic nitrogen was about ten times nitrogen storage in litter and biomass, and assimilation by vegetation was balanced by the release of nitrogen from litter decay, soil mineralization, and nitrification (Williams et al. Subalpine Forests   599

TA B L E 2 8 .4 Biomass of plant material in the 120 hectare Emerald Lake watershed, Sequoia National Park

Basin coverage (ha)

Aboveground biomass (kg ha-1)

Belowground biomass

Willow

8.55

965

567

Mesic shrub

0.73

26.5

42.4

29.5

Mesic crevice

15.15

36.9

58.8

40.0

Wet meadow

4.14

Xeric crevice

13.40

Dry meadow

Community type

439

999

47.3

6.0

33

48.7

7.73

62.9

361

49.7

Fellfield

0.84

0.4

2.1

0.3

Colluvium

3.44

1.5

8.5

1.2

Trees

ND

Total

130

Litter biomass

16,000

5,630

17,200

7710

37.7 694

Source: Adapted from Rundel et al. 1989. See Rundel et al. 2009 for a description of communities.

1995, Wolford et al. 1996, Wolford and Bales 1996, Meixner and Bales 2003, Sickman et al. 2001, 2003). Trees make up about 90% of the aboveground biomass and nitrogen in plant tissue in the watershed, but only about three-fourths of belowground biomass and half of belowground nitrogen (Table 28.4). As much as 90% of annual wet deposition of nitrogen was stored in the seasonal snowpack, and both nitrate and ammonium ions were released in a strong ionic pulse with the first fraction of spring snowmelt. This nitrate release was evident in a small but significant pulse in streamwater concentration with early snowmelt. However, almost all of the ammonium input from both wet and dry deposition was retained in the watershed by biological assimilation (Williams et al. 1995).

Ecosystem Services Snowpack and Water Supply By far the greatest ecosystem service provided by high mountain watersheds is their capacity for storage and delivery of critical water supply for downstream agricultural, industrial, urban use. Subalpine hydrologic reserves, stored primarily as winter snowpack, provide water both to the west in California and to the increasingly urbanized eastern front of the Sierra Nevada/Carson Ranges. Snowpack retention is an extremely important function of subalpine forests given the long Mediterranean summer drought regime of the California climate. Density of trees and canopy cover are important determinants in the amount of snow that develops on the forest floor and also on the retention into spring (Raleigh et al. 2013, Lundquist et al. 2014). Although many subalpine forests are sparse, those with moderate density provide the greatest snowpack retention of all forest types. In addition to forest canopy and density, many factors affect the stability and hydrology of high-elevation environments. Subalpine ecosystems are sensitive to small changes 60 0  Ecosystems

in growing season conditions of temperature and water availability, and their stability impacts hydrological conditions of lowland ecosystems (Bales et al. 2006, Trujillo et al. 2012). Given the high interannual variability of precipitation in the California region, and corollary differences in snowpack depth, water supplies in the critical spring forecast season vary drastically (Figure 28.11). In that precipitation and snowpack vary greatly from year to year as a result of natural forcing (McCabe and Dettinger 1999), further impacts on these sensitive functions have large cascading impacts. The most significant effect on snowpack and water supply is from warming temperatures as a result of anthropogenic forcing on climate (Bonfils et al. 2008). With snowpack amount declining and snowmelt advancing, the ability to provide water to increasingly hot and arid urban users during a prolonging summer drought will be severely challenged (Stewart et al. 2005). The many skeletal and oligotrophic subalpine watersheds of the Sierra Nevada also have the potential to be strongly affected by atmospheric nitrogen deposition and acidification associated with the expanding urbanization of the San Joaquin Valley (Sickman, Leydecker et al. 2003).

Biodiversity Subalpine ecosystems in California support important and distinct biodiversity. In addition to the tree diversity of subalpine forest ecosystems, many thousands of vascular and nonvascular plant species grow in the diverse habitats of California’s high elevations. The Jepson eFlora lists taxa by floristic subprovinces, from which estimates for the number of plant taxa in several subalpine regions can be derived. Of more than 6,500 plant taxa in California as a whole, these include 1,710 species in the high North Coast Ranges, 1,860 species in the high Cascades, and 2,740 species in the Sierra Nevada (Jepson eFlora 2014). Although these numbers are only rough estimation of actual subalpine zone diversity, they

A

B

Percent 1971–2000 Average > 180 150–180 130–149 110–129 90–109 70–89 50–69 25–49 180 150–180 130–149 110–129 90–109 70–89 50–69 25–49 3 m) Timberline (timber-size trees)

F

F

G

Example of a species specific elevational limit of a non-treeline-forming tree species

F I GU R E 2 9.1 Schematic representation of the subalpine-alpine ecotone, showing various ways in which the thermal treeline can be displaced downslope by disturbance, snowbeds, or substrate. The alpine-nival ecotone is shown high on the peaks. Between these two ecotones is the alpine zone. Source: Modified from Körner 2012.

alpine ecosystems have excursions upward and downward for reasons including geology, geomorphology, and microclimate, to classify California alpine ecosystems at the regional scale, we adopt the global thermal limit for treeline as a convenient low-elevation baseline. Treeline has been defined as the zone on the landscape where average growing season temperature is 6.4°C or lower (Körner and Paulsen 2004). In this context, trees are defined as plants having upright stems that attain height ≥3 meters regardless of taxonomy, and the treeline community is characterized as more-or-less continuous patches of trees whose crowns form at least a loose canopy (Körner 2007). Typically, the upper-elevational zones around treeline support mosaics of subalpine forest, dwarfed or krummholz trees, shrublands, and low alpine perennials, each responding to a complex mix of environmental conditions, soil, and disturbance history (Figure 29.2). Environmental stresses associated with temperate alpine ecosystems include extreme winter temperatures, short growing season, low nutrient availability, high winds, low partial pressures of CO2, high UV irradiance, and limited water availability (Billings 2000, Bowman and Seastedt 2001, Körner 2003). The California alpine ecosystem lies in regions colder, and usually above, this zone. For the purpose of this chapter, we allow alpine ecosystems to include also the small area of nival regions that occurs in the state. The latter is defined by another globally occurring threshold: the zone where average growing season temperatures are ≤3°C, below which even plants shorter than 3 meters cannot endure freezing damage (Körner 2007). Only hardy lichens, isolated plants in protected microenvironments, snow algae, and other snow ice–​dwelling invertebrates can survive these conditions. The definition of alpine Photo on previous page: Granitic cirque at head of Burt Canyon, Sierra Nevada, with wetlands along the creek and aster- and paintbrush-festooned upland slopes. Photo: Jeffrey Wyneken. 614  Ecosystems

we adopt differs from what is often generically called tundra. Technically, “tundra” refers to specific vegetation formations as well as regions of permanently frozen soils and usually is applied to Arctic latitudes (Billings 1973). Alpine ecosystems extend beyond the typically envisioned high-elevation open slopes and summits of cold-adapted shrubs and herbs to also include lithic environments of cliffs, talus fields, boulder fields and rock glaciers; permanent and persistent snow and icefields, including glaciers; and various water bodies such as streams, tarns, and large lakes.

Geographic Distribution of Alpine Ecosystems in California The lower (warm) limit of the alpine ecosystem, or climatic treeline, varies with latitude across California, ranging from 3,500 meters in the southern California mountains, to 3,200 meters in the Yosemite region, to 3,000 meters near Donner Pass, to 2,800 meters with somewhat higher elevations in ranges of the western Great Basin east of the Sierra Nevada (Rundel 2011) (Figure 29.3). To the north, in the Cascade Range, climatic treeline begins at approximately 2,800 meters on Lassen Peak and 2,700 meters on Mount Shasta. Moving from south to north, high elevations with alpine communities are first encountered in the Transverse and Peninsular Ranges of southern California. These ranges support local areas of weakly developed, alpine-like communities populated by a subset of Sierran alpine species (Hall 1902, Parish 1917, Horton 1960, Hanes 1976, Major and Taylor 1977, Meyers 1978, Gibson et al. 2008). Mount San Gorgonio in the San Bernardino Mountains reaches 3,506 meters and had local glacial activity in the Pleistocene (Sharp et al. 1959). Other high points are Mount San Jacinto in the San Jacinto Mountains at 3,302 meters and Mount Baldy (San Antonio) in the San Gabriel Mountains at 3,068 meters. Alpine species

FIGURE 2 9.2 The forest-alpine ecotone at upper treeline is a zone on the landscape that can include krummholz whitebark pines, as here on the uplifted, metamorphic plateau of the Tamarack Crest, Sierra Nevada. Photo: Constance Millar.

are present in both xeric and mesic habitats at high elevations, but alpine communities in the form of extended areas dominated by assemblages of alpine species are only weakly developed. The greatest area of alpine ecosystems in California occurs in the Sierra Nevada. The elevational contour of 3,500 meters, a limit that roughly corresponds to treeline in the southern Sierra Nevada, has been used as one simple parameter to delineate alpine ecosystems (Sharsmith 1940). This boundary defines a relatively continuous area from Kings Canyon and Sequoia National Parks along the crest of the central and southern Sierra Nevada extending to northern Tuolumne and Mono Counties. The alpine zone of the southern Sierra Nevada first appears on Olancha Peak (3,698 meters) on the Tulare-Inyo County line, the southernmost glaciated summit of the range (Howell 1951, Tatum 1979). Cirque Peak (3,932 meters) in Sequoia National Park forms the southern limit of an extensive and virtually contiguous alpine zone of glaciated peaks in the Sierra Nevada. Here occur extensive areas of alpine habitat and high peaks that reach above 4,000 meters, with Mount Whitney at 4,421 meters the highest point in the contiguous United States. Alpine habitats in the central Sierra Nevada are well developed in the area of Leavitt Peak (3,527 meters) near Sonora Pass and south across Yosemite National Park, whose highest peak is Mount Lyell (3,999 meters). Further south, this belt of alpine habitat continues into Kings Canyon and Sequoia National Parks. Tioga Pass in Yosemite National Park (3,031 meters) and Mammoth Pass (Minaret Summit,2,824 meters), which is the route for California Highway 203, provide two major breaks containing subalpine elevations but not true alpine habitats. North of the Tioga Pass area, the crest of the Sierra Nevada lies at lower elevations with only scattered areas of typical alpine habitat present. Fragmented communities of alpine species are present at elevations well below 3,500 meters, particularly along exposed ridgelines and on steep, north-facing slopes that were once heavily glaciated. Alpine habitats are weakly developed in Alpine County (Sonora Peak, 3,493 meters) and eastern El Dorado County (Freel Peak, 3,318 meters), extending to their northern limit on Mount Rose (3,285 meters) in the Carson Range east of Lake Tahoe along the California-Nevada border. Nevertheless, scattered com-

munities of alpine-like habitat exist at upper elevations in the northern Sierra Nevada, positioned above and around local, edaphically controlled treelines, and an alpine flora is well represented (Smiley 1915). The substrate north of Sonora Pass is largely volcanic and thus quite distinct from the granitic bedrock of the central and southern Sierra Nevada. Notable exceptions exist where granitic plutons are exposed in the Desolation Wilderness, Donner Pass region, and adjacent parts of the Tahoe Basin. Several high mountain ranges lie to the east of the Sierra Nevada at the western margin of the Great Basin, and these support small but diverse areas of alpine ecosystems. The White Mountains have an extensive alpine area, with 106 square kilometers above 3,500 meters and the third highest peak in California, on White Mountain Peak, at 4,344 meters (Figure 29.4). To the south, Mount Waucoba forms the high point in the Inyo Mountains at 3,390 meters. The Panamint Mountains east of the Inyo Mountains reach a maximum elevation of 3,366 meters on Telescope Peak. The Sweetwater Mountains, located 33 kilometers east of the north-central Sierra Nevada north of Bridgeport, reach 3,552 meters on Mount Patterson and contain a significant zone of diverse alpine ecosystems (Bell-Hunter and Johnson 1983; Figure 29.5). At the south end of the Mono Basin, volcanic Glass Mountain reaches 3,392 meters and supports a small, mostly edaphically controlled alpine zone. To the north of the Sierra Nevada, the southern Cascade Mountains provide local areas of alpine habitat and were likely once stepping-stones for high-elevation plants and animals migrating into California with the late Neogene and Pleistocene tectonic dynamics of the Sierra Nevada (see Chapter 8, “Ecosystems Past: Vegetation Prehistory”). Mount Shasta reaches an elevation of 4,322 meters, while Lassen Peak extends to 3,187 meters (Gillett et al. 1995). Magee Peak (2,641 meters), located midway between Mounts Lassen and Shasta, supports limited areas of alpine vegetation in cirques on its north face (Major and Taylor 1977). In northeast California, the Warner Mountains, another fault-block range of the Great Basin province, run 140 kilometers from south to north and attain plateau heights near 3,000 meters. Alpine ecosystems are scattered along their crest, primarily in the South Warner section. Alpine Ecosystems   615

FIGURE 2 9.3 Distribution of alpine ecosystems in California. Source: Data from U.S. Geological Survey, Gap Analysis Program (GAP); and Cal Fire, Fire Resource and Assessment Program (FRAP). Map: Parker Welch, Center for Integrated Spatial Research (CISR).

616  Ecosystems

FIGURE 2 9.4 Extensive alpine fellfields extend across the broad alpine plateaus of the White Mountains, interrupted occasionally by inselbergs, patterned ground, and other periglacial features. Photo: Constance Millar.

FIGURE 2 9.5 The diverse volcanic soils of the Sweetwater Mountains in the California Great Basin appear barren from a distance but support a great number of alpine herbs, including many endemics. Photo: Constance Millar.

In northwestern California the higher peaks of the Klamath Mountains, especially in the Trinity Alps, Marble Mountains, and Scott Mountains, support alpine ecosystems and contain areas of permanent or long-lasting snowfields on north-facing slopes (Howell 1944, Major and Taylor 1977). The highest peaks are Mount Eddy (2,750 meters) in Siskiyou County, Thompson Peak (2,744 meters) in Trinity County, and Mount Ashland (2,296 meters) in Jackson County, Oregon. Major and Taylor (1977) note that alpine species distributions dip as low as 2,000 meters or less on karst topography associated with marble substrates in the Marble Mountains, and alpine-like communities also have lower-elevation excursions onto ultramafic (e.g., serpentine) soils (Kruckeberg 1984).

Climate Regimes and Abiotic Stress At the regional or synoptic scale, the alpine zone of the Sierra Nevada experiences a Mediterranean-type climate regime with dry summers and with precipitation heavily centered on the winter months. This regime differs significantly from that

present in the more continental and monsoon-dominated regions of western and southwestern U.S. and in most of the continental alpine habitats of the world, where summer precipitation predominates. This seasonality is a significant element of the alpine environments of California and a strong factor in explaining the relatively high rate of endemism in the alpine flora. Much of the historic literature on plant and animal adaptations to high-elevation, alpine habitats has come from studies in the colder and more continental alpine ecosystems of the Rocky Mountains and the European Alps. In alpine habitats at the upper treeline in the Sierra Nevada, about 95% of annual precipitation falls as winter snow. By contrast to the Pacific Northwest, where snowpacks accumulate during regular winter storms throughout the cold season, much snow in the California alpine zones falls during a very small number of storms separated by long, dry intervals. The absence of one or two of these storm episodes in a year can cause a dry snowpack year relative to average levels. By contrast, fortuitous landing of one or more atmospheric river storms (“pineapple express”) on the California region can result in record wet years and deep snowpacks (Dettinger et al. 2011). Deep snowpacks and cool temperatures at higher elevations mean that snowmelt extends into spring, but the length and magnitude of the summer drought period experienced by plants and animals is significant. Patterns of rainfall decline gradually from north to south in the main California cordillera, and summer drought decreases as elevation increases because of both increased levels of precipitation and cooler temperatures with lower evaporative demand at higher elevations (Stephenson 1998, Urban et al. 2000). Very few climate stations are situated in the California alpine zone. In general, winter mean monthly low temperatures are moderate in the Sierra Nevada (Table 29.1) compared to the more continental climates of the interior Great Basin and Rocky Mountains, and in general soils rarely freeze beyond moderate depths. While the mean minimum temperature above treeline is generally below freezing for ten months of the year, nighttime lows typically reach only –3°C to –6°C, although temperature extremes can fall below –​ 20°C on high, north slopes and in cold air sinks (Millar et Alpine Ecosystems   617

TA B L E 2 9.1 Climate data for select alpine locations in California

Temperature °C Jan

July

Annual

Precipi­tation (mm)

Location

Mountain range

Latitude (° W)

Mount Shasta

South Cascades

41.42997

122.19914

3261

-2.4

-10.9

15.2

0.8

4.9

-6.3

-0.7

275

Thompson Peak

Trinity Alps

41.00065

123.04834

2717

0.9

-6.3

18.6

6.7

8.3

-1.0

3.7

Mount Lassen

South Cascades

40.48916

121.50747

3130

-0.6

-9.1

18.0

3.0

7.5

-4.4

Mount Rose

Sierra Nevada

39.34403

119.91807

3283

-0.4

-11.0

17.6

3.2

7.0

Barcroft Station

White Mountains

37.58281

118.23721

3790

-4.5

-12.4

12.6

2.7

Mammoth Crest Sierra Nevada

37.55376

118.98012

3514

0.6

-10.1

16.9

Whitney N (Tulainyo Lake)

36.60582

118.27619

3754

-1.7

-11.9

12.7

Sierra Nevada

Longitude Elevation (° N) (M)

max

min

max

min

max

min

temp

Jan

July

Aug

Sept

Annual

17

15

43

1898

436

9

56

35

1974

1.6

532

12

40

58

3064

-5.5

0.8

229

21

23

47

1467

2.7

-6.5

-1.9

36

23

30

22

450

4.5

7.3

-4.1

1.6

297

6

3

23

1473

-0.5

4.1

-7.6

-1.8

114

8

10

19

697

Source: Data excerpted from the PRISM climate model (Daly et al. 1994) for point locations selected as representative of the mid-upper alpine zone for the region. PRISM data represent 1971–2000 normals, with 800 m grid.

al. 2014b, and unpublished data). A cooperative climate station on the summit of Mount Warren (3,757 meters, Western Regional Climate Center) operated for six years before high winds destroyed the major infrastructure in 2011. In 2009, when data were most complete for all months, average annual temperature was –​1.3°C; the lowest monthly minimum temperature (December) was –​22.3°C and the highest monthly maximum temperature (August) was 20.1°C. Winds throughout 2009 were mainly westerly (cold seasons) and southwesterly (warm seasons), with gusts commonly over 54 m s-1 (120 mph). The maximum monthly wind gust (March) was 42 m s-1 (94 mph), and that month had two days with maximum gusts over 134 m s-1 (300 mph) including one at 171 m s -1 (382 mph)! Climatic data to characterize the alpine environment of the White Mountains have been collected for many years at the Barcroft Station at 3,801 meters (Pace et al. 1974, Powell and Klieforth 1991). The mean monthly maximum temperatures at Barcroft vary from a high of 11.9°C in July to a low of –​5.3°C in February. Record maximum temperatures of 22°C have been reached in July and August. Mean monthly maximum temperatures remain below freezing for six months of the year from November through April. Mean minimum temperatures range from a high of 2.4°C in July to a low of –​14.0°C in March. Mean minimum temperatures drop below freezing for every month of the year except July and August (Rundel et al. 2008). Still, many winter days have midday temperatures that rise above freezing. Such temperatures, combined with soils that do not freeze below shallow surface layers, allow for diurnal water uptake by plant species. Because of their position in the rain shadow of the Sierra Nevada, precipitation in the White Mountains is only about a third of that in the Sierra Nevada at the same elevation (see Table 29.1). A Mediterranean-climate pattern remains of winter precipitation in the White Mountains, but a strong added influence of summer convective storms from the south and east brings scattered precipitation events throughout the growing season. Mean annual precipitation at Barcroft is 478 millimeters, with all of this precipitation falling as snow except in July through September. Mean monthly precipitation ranges from a high of 56 millimeters in December to a low of 18 millimeters in September, but year-to-year variation is high. The extremes in annual precipitation over the record period have ranged from 242 to 852 millimeters (Rundel et al. 2008). California alpine ecosystems are influenced by interannual and interdecadal climate modes, such as the El Niño/La Niña system (ENSO) (Diaz and Markgraf 2000) and the Pacific Decadal Oscillation (PDO) (Mantua et al. 1997). These oceanmediated climate modes have alternate states that bring to the California region years of wet, warm winters (El Niño) or cool, dry winters (La Niña). Decadal oscillators such as PDO amplify the effects of one or the other condition. While extreme conditions of ENSO usually occur about every two to seven years, in California more El Niño conditions occurred during the 1980s and 1990s, while La Niña conditions have had a stronghold in the 2000s. The latter tend to modify already dry winters to be record low precipitation years, even in alpine ecosystems. The ENSO system expresses itself in opposite patterns (wet-warm versus cold-dry) in the Pacific Northwest compared to the American Southwest, and northern California lies in the zone of transition between these patterns. Thus northern California alpine ecosystems as far south as the central Sierra Nevada can experience quite dif-

ferent winter temperatures and snowpacks than the southern parts of the range. Other, longer-term trends that affect the alpine zone include multiyear droughts. During the twentieth century in the central Sierra Nevada, these tended to recur every fifteen years or so and persist for five to seven years (Cayan et al. 1998, Millar et al. 2007).

Geologic and Geomorphic Setting Historical Geology of Uplift and Erosion (Subsidence) The geologic history of the Sierra Nevada has been discussed in more detail in previous chapters (see Chapters 4, “Geomorphology and Soils,” and 8, “Ecosystems Past: Vegetation Prehistory”), which have highlighted changing interpretations of the geomorphic development of the range. Traditional understanding held that the present-day Sierra Nevada was a young, uplifted mountain range resulting from Great Basin extensional forces and faulting. Although scientists have long recognized that mountains of volcanic origin existed in the late Mesozoic and early Tertiary at the site of the presentday Sierra Nevada, the prevailing view was that this ancient range had never gained elevation greater than approximately 2,000 meters and had eroded to lowlands during the early to mid-Tertiary. Fault-block tilting in the past 10–​5 Ma was believed to have created the high elevation of the modern Sierra Nevada. New evidence suggests that the Sierra Nevada in fact reached heights of more than 2,800 meters in the early Tertiary and remained high through subsequent millennia (see Chapter 8, “Ecosystems Past: Vegetation Prehistory”). Nevertheless, the form, topography, and elevation of the modern Sierra Nevada were strongly influenced by effects of more recent extensional and faulting processes—​processes that today continue through tectonic action along the Sierra Micro-Plate and Eastern California Shear Zones.

Glacial and Periglacial History The major glacial activity present in California has been in the Sierra Nevada, and this history of multiple glacial advances and retreats has had a major impact on the distribution and fragmentation of the biota. During the last glacial maximum of the Pleistocene (approximately 20 ka; 1 ka = 1000 years), an ice cap 125 kilometers long and 65 kilometers wide spread over most of the high parts of the Sierra Nevada and reached downslope to an elevation of about 2,600 meters. Valley glaciers moving from the ice cap extended as far as 65 kilometers down canyons on the west slope of the range and 30 kilometers down canyons of the steeper eastern escarpment (Raub et al. 2006). Smaller areas of glaciers formed in the Trinity Alps, Salmon Mountains, Cascade Ranges (Mount Shasta, Mount Lassen, Medicine Lake), Warner Mountains, Sweetwater Range, White Mountains, and San Bernardino Mountains. Prior mountain glaciation events are evident most dramatically in the multiple moraines of the eastern Sierra Nevada, some of which reached greater extents than the last glacial maximum. Evidence suggests that Pleistocene glaciers in the Sierra Nevada completely melted at the latest during the thermal optimum of the Holocene 4,000 to 6,000 years ago (Bowerman and Clark 2011). Neoglaciation began with a moderate advance, 3,200 years ago, followed by a possible glacier maxiAlpine Ecosystems   619

FIGURE 2 9.6 Cirque, cliff, and valley slope environments of the eastern Sierra Nevada show erosional and depositional effects resulting from multiple Pleistocene glaciations. Small neoglacial icefields now occupy the highest of these cirque headwalls. Photo: Jeffrey Wyneken.

mum at ~2,800 years ago and four distinct glacier maxima at approximately 2,200 years, 1,600 years, 700 years, and 250 to 170 years ago, the most recent being the largest (Bowerman and Clark 2011). The advent of the global Little Ice Age about 600 years ago (Grove 1988), brought on by shifts in the solar cycle and significant volcanic eruptions, resulted in a period with the coldest conditions of the past 4,000 years. The coldest part of the Little Ice Age in California occurred during the late 1800s and into the early decades of the twentieth century and left a legacy of rock ice formations and highelevation microclimates that continue to have a significant impact.

Geomorphic Settings and Habitats Community composition in alpine communities is strongly influenced by geomorphic structures and their relationship to erosion, snow accumulation, and snowmelt. Such features as soil accumulation and stability, water availability, and exposure to wind are strongly shaped by these settings.

Mountain Summits and Upland Alpine Plateaus Mountain summits in the alpine zone of California are modifications of two primary shapes. One is the classic cone shape, most symmetric in volcanic cones such as Mount Lassen and Mount Shasta, while the other is the highly irregular result of fault-block tectonics. The latter, such as Mount Whitney, often have sheer cliffs on one side (the escarpment) while the other slopes often grade to broad, often surprisingly flat, alpine plateaus. These low-relief uplands, often terraces (treads) with steep cliffs (risers), have long been recognized as important features of the Sierra Nevada (Lawson 1904). These plateaus are interpreted as former lowland erosion surfaces brought to their high-elevation locations during the late Tertiary and Quaternary through episodic tectonic uplift and fault action (Wahrhaftig 1965). The plateaus remain broad and flat, usually with a slight gradient and mostly not incised due to lack of opportunity for snowpack accumulation 620  Ecosystems

and melt from higher elevations. The plateaus present a large habitat landscape for alpine communities at high elevations. Because of their elevations, the summits and high plateaus of the Sierra Nevada and White Mountains, Mount Shasta, and some regions of the Klamath Mountains experience repeated freeze-thaw action, which breaks what would in many cases be exposures of underlying plutons into extensive fields of shattered rock, known as felsenmeer. The dynamic processes present in these habitats makes it difficult for plant communities to develop significant cover in them. Together summits and alpine plateaus experience rather unique climates for complex mountain regions. Given their extension upward to high altitudes and their relative isolation from adjacent barriers, these high points are strongly influenced by synoptic or regional climatology and more likely to “take the pulse” of changes in regional or hemispheric climate trends than are lower mountain elevations (Barry 2008).

Upland Slopes and Basins: Cirques, Cliffs, and Depositional Chutes Below the highest mountain summits and upland plateaus of California’s alpine zone are cirque, cliff, and valley slope environments (Figure 29.6). Valleys in glaciated areas head in cirques, amphitheater-shaped basins with broad, flat floors and sloping walls, whereas unglaciated valleys head in slopes that often are narrow and can have very steep walls. Glaciated valley slopes, especially near the range crests and on escarpment direction, often have steep walls above U-shaped valley floors. The “trim line” represents the upper height of glacial activity. This is often visible as a change in slope and composition of the valley wall, with steep, ice-scoured rocks below the trim line and frost-shattered slopes rising to summits, ridgetops, and upland plateaus above. As a result of differences in substrate texture, vegetation is often quite different above and below trim lines. Glaciated and unglaciated valley walls support cliffs, talus slopes, and avalanche, debris, and landslide chutes that attract lithic-adapted flora and/or those adapted to unstable substrates. Shallower-gradient slopes support stable vegetation communities.

FIGURE 2 9.7 Rock glaciers, such as on the east slope of Mount Gibbs in the Sierra Nevada, are common periglacial features in many alpine canyons. Photo: Constance Millar.

Broken Rock Habitats: Rock Glaciers, Scree, Fellfield, and Talus Slopes Rock glaciers and periglacial talus slopes are widespread geomorphic features associated with cirques and high valleys of the central and southern parts of the Sierra Nevada (Figure 29.7). Although widely overlooked as important features in the past, these now are understood to play an important role in mountain hydrology (Millar and Westfall 2008, 2010). Unlike typical ice glaciers and exposed areas of snowpack, the ice and groundwater contained within rock glaciers and talus slopes are insulated from the direct effects of solar radiation by blankets of rock debris (Clark et al. 1994). Amplifying the insulation effect of rock mantling are internal thermal regimes created by air circulation within the matrix of rock (Millar et al. 2013, 2014b). As a result, thaw of ice in rock glaciers lags behind thaw in typical ice glaciers. The former appear to be in disequilibrium with climate, especially when climates are changing rapidly such as at present. The unique thermal regimes of high and north-facing talus slopes and rock glaciers in the Sierra Nevada are cold enough to support persistent internal ice. Thus persistent cold conditions associated with rock glaciers can provide microclimates equivalent to alpine conditions 1,000 meters higher in elevation. Increasingly, these rocky environments are recognized as unique mountain ecosystems (Kubat 2000), with cold-displaced species such as the predatory rhagidiid mite (Rhagidia gelida) inhabiting internal matrices far below its usual elevation limits (Zacharda et al. 2005) and plant species adapted to unstable lithic environments growing on the rocky mantles (Burga et al. 2004). In California these alpine ecosystems are poorly described, but pilot studies indicate that distinct vegetation associations are found on the surfaces of talus slopes, and that the wetlands supported by talus and rock glacier springs also support distinct plant and arthropod communities (Millar, Westfall, Evenden et al. 2014). Meltwater from snow, internal ice, permafrost, and/or stable groundwater within and below these features appears to provide an important hydrologic reservoir through the summer months (Raub et al. 2006, Maurer 2007, Millar et al. 2013, Millar et al. 2014b), contribut-

ing to streamflow and downslope recharge. In this way, these features support abundant wetlands in high-elevation canyons and provide critical habitat for a host of alpine biota, of which some, like the American pika (Ochotona princeps), depend on wetland habitats supported by adjacent talus fields and rock glaciers (Millar and Westfall 2010). Moreover, wetlands act as sponges to retain water in upper-elevation basins in contrast to meltwaters from annual snowpacks and ice glaciers, which more typically flow out of the uplands in incised channels.

Patterned Ground Patterned ground and related permafrost features are common in arid, cold climates of the world, and permafrost dynamics can have a strong impact on the development of plant communities (Washburn 1980). However, these have been little-mentioned for the Sierra Nevada. This is because permafrost generally has not been assumed to exist in the range generally, although permafrost features now are known to exist in both rock glaciers and high, cold talus slopes (Millar et al. 2013, 2014b). Ample evidence of patterned ground, especially sorted circles and slope stripes from historical (likely Pleistocene) periglacial action, exists in many alpine zones of the Sierra Nevada, especially upland plateaus. Some areas other than rocky slopes such as shallow edges of tarns (Figure 29.8) suggest that these processes are ongoing. Intensive research in the adjacent White Mountains of California has demonstrated the presence of discontinuous, modern, patterned ground features and processes at 3,800 meters. Modern patterned ground processes with sorted circles, nets, and stripes become the dominant landscape phenomena above 4,150 meters (Wilkerson 1995). Freeze-thaw cycles throughout the year are common in some parts of the White Mountains, with over 220 cycles per year observed (LaMarche 1968). While the larger patterned ground features in the White Mountains have been assumed to be relict (Wilkerson 1995), the presence of active permafrost processes there suggests that similar active processes may exist in the Sierra Nevada, especially on exposed plateaus and ridgetops Alpine Ecosystems   621

where water collects yet wind sweeps away snow, maintaining exposure of the ground surface to freezing air (Millar and Westfall 2008, Millar et al. 2013).

Wetlands Wetland environments, characterized by high groundwater moisture, are important generally in mountain regions of the world for the distinctive biodiversity they support. This is especially true in the California alpine zone, where Mediterranean-type and dry continental (Great Basin) climates and low latitudes combine to drive temperatures high and make water during the growing season scarce. Where wetlands occur, plant communities differ significantly from those occurring on uplands (Sawyer et al. 2009, Weixelman et al. 2011). Invertebrate assemblages also differ substantially between alpine upland and wetland habitats (Holmquist et al. 2011), with wetlands tending to be more speciose. In the California alpine zone, wetlands around springs and seeps and wet meadows derive their high soil moisture primarily from persistent subterranean groundwater sources that often are augmented by snowpack and snowmelt, especially where slope gradients are low (Figure 29.9). Snowbeds (small wetlands fed by late-lying snowfields that recur in specific locations) and glacial forefields (including rock glacier and talus slope forefields) are fed by closely adjacent water sources (melting snow and ice bodies). Riparian corridors tend to be very narrow and to track valley bottoms along streams as well as along creek drainages of valley slopes. The character of wetlands and thus the habitat they create for plants and animals varies depending on the substrate, which affects soil water-holding capacity and acidity. Some soils (e.g., volcanic) drain rapidly while others (e.g., granitic) can develop so-called Teflon basins that hold water near the surface. Aquatic environments including the moving water of streams, creeks, and rivers and the still water of lakes, tarns, and ephemeral pools are also extremely important habitats in the otherwise relatively dry alpine zones of California (see Chapter 32, “Lakes”). In glaciated regions, tarns, paternoster ponds, and moraine-impounded lakes constitute the bulk of still waters, which number many in the Sierra Nevada ( more than four thousand) (Knapp 1996) and Klamath Ranges and very few in the drier interior Great Basin ranges (Warner Mountains, Sweetwater Mountains, White Mountains). Snowmelt-derived streams flowing from high slopes and cirques are sources for important rivers of mid-montane and lowland reaches. Due to the presence of an extensive ice cap during the last glacial maximum in the Sierra Nevada, lakes and streams above 1,800 meters were fishless during the late Pleistocene and Holocene and prior to European settlement supported rich amphibian and invertebrate diversity (Jennings 1996, Erman 1996). Stocking of non-native fish starting in the early twentieth century, however, widely transformed these alpine waters; amphibian species have greatly declined in abundance and distribution, while aquatic invertebrate diversity has shifted drastically in response to predation by fish (Knapp 1996).

Glaciers and Permanent Snowfields More than seventeen hundred permanent snow or ice bodies are located in California, with seventy of these larger than 622  Ecosystems

0.1 square kilometers (Figure 29.10). Twenty of these glaciers have been named—​seven on Mount Shasta and thirteen in the Sierra Nevada. Snow and ice bodies are also located in the Trinity Alps and near Mount Lassen. In total, permanent snow and ice bodies cover over 46 square kilometers of California (Fountain et al. 2007). An inventory in 2006 identified 497 glaciers covering a total area of 50 square kilometers in the Sierra Nevada (Raub et al. 2006), while 421 rock glaciers and related features were inventoried from the central Sierra Nevada alone (Millar and Westfall 2008). Aside from the Sierra Nevada, the alpine zone of California supports glaciers primarily on Mount Shasta, where seven to ten glaciers were recognized as of 1987 (USGS 1986, Rhodes 1987). Despite regional warming over the past half century, the glaciers of Mount Shasta have shown a greater sensitivity to precipitation than to temperature and have continued to expand following a contraction during a prolonged drought in the early twentieth century (Howat et al. 2007). However, the strong warming trend predicted by regional climate models will be the dominant forcing with an expected near-total loss of glaciers on Mount Shasta and elsewhere in California by the end of the twenty-first century (Basagic 2008).

Processes and Ecosystem Dynamics Alpine and subalpine watersheds in California play critical hydrologic roles for downstream agriculture and urban development in California, particularly with respect to the seasonality and amount of snowmelt (Bales et al. 2006). The hydrologic flow and biogeochemical processes of these highmountain ecosystems are sensitive to small changes in growing-season temperature and water availability. The thin oligotrophic soils of alpine watersheds also have the potential to be significantly affected by atmospheric nitrogen deposition and acidification associated with the expanding urbanization of the San Joaquin Valley (Sickman et al. 2003). While there has been a long history of research on high-mountain hydrology and biogeochemistry of lakes and streams in the Sierra Nevada, these studies have largely focused on subalpine watersheds (see Chapters 28, “Subalpine Forests,” and 32, “Lakes”). Hydrologic studies in high-elevation watersheds in the Sierra Nevada have modeled streamflows and water balance and the association of snowmelt and runoff with the solute chemistry of aquatic systems (Kattelmann and Elder 1991, Williams and Melack 1991). Climate change is expected to affect hydrology by increasing snowmelt rates, promoting earlier runoff (Wolford and Bales 1996). Acid deposition has the potential to alter streamwater pH and increase sensitivity to acidification (Sickman et al. 2001). A key factor influencing nitrogen cycling is duration of snow cover, which influences plant uptake, mineralization, and mineral nitrogen export (Meixner and Bales 2003). Because of the extreme heterogeneity of soil depth and vegetation cover in alpine watersheds, soil mineralization rates for nitrogen are highly site-specific (Miller et al. 2009). Although data on aboveground production are available for subalpine communities in the Sierra Nevada, very little attempt has been made to collect such data in alpine habitats above treeline. Studies from the central Rocky Mountains and the European Alps have generally found aboveground production rates of approximately 100 to 400 g m-2 yr-1, with

FIGURE 2 9.8 Sorted circles result from repeated freeze-thaw action when lake waters are shallow in alpine environments, as along the borders of the deep basin of Silverpine Lake, Sierra Nevada. Photo: Constance Millar.

FIGURE 2 9.9 Wetlands commonly surround alpine lakes, as at Greenstone Lake in the Sierra Nevada, where talus contributes persistent springs in addition to streamflow from upland glaciers. Photo: Constance Millar.

FIGURE 2 9.10 North Palisade Glacier, Sierra Nevada. Glaciers in the California alpine zone are remnants from the Little Ice Age, 1450–​1920 CE and, except for glaciers on Mount Shasta, exist only in highest headwall cirques. Photo: Constance Millar.

typical values closer to 200 g m-2 yr-1 (see Bowman and Seastedt 2001, Körner 2003). However, these mean values mask the high spatial heterogeneity in rates of aboveground net primary productivity, with plot-level values ranging from as low as 50 g m-2 yr-1 to as much as 500 g m-2 yr-1 or more (Bowman et al. 1993, Walker et al. 1994). This heterogeneity is strongly influenced by topographic controls on microclimate conditions as well as biotic impacts from grazers and burrowing animals (Scott and Billings 1964). Interannual variation in productivity is also typical, with summer temperatures, patterns of snowmelt, and levels of summer drought strongly influencing the length of growing season. On an ecosystem level, much of the biomass accumulation and net primary productivity of alpine communities occurs belowground. Ratios of belowground to aboveground biomass range from approximately 2.5 to 8.8 in studies carried out in alpine systems in the Rocky Mountains and European Alps (see Bowman and Seastedt 2001, Körner 2003). Studies of alpine meadows at Niwot Ridge in the Rocky Mountains have reported ratios of belowground to aboveground productivity that vary from approximately 1.0 in moist alpine meadows to 1.6 in wet meadows and 2.3 in dry meadows (Fisk et al. 1998).

Vegetation and Flora Local distribution of individual plant habitats is determined strongly by features of the physical environment including topographic position, wind exposure, snow accumulation, and soil depth and drainage (Taylor 1977, Sawyer and KeelerWolf 2007). These diverse habitats include windswept ridges, snowbeds, dry meadows, basins and gentle slopes, and shrubdominated drainage channels. The most severe conditions for plant growth occur on windswept summits and ridges. These communities may be snow-free through much of winter and typically consist of well-drained, coarse soils. Under extreme conditions, patterned soils can occur in these settings from frost heaving. Fellfield communities fall into this category and often exhibit a dominance of cold- and drought-tolerant mats and cushions along with low-growing herbaceous perennials. Areas with late-lying snowbeds are characterized by a short growing season with limited water stress and typically support communities dominated by grasses and sedges. Dry meadows occur in areas with shallow soils and with little access to soil moisture once drought conditions extend into the summer. These communities are dominated by a mix of low-growing herbaceous perennials and graminoids and typically are more limited in growth by water availability than by the length of the growing season. Broad alpine valleys and meandering streams exhibit a mix of herbaceous communities dominated by a mix of graminoids and forbs on shallower soils and shrub cover on areas with water availability and some shelter from wind exposure. Willow (Salix) species often form dense, low-growing stands or mats, and a diversity of low, ericaceous shrubs are often present. More arid areas of glaciated bedrock with shallow and rapidly draining sandy soils may support stands of Artemisia, particularly east of the Sierra crest. A number of studies have developed relatively detailed classifications of flora and plant community alliances and associations for major habitats (Howell 1951, Klikoff 1965, Pemble 1970, Taylor 1976a, Taylor 1976b, Major and Taylor 1977, Tatum 1979, Burke 1982, Porter 1983, Constantine-Shull 624  Ecosystems

2000). Alpine herbaceous and shrub alliances are described comprehensively by Sawyer et al. (2009) in the broader context of California vegetation types, although these types are often difficult to reconcile with specific stands of alpine vegetation. Ecological habitat descriptions for the alpine zone of the White Mountains have been made by Morefield (1988, 1992) with a somewhat simpler system of seven categories proposed by Rundel et al. (2008). Along a rough gradient from mesic to xeric these ecological habitats comprise aquatic sites, wet sites (areas with saturated soils and riparian habitats), moist sites (wet meadows and areas with snowmelt accumulation), fellfields with seasonal moisture availability, talus slopes, open slopes, and dry rocky slopes. Howell (1944) commented on the elements of boreal flora in the Klamath Mountains.

Plant Functional Groups and Adaptive Traits Regardless of specific growth form, alpine plants are typically small and grow close to the ground. Spacing between plants is often wide with intervening areas of soil or rock. Such patterns illustrate the role of the physical environment in finely influencing microclimate to create favorable or nonfavorable microsites for plant establishment and growth. A few centimeters’ difference in microtopography can have significant influence on air and soil temperatures, wind desiccation, and snow accumulation. While plant life forms are commonly discussed in modern treatments of alpine vegetation (Billings 2000, Bowman and Seastedt 2001, Körner 2003), only limited attention has been given to quantifying these as plant functional groups. Herbaceous perennials of a variety of forms and architectures (i.e., broad-leaved herbaceous perennials, mats and cushions, graminoids, and more rarely geophytes) form the dominant community cover in temperate alpine ecosystems. Also present with lower species richness are shrubs, as subshrubs (chamaephytes) with a low form of woody growth. Other plant life forms such as taller woody shrubs (phanerophytes) and annuals (therophytes) are rare in most alpine habitats. Life forms of the Sierra Nevada alpine flora, as in other temperate alpine floras, are heavily dominated by broad-leaved herbaceous perennials (48% of species) followed by graminoid perennials (22%) and mats and cushions (12%). Annuals and woody shrubs account for 6% each of the flora. Life forms are similar in the alpine flora of the White Mountains with broad-leaved herbaceous perennials dominant (53%) followed by graminoid perennials (22%), mats and cushions (11%), annuals (8%), and woody shrubs including low subshrubs less than 50 centimeters in height (6%). The proportion of annual species is relatively high in the Sierra Nevada and White Mountains compared to other continental alpine regions. Annual plants face a strong disadvantage in alpine ecosystems because of the short, cool growing season during which they must complete their entire life cycle. The relatively warm growing season temperatures of alpine habitats in California likely explains this greater success. Clear differences in root morphology exist between growth forms, suggesting functional differences in adaptive strategies. Cushion plants and subshrubs exhibit characteristic tap roots, while mat-forming cushions also have shallow, spreading, adventitious roots arising along stems to take advantage of temporary surface soil moisture (Billings and Mooney 1968). Perennial graminoids have spreading, fibrous roots. A consistent

pattern across growth forms exists wherein most alpine species have stomates on both their upper and lower leaf surfaces, reflecting the high light environments in which they grow (Rundel et al. 2005). Studies of seed germination and seedling establishment across a gradient of changing soil substrate in the Sierra Nevada have shown that water might be the limiting factor for species germination and that differential nutrient availability across soil types strongly influences early seedling growth (Wenk and Dawson 2007). Ecophysiological traits of alpine plants have been broadly reviewed (Billings and Mooney 1968, Billings 1974, Bowman and Seastedt 2001, Körner 2003) but are not well studied in California alpine plants. Rundel et al. (2005) compared ecophysiological traits of leaf structure, midday leaf temperature, mean maximum photosynthetic rate, and predawn and midday water potentials among four perennial life forms in an alpine fellfield in the White Mountains without finding consistent patterns associated with plant functional types. However, plant growth form may significantly influence diurnal leaf temperatures. The ability of some canopy architectures to promote high leaf temperatures helps explain the ability of the C4 grass Muhlenbergia richardsonis to grow successfully at elevations up to nearly 4,000 meters in the White Mountains (Sage and Sage 2002). In adjacent communities the leaf temperatures of two upright species (Chrysothamnus viscidiflorus and Linanthus nuttallii subsp. pubescens) track ambient air temperature, while the mat-forming Penstemon heterodoxus has leaves that are heated significantly compared to air temperatures at midday (Rundel et al. 2005).

Floristics Diversity and Phylogenetic Breadth and Depth The alpine zone of the Sierra Nevada, if defined as nonforested areas at or above 3,500 meters, includes 385 species of native vascular plants (Rundel 2011). If the alpine boundary were defined as at or above 3,300 meters, the alpine flora would grow to 536 species. Ninety-seven species reach elevations of 4,000 meters, and 27 species reach to 4,200 meters. Only a relatively small number of species are high-elevation specialists; 9 species have ranges restricted to elevations above 3,500 meters, and an additional 67 species (17% of the flora) are restricted to subalpine and alpine habitats. These highelevation specialists are spread across multiple families but as a group share the feature of relatively high endemism compared to the overall Sierra Nevada flora. More than a quarter of the species in the Sierran alpine flora have elevational ranges that extend as low as foothill habitats below 1,200 meters (Rundel 2011). Over half of the Sierran alpine species occur in just six families, led by the Asteraceae (55 species), Poaceae (39 species), Brassicaceae (34 species), and Cyperaceae (39 species). The largest genus present is Carex with 29 species, and 18 more species would be added by lowering the alpine boundary to 3,300 meters. Next in size are Draba (14 species) and Lupinus (11 species). The level of endemism in the Sierra Nevada alpine flora is moderate to high depending on how the geographic unit for endemism is defined. The 36 Sierran endemics present in the alpine flora compare with 205 endemic taxa for the montane areas of the range and thus compose 18% of the endemic flora of the higher Sierra Nevada (Rundel 2011). The unique Californian component of the alpine flora of the Sierra Nevada is considerably greater if one con-

siders 31 species present in the alpine flora that are not limited to the Sierra Nevada but occur elsewhere in California or in ranges adjacent to the Sierra Nevada. Under this definition, 66 endemic taxa represent 16% of the Sierran alpine flora. This is high compared to other alpine ranges in continental North America and Europe and reflects both the environmental stress associated with the summer-dry, Mediterraneantype climate of the Sierra Nevada and the relative isolation of the range. The alpine zone of the White Mountains, defined as nonforested areas above 3,500 meters, includes 163 native species of vascular plants. Seven families account for nearly twothirds of the flora, led by the Asteraceae (30 species), followed by the Brassicaceae (18 species), Poaceae (17 species), Cyperaceae (15 species), Rosaceae (9 species), Caryophyllaceae (9 species), and Polygonaceae (7 species). While 31% of the alpine flora are restricted to alpine habitats, more than two-thirds of this flora extend to lower-elevation communities in the White Mountains of montane forest, pinyon-juniper woodland, or cold desert. Fellfields form the characteristic habitat for 41% of the flora, while moist meadows and open-slope habitats contain 24% and 22% of the flora, respectively. Endemism is low in the alpine zone of the White Mountains, with just three endemic species, although three more come close to being endemic. Two of these have small populations on the east slope of the Sierra Nevada, and one has a disjunct population in the Klamath Mountains. The Sweetwater Mountains, 33 kilometers east of the Sierra Nevada, support an alpine flora of 173 species in 16 square kilometers of alpine habitat and share 94% of this flora with the Sierra Nevada (Bell-Hunter and Johnson 1983). The small alpine zone on Mount Grant to the north of the Sweetwater Mountains in western Nevada supports a flora of 70 species dominated by Sierra Nevadan elements in just 2.6 square kilometers (Bell and Johnson 1980).

Evolution of the Flora The evolution of the alpine flora of the Sierra Nevada has involved a variety of factors including geologic history, climate history, modes of colonization, and factors promoting regional speciation (Stebbins 1982). On a biogeographic basis, there are strong indications of a north-to-south route of colonization of high mountain areas of the Sierra Nevada during the late Pliocene and Pleistocene. This evidence comes from a pattern of decreasing presence of Rocky Mountain floristic elements and an increased number of endemic alpine species as one moves from the northern to southern crest of the range where elevations are higher (Chabot and Billings 1972, Raven and Axelrod 1978, Rundel 2011). This gradient is shaped not just by geographic distance but also by steadily decreasing levels of precipitation moving to the south. Floristic elements of subalpine, subalpine wet meadows, and other moist sites typically have broad geographic ranges but become increasingly restricted to the most mesic sites as precipitation decreases southward in the Sierra Nevada (Kimball et al. 2004). Species growing in xeric rocky habitats show higher levels of endemism and smaller range sizes due to isolation and divergence from ancestral populations distributed in wetter habitats to the north. Endemism also increases in the Sierra Nevada with increasing elevation. Of species obligately occurring above 3,000 meters, fully one-third are California endemics, double the level of endemism for the entire mountain range. Alpine Ecosystems   625

A number of alpine species reach their southern occurrence limit on Mount Lassen, suggesting that some of these and other Cascade Range species might have been present in the Sierra Nevada in the late Pliocene or early Pleistocene. Although the species compositions of lower- and middle-elevation conifer forests of Lassen National Park are strongly related to those of the Sierra Nevada, the summits of the highest peaks in Lassen support an alpine flora with stronger floristic links to Mount Shasta and the Cascade Range to the north (Gillett et al. 1995). The Klamath Mountains also mark the southern distribution limit for a number of high-elevation species that do not occur in the Sierra Nevada (Howell 1944). The relative isolation of the Sierra Nevada from northern ranges and the summer drought have acted as a filter to exclude some widespread, circumpolar, arctic-alpine species such as Dryas integrifolia and Silene acaulis, which do not occur in California. More controversy exists about the possible migration of significant components of the Sierran alpine flora from the Rocky Mountains across the Great Basin. Both geological and paleobotanical evidence exist to suggest that the mean elevation of the Great Basin was up to 1,500 meters higher in the Miocene and that the current basin and range topography is the result of subsidence rather than uplift (Wernicke et al. 1988, Wolfe et al. 1997). The presence of higher elevations in the Great Basin during the Pleistocene could have provided stepping-stones for the dispersal of alpine organisms from the east. Several notable examples exist of disjunct Rocky Mountain species with restricted distributions in the central and southern Sierra Nevada, often growing in azonal soil conditions (Major and Bamberg 1967, Taylor 1976a). Molecular evidence has shown that at least one lineage of butterflies entered the Sierra Nevada by this route (Nice and Shapiro 2001). However, other authors feel that the majority of these disjunct plant species reached the Sierra Nevada by the same dominant route via the Cascade Range (Chabot and Billings 1972). Only a preliminary understanding exists of the origins of the endemic alpine flora of the Sierra Nevada and White Mountains, and modes of speciation are clearly complex (Rundel 2011). Factors promoting endemism in the alpine flora include the recent uplift of the mountains, the relative isolation of the Sierra Nevada from other ranges, glacial restrictions on migrations, Holocene climate variability, the mixing of desert and mountain floras, and the unusual conditions of summer drought (Chabot and Billings 1972). There are many examples of genera in the alpine flora where apomixis (emergence of asexual reproduction) has been important in speciation. These include Boechera (Schranz et al. 2005, Dobeš et al. 2007) and Draba (JordonThaden and Koch 2008) (Brassicaceae), Antennaria (Bayer and Stebbins 1987) and Arnica and Crepis (Noyes 2007) (Asteraceae), Poa and Calamagrostis (Poaceae), and Potentilla (Rosaceae) (Asker and Jerling 1992). Diploid lineages of polyploid complexes often occupy unglaciated areas and resist introgression, hypothetically due to a significantly higher seed set. However, asexual apomictic populations are often more widespread than their sexually reproducing relatives in glaciated areas. The advantages of apomixis include reproductive isolation and stability of vegetative lineages in their area of distribution. Other modes of alpine speciation include population disjunction, reproductive isolation (Chase and Raven 1975), and upslope migration and colonization by arid-adapted lowland taxa at the end of the Pleistocene (Went 1948, 1953). 626  Ecosystems

The White Mountains present a particularly interesting area for study of evolutionary history of the flora and fauna given their position at the interface between two major geomorphic provinces, the Sierra-Cascade Province and the arid Basin and Range Province. Despite this interface, the White range is isolated from direct contact with high elevations of these provinces. Moreover, warmer and more xeric climatic conditions of the middle Holocene climatic optimum allowed an upward movement of subalpine conifers, restricting the area available for growth of alpine communities (Jennings and Elliot-Fisk 1991, 1993). Thus the White Mountains present an example where both climate history and geographic isolation have played significant roles in the evolution of the biota.

Fauna The alpine zone of California harbors relatively low mammal diversity. Few species are restricted to alpine habitats, while many more use the alpine environment transiently or seasonally (Table 29.2). Large herbivores such as mule deer (Odocoileus hemionus) (Anderson and Wallmo 1984) and desert and Sierra Nevada bighorn sheep (Ovis canadensis nelsoni and O. c. sierrae, respectively) (Wehausen et al. 2007) use the alpine zone in summers both for foraging and as retreat and escape from predators. The prime predator of these ungulates, mountain lion (Puma concolor), occasionally follows them to the alpine regions. Sierra Nevada bighorn sheep (Ovis canadensis sierrae) makes major use of alpine areas of the Sierra Nevada during summer. Individuals range over a broad elevational distribution, with winter range as low as 1,450 meters. Individuals near the Mono Basin also migrate far to the east. Sierra Nevada bighorn sheep favor open areas where predators can be seen at a distance and steep rocky slopes can serve as refuge from attacks by mountain lions. There once were as many as a thousand bighorn sheep in the Sierra Nevada, but this number had declined to about 125 adults at the time of their listing as endangered in 1999 (Wehausen et al. 2007). By 2007 they had recovered to more than 400 individuals distributed in a metapopulation of eight subpopulations. They are approaching the target population size of 500, thanks to concerted efforts both to augment herd units and to reduce predator pressure. Threats to Sierra Nevada bighorn sheep survival include disease transfer from domestic sheep, mountain lion predation, and extreme climate conditions. Many more small and meso-mammals occur in the alpine zone than large animals, and a few are highly restricted to that zone. Among the more charismatic of small mammals is the American pika (Ochotona princeps), a small rabbit relative highly restricted to talus slopes and similar broken-rock landforms (Smith and Weston 1990) (Figure 29.11). Pikas range throughout the mountains of western North America. In California they are found throughout the Sierra Nevada, White Mountains, Bodie Mountains, Sweetwater Mountains, southern Cascades, and Warner Mountains (USFWS 2010, Millar and Westfall 2010.) Like other lagomorphs, pikas do not hibernate and collect a diverse range of herbaceous and shrubby vegetation during the warm season, which they cache in stacks (referred to as “haypiles”) within the talus and consume during the winter. Pikas are poor thermoregulators, tolerant of cold yet sensitive to heat, and are often assumed to be alpine-restricted and at risk from global warming. In

TA B LE 2 9. 2 Mammal species known to occur in alpine ecosystems of California

Common name

Scientific name

Elevation (m)

Citation for species’ elevation

Large mammals Desert bighorn sheep

Ovis canadensis nelson

4300

California Department of Fish and Wildlife

Sierra Nevada bighorn sheep

Ovis canadensis sierra

1500–4270

California Department of Fish and Wildlife

Mule deer

Odocoileus hemionus

< timberline

Anderson and Wallmo 1984

Black bear

Ursus americanus

100–3000+

California Department of Fish and Wildlife

Sierra Nevada red fox

Vulpes vulpes necator

1200–3600

Perrine et al. 2010

*Mountain lion

Puma concolor

< 4000

California Department of Fish and Wildlife

Mesocarnivores and small predatory mammals *Sierra Nevada marten

Martes americana sierra

2300–3150

Kucera et al. 1996, Storer et al. 2004

Long-tailed weasel

Mustela frenata

< 3500?

Storer et al. 2004

Small mammals White-tailed jackrabbit

Lepus townsendii

< 3650

Storer et al. 2004

Bushy-tailed woodrat

Neotoma cinerea

1500–4000

Grayson and Livingstone 1989; Storer et al. 2004

Golden-mantled squirrel

Callospermophilus lateralis

1646–3200

Moritz et al. 2008

Mount Lyell shrew

Sorex lyelli

2100–3630

Epanchin and Engilis 2009

Water shrew

Sorex palustris

1658–3535

Epanchin and Engilis 2009

Sorex monticolus

3400

Epanchin and Engilis 2009

Sorex vagrans

3400

Epanchin and Engilis 2009

Sorex tenullus

4267

Epanchin and Engilis 2009

Belding ground squirrel

Urocitellus beldingi

2286–3287

Moritz et al. 2008

Alpine chipmunk

Tamias alpinus

2307–3353

Moritz et al. 2008

American pika

Ochotona princeps

2377–3871

Moritz et al. 2008

Deer mouse

Peromyscus maniculatus

57–3287

Moritz et al. 2008

Yellow-bellied marmot

Marmota flaviventris

2469–3353

Moritz et al. 2008

Source: * indicates occasional use in the alpine zone.

fact, pikas range from sagebrush-steppe ecosystems through the montane zones to the highest summit reaches and find their optimal elevation range from the mid-subalpine to midalpine zone (Millar and Westfall 2010; Millar et al. 2014a). The capacity of pikas to tolerate warm ambient temperatures despite their thermal sensitivity relates both to the unique microclimates of talus and related landforms (Millar et al. 2014b) and to their behavioral adaptations. Talus thermal regimes are highly buffered from external air, remaining cool in summer and warm in winter. Pikas adapt behaviorally by using this “air-conditioned” habitat for refuge as needed in response to external air temperatures (Smith 1974). Yellow-bellied marmots (Marmota flaviventris) are another important mammal species for alpine ecosystems in the White Mountains, Sierra Nevada, and higher mountain ranges to the north. Yellow-bellied marmots have a harem-

polygynous social system whereby a male defends and mates with one or more females. Female daughters often do not disperse and settle around their mothers. Sons invariably disperse as yearlings and try to find and defend one or more females. Females tend to breed as two-year olds. Litter sizes average a bit over four pups, of which about half survive their first year. Yellow-bellied marmots chuck, whistle, and trill when alarmed by predators. They breed in alpine and subalpine meadows. A number of other small mammal species have ranges extending up to elevations with alpine conditions but are more characteristic of upper montane and subalpine habitats. These include the bushy-tailed woodrat (Neotoma cinerea) (Grayson and Livingstone 1989), Belding’s ground squirrel (Urocitellus beldingi), golden-mantled squirrel (Callospermophilus lateralis), alpine chipmunk (Tamias alpinus), deer mouse Alpine Ecosystems   627

FIGURE 2 9.11 The American pika, a typical small mammal of the alpine zone. Photo: Andrey Shcherbina.

(Peromyscus maniculatus), and Mount Lyell shrew (Sorex lyelli) (Epanchin and Engilis 2009). Recent studies resampling century-old plots in Yosemite National Park have shown that all of these species as well as pikas (although the sample size for pikas was low and the patterns not strongly significant) and yellow-bellied marmots have increased their lower elevational limit of occurrence, likely reflecting warming climatic conditions (Moritz et al. 2008, Tingley et al. 2009). Several species of small to medium-sized carnivores have ranges that extend beyond the conifer zone into open rocky alpine areas of the Sierra Nevada and Cascade Range. These include the rare wolverine (Gulo gulo), which may prey on marmots, and the rare Sierra Nevada red fox (Vulpes vulpes necator) (Perrine et al. 2010). American martens (Martes americana) feed on a variety of vertebrates including chipmunks and ground squirrels and occasionally prey on pikas (Kucera et al. 1996, Jameson and Peeters 2004). While bird use of alpine habitats is largely seasonal and transitory, a notable exception is the gray-crowned rosy finch (Leucostichte tephrocotis), which is a common winter resident of the high Sierra Nevada, Mount Shasta, Mount Lassen, and the Sweetwater and White Mountains at elevations up to 4,000 meters. It forages on the ground, in dwarf shrub habitat, and in alpine barrens above timberline for seeds and insects. The finch breeds in talus slopes and in rock-face crevices. In their California range, these birds are nomadic rather than migratory. Another common member of the alpine avifauna that breeds and summers in California is the water pipit (Anthus spinoletta alticola), which ranges from the central to southern Sierra Nevada and occasionally uses high elevations of the San Gorgonio Mountains of southern California (Miller and Green 1987). Water pipits prefer wet meadow habitat but also frequent open rocky plateaus and slopes. Sage grouse (Centrocercus urophasianus) inhabit high-elevation sagebrush and grassland habitat across the Great Basin and occur as high as 3,700 meters in the White Mountains. Resident birds in subalpine forests such as pine siskin (Carduelis pinus), darkeyed junco (Junco hyemalus), mountain quail (Oreortyx pictus), Clark’s nutcracker (Nucifraga columbiana), and others may extend their foraging above timberline but are not regular residents. Red-tailed hawks (Buteo jamaicensis), peregrine falcons (Falco peregrinus), and golden eagles (Aquila chrysaetos) 628  Ecosystems

are occasionally observed flying over alpine habitats but are not regular residents and breed at lower elevations (Charlet and Rust 1991). The Sierra Nevada yellow-legged frog (Rana muscosa), native to high subalpine and alpine lakes in the mountains, was once the most common frog species over broad areas. Over the past century, however, this species has dramatically declined in abundance (Drost and Fellers 1996, Vredenburg et al. 2005). Although this decline was largely attributed for many years to the introduction of non-native trout and/or to pesticides, recent declines have continued even in apparently unpolluted lakes without fish. Studies have now identified pathogenic chytrid fungi (Batrachochytrium dendrobatidis) as an additional cause of population loss (Briggs et al. 2005). It has been suggested that the former abundance of this species made them a keystone predator and prey and a crucial agent of nutrient and energy cycling in Sierra Nevada aquatic and terrestrial ecosystems (Drost and Fellers 1996). A second species of once-common amphibian in wet subalpine and alpine meadows that has experienced sharp drops in population numbers in recent decades is the Yosemite toad (Anaxyrus canorus). The species epithet references the melodic call of the male toads. A notable feature of this species is the sharp differentiation in color patterns of males and females, arguably the greatest sexual dimorphism of any anuran in North America. These are long-lived toads with upper age limits estimated from fifteen to twenty years—​a life span thought to be an adaptation to their seasonal, high-elevation environment.

Human Interactions Despite their high elevation, abundant evidence indicates that Native Americans made at least limited use of alpine ecosystems in both the Sierra Nevada and White Mountains. Early use of alpine zones in California included transient use by male hunters to track and procure large game, likely with significant impacts on desert and Sierra Nevada bighorn sheep populations (Wehausen et al. 2007). After about two thousand years, family groups started to move to alpine zones for the warm season, where they established village sites in the high White Mountains as well as in two alpine ranges of Nevada (Bettinger 1991, David Hurst Thomas, pers. comm. 2012). With entire family units residing in alpine regions, much wider use of mammal species occurred for consumption, clothing, and shelter materials, and small mammals (rodents and lagomorphs) were hunted regularly (Grayson and Livingstone 1989). It is clear also that prehistoric humans influenced the abundance and distribution of deadwood in alpine landscapes, complicating interpretations of paleotreelines (Grayson and Millar 2008). The heavy summer grazing of high Sierran meadows by sheep was widespread up into alpine meadows in the last four decades of the nineteenth century. Several accounts describe how overstocking and overgrazing altered vegetation composition in high-elevation Sierran meadows (Ratliff 1985; see Chapter 31, “Wetlands”). A permit process to limit grazing in national park lands began in 1905, and much recovery has occurred, although cattle grazing is still permitted in some high-mountain meadows on national forest land. Grazing in high-elevation meadows of the White Mountains was halted in 1988, and recovery has been taking place (Ababneh and Woolfenden 2010).

Invasive Species Alpine plant communities in California have remained free of any significant invasion by non-native plant species, likely due to the extreme conditions presented by the physical environment. A small number of non-native species have been collected around areas of human activity in the alpine zone of the White Mountains, but none of these appear to have become permanently established (Rundel et al. 2008). Information is lacking on significant establishment of non-native plant species in the high Sierra Nevada. Past disturbance by grazing and other human activities suggests that an absence of propagule dispersal is not the limiting factor in the rarity of non-native species. Nevertheless, aggressive species from other global regions of high-elevation habitat could become widely established in the future if introduced. Subalpine and alpine lakes in the Sierra Nevada originally lacked fish, but the widespread introduction of non-native trout species began in the mid- to late nineteenth century and populated all watersheds. Fish stocking was completely halted in the Sierran national parks in 1991 but continues on national forest lands. Studies in high-elevation Sierran lakes and streams have shown significant impacts of introduced trout on native trout, amphibians, zooplankton, and benthic macroinvertebrates (Knapp 1996; see Chapter 32, “Lakes”). White-tailed ptarmigan (Lagopus leucurus) were introduced to the high Sierra Nevada in 1971–​1972 by the California Fish and Wildlife Service and have become well established locally in alpine grassland and fellfield habitats. No studies to date have assessed possible impacts of this introduction.

Conservation Alpine ecosystems in California today remain largely free of major human impacts due to their isolated locations and positions within protected areas and national parks. Some highelevation degradation still takes place on national forest lands, but the major effects today are the scattered impacts of summer cattle grazing and recreational activities including pack trains and mountain biking. These activities are much more common in subalpine meadows rather than alpine meadows. Alpine watersheds with pack animal presence and summer cattle grazing have increased periphytic algal biomass, attached heterotrophic bacteria, and E. coli compared with nongrazed areas. Thus pollution from cattle grazing might be a significant cause of deteriorating water quality within some watersheds (Derlet and Carlson 2006, Derlet et al. 2012, Myers and Whited 2012). Invasive plants have not yet become a conservation issue, but this could change in the future. Introduced trout in alpine streams and lakes have certainly had an impact on native amphibian populations. Possible impacts of introduced white-tailed ptarmigan have not been studied. Overall, the collective impacts from all of these invasions are relatively small. Of much greater concern, as described earlier, is the potential impact of climate change on alpine ecosystems.

Climate Adaptation The significant roles of climatic variables in shaping alpine ecosystem productivity suggest that climate change impacts on alpine plant communities could be more pronounced than

on lower-elevation communities (Grabherr et al. 2000). Moreover, alpine ecosystems are predicted to experience some of the highest levels of warming globally and are expected to exhibit signs of change before other terrestrial ecosystems because of their high sensitivity to disturbance. Alpine ecosystems likely also will be affected by other attendant factors such as declining snowpack, earlier spring runoff, and earlier phenology (Cayan et al. 2001, Duffy et al. 2007, Mote et al. 2005, Stewart et al. 2005). The international program for monitoring response of alpine plants to climate change, GLORIA (Global Observation Research Initiative in Alpine Environments; see Grabherr et al. 2000, Malanson and Fagre 2013), promotes stations on mountain summits worldwide. For each “target region,” standardized monitoring designs are installed on four mountain summits that span the elevational extent from upper treeline to the highest peak in the local region. Seven target regions have been established in California: the Panamint Range; the White Mountains; the Sierra Nevada; and the Sweetwater Mountains (see the North American GLORIA website at http://www.fs.fed.us/psw/cirmount/gloria/). The earliest were established in 2004, and several will undergo the second round of five-year remeasurements in 2014. Early results show no striking or significant changes in vegetation or floristics; the most obvious changes are increases in soil temperature. Data from the California GLORIA target regions document local floristic diversity of the summits and provide an excellent reference for local conditions. In addition to these standard target regions, the White Mountain Research Center (at the University of California) operates as one of two GLORIA master sites (the other is in Austria) where interdisciplinary alpine studies are conducted in addition to the multisummit protocol to monitor the impacts from and adaptation to climate change of alpine biota and ecosystems (see http://www .fs.fed.us/psw/cirmount/gloria/). At the broad scale of environmental modeling and climate change, global change models (GCMs) predict that alpine areas will experience higher levels of temperature increase than global averages (Theurillat and Guisan 2001, Beniston 2005). Concerns about the potential impacts of higher temperatures on high-elevation communities have led to a variety of studies, including experimental warming manipulations to look at impacts of increased temperatures on alpine and subalpine plant phenology and community structure (Harte et al. 1995, Price and Waser 2000, Klein et al. 2005). As useful as GCMs can be, they operate on grid scales of kilometers along horizontal axes and tens of meters along the vertical. Thus they are most effective at heights well above the soil, excluding the plant canopy levels where alpine microclimate affects biological and local ecosystem processes. For alpine ecosystems a range of factors complicate the straightforward interpolation from macroclimate to microclimate (Wundram et al. 2010). Boundary layer dynamics at the ground surface complicate predictions because the complexity of interactive factors such as wind shear, pressure gradients, and energy balance cause the environment to become decoupled from free-air conditions above the ground surface. The topographic heterogeneity of alpine habitats creates a fine pattern of thermal microhabitat conditions at a scale of centimeters. The magnitude of these temperature differences is greater than the range of warming scenarios over the next century in IPCC projections (Graham et al. 2012). If short dispersal and establishment is possible for alpine plants, then fellfield habitats may offer significant buffering from global Alpine Ecosystems   629

warming because of the mosaic of thermal microclimates present. However, we know very little about the significance of moisture availability for plant distributions or its interactions of temperature and soil moisture (Winkler 2013). GCM models are able to make temperature predictions with far more confidence than precipitation ones, leaving open the question of moisture availability. The roles of microclimate in alpine habitats suggest that models predicting upslope movements of species under increasing temperatures might not be entirely realistic and that sufficient microclimate heterogeneity might exist to slow species range shifts.

Summary Alpine ecosystems are typically defined as those areas occurring above treeline, but alpine ecosystems at a local scale can be found below this boundary for reasons including geology, geomorphology, and microclimate. The lower limit of alpine ecosystems, the climatic treeline, varies with latitude across California, ranging from about 3,500 meters in the southern California mountains and southern Sierra Nevada to 3,200 meters in the Yosemite region, 3,000 meters near Donner Pass, 2,800 meters at Lassen Peak, and finally 2,700 meters on Mount Shasta. Alpine ecosystems extend beyond the typically envisioned high-elevation open slopes and summits of cold-adapted shrubs and herbs to include as well lithic environments of cliffs, talus fields, boulder fields and rock glaciers; permanent and persistent snow and icefields, including glaciers; and various water bodies such as streams, tarns, and large lakes. Alpine ecosystems provide severe physiological stresses for both animal and plant populations. These environmental stresses in California include low winter temperatures, short growing season, low nutrient availability, high winds, low partial pressures of CO2, high UV irradiance, and limited water availability under summer drought. The alpine regions of California typically experience a Mediterranean-type climate regime with dry summers and precipitation heavily centered on the winter months. This regime differs significantly from that present in most of the continental alpine habitats of the world, where summer precipitation predominates. At the upper treeline in the Sierra Nevada about 95% of annual precipitation falls as winter snow, with much of this accumulating during regular winter during a very small number of storms separated by long, dry intervals. This pattern produces extreme interannual variability in precipitation and water availability. Alpine plant communities are dominated by herbaceous perennials (broad-leaved herbaceous perennials, mats and cushions, graminoids, and geophytes) that form the dominant community cover. Also present with lower species richness are low shrubs and semiwoody subshrubs. Other plant life forms such as taller woody shrubs and annuals are rare. Alpine ecosystems support a low diversity of resident mammal species, but many others use the alpine environment occasionally or seasonally. Notable are large herbivores such as mule deer and desert and Sierra Nevada bighorn sheep that forage in the alpine zone in summer. Many more small and midsized mammals occur in the alpine zone, with yellow-bellied marmots and pikas commonly seen in such habitats. Alpine ecosystems are predicted to experience strong levels of temperature increase from global warming globally but will likely be most impacted by indirect effects such as declining snowpack, earlier spring runoff, and earlier growth and flowering phenology. 630  Ecosystems

Acknowledgments We thank the staff of Sequoia and Kings Canyon and Yosemite National Parks for making their data available; the staff of the White Mountains Research Station for their support; and the volume editors for their guidance.

Recommended Reading Billings, W. D. 1974. Adaptations and origins of alpine plants. Arctic and Alpine Research 6:129–​142. Bowman, W. D., and T. R. Seastedt, editors. 2001. Structure and function of an alpine ecosystem: Niwot Ridge, Colorado. Oxford University Press, Oxford, UK. Körner, C. 2003. Alpine plant life: Functional plant ecology of high mountain ecosystems. Springer Verlag, Berlin, Germany. Millar, C. I. 2012. Geologic, climatic, and vegetation history of California. Pages 49–68 in B. G. Baldwin, D. Goldman, D.J. Keil, R. Patterson, T.J. Rosatti, and D. Wilken, editors. The Jepson Manual: Higher plants of California. Second edition. University of California Press, Berkeley, California. Nagy, L., and G. Grabherr. 2009. The biology of alpine habitats. Oxford University Press, Oxford, UK. Rundel, P. 2011. The diversity and biogeography of the alpine flora of the Sierra Nevada, California. Madroño 58:153–​184. Sawyer, J. O., and T. Keeler-Wolf. 2007. Alpine vegetation. Pages 539–​ 573 in M. Barbour, A. Schoenherr, and T. Keeler-Wolf, editors. Terrestrial vegetation of California. Second edition. University of California Press, Berkeley, California.

Glossary Atmospheric river  A narrow atmospheric band of concentrated moisture that can cause extreme precipitation events at midlatitudes. Atmospheric rivers affecting the coast of western North America are informally referred to as “pineapple express” phenomena. Cirque  An amphitheater-shaped basin below a mountain peak carved by glacial action. Fellfield  Alpine habitat with shallow, stony, and poorly developed stony soils. Felsenmeer  Exposed rock surface that has been broken up by frost action so that much rock is buried under a cover of angular, shattered boulders. Freeze-thaw cycle  A weathering cycle in which water seeps into cracks and then freezes and expands, promoting breakdown of the rock. Graminoid  Having a grasslike form of growth, as in the Poaceae, Cyperaceae, and Juncaceae. Karst topography  A region where the terrain has been impacted by the physical and chemical weathering of carbonate rocks such as dolomite and limestone. Krummholz  A stunted and often deformed growth of trees at the treeline limit. Little Ice Age  A global period of cooling, extending from about 1550 or earlier to about 1850, marked by a significant expansion of glaciers. Moraine  An accumulation of unconsolidated glacial debris. Nival  Growing with or under snow; also used to connote an upper alpine region continuously under snow or ice throughout the year. Paternoster pond  A glacial pond or lake connected to multiple others in a string by a single stream or a system of linked streams.

Periglacial  Describes any place where geomorphic processes related to freeze-thaw cycles of water occur. Permafrost  Permanently frozen subsurface layers of soil. Rock glacier  Geomorphological landforms consisting of angular rock debris frozen in interstitial ice. Talus field, talus slope  Describes a landform of jumbled rock debris lying with an inclination up to the maximum angle of repose. Tarn  A small lake at the base of a cirque formed by past glacial action.

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THIRT Y

Deserts JAYNE BELNAP, ROBERT H. WEBB, TODD C. ESQUE , MAT THE W L . BROOKS, LESLE Y A . DEFALCO, and JAM ES A . M AC MAHON

Introduction The deserts of California (Figure 30.1) occupy approximately 38% of the state’s landscape (Table 30.1) and consist of three distinct deserts: the Great Basin Desert, the Mojave Desert, and the Colorado Desert, which is a subdivision of the Sonoran Desert (Brown and Lowe 1980). The wide range of climates and geology found within each of these deserts produce very different vegetation communities and ecosystem processes and therefore different ecosystem services. In deserts, extreme conditions such as very high and low temperatures and very low rainfall result in control by abiotic ­factors—​climate, geology, geomorphology, and soils—​of the composition and function of ecosystems, including plant and animal distributions. This situation is in contrast to ecosystems where wetter and milder temperatures occur and the dominant organizing factors are often disturbance (such as fire, landslides, and floods) and biotic interactions (such as competition, herbivory and predation) (Clark 1991). Despite their harsh conditions, deserts are home to a surprisingly large number of plants and animals. Deserts are also places where organisms display a wide array of adaptations to the extremes they encounter, providing some of the

best examples of evolution by natural selection (MacMahon and Wagner 1985, Ward 2009). Humans have also utilized these regions for thousands of years, despite the relatively low productivity and harsh climates of these landscapes. Unlike much of California, most of these desert lands have received little high-intensity use since European settlement, leaving large areas relatively undisturbed. Desert landscapes are being altered, however, by the introduction of fire following the recent invasion of Mediterranean annual grasses (D’Antonio and Vitousek 1992). As most native plants are not adapted to fire, they do not recover, whereas the non-native grasses flourish. Land uses such as energy exploration and development, recreational use, and urban development are rapidly increasing as well, and because desert lands are slow to recover, such disturbances will alter these landscapes for many years to come. This chapter begins with a brief description of where the different deserts of California are located and their dominant vegetation communities. The abiotic factors that define these deserts, and how these factors control vegetation and thus also animal distributions, are examined next. After this 635

TA B L E 3 0 .1 Areas of the California desert units

Land mass

Area (km2)

California

423,970

100.0

Mojave Desert

129,523

30.5

26,317

6.2

Modoc Plateau

4,000

0.9

Owens Valley

1,936

0.5

Mono Lake basin

2,030

0.5

Colorado Desert

Total

161,775.7

Percentage area

38.2

Source: Calculated by RH Webb. note: Modoc Plateau, Owens Valley and Mono Lake Basin are zones of the California Great Basin Desert.

s­ ection, ecosystem processes and iconic species of these deserts are discussed, followed by a concluding section on the future of these landscapes. This last section focuses primarily on the Mojave Desert, as it is both the largest California desert and the focus of most existing research on California deserts.

teau, this region has high soil diversity and hotspots of plant endemism. The southeastern Mojave Desert grades into the Sonoran Desert and contains, across the border in Arizona, the co-occurrence of Joshua tree (Yucca brevifolia) and saguaro (Carnegiea gigantea). The south-central Mojave Desert contains large, iconic Joshua tree forests. The central Mojave Desert encompasses the lower reaches of the internally drained Mojave River, including a series of relatively low-elevation playas (dry lakes). The separation of the central Mojave Desert from the Great Basin and Colorado Deserts, along with its relatively moderate topography, results in less ecological variability than the other subregions, including the fewest number of vascular plant species (458 species) (Rowlands et al. 1982, Bell et al. 2009, Wood et al. 2012). Lastly, the western Mojave Desert has intermediate plant species richness (663 species) (Rowlands et al. 1982). The Colorado Desert occurs in southeastern Arizona and southwestern California, extending southward into Mexico. It is dominated by creosote (Larrea tridentata) and also has abundant succulent plants, notably cholla (Cylindropuntia spp.) (Pinkava et al. 2001) and saguaro, and leguminous trees, particularly paloverde (Parkinsonia microphylla) and ironwood (Olneya tesota). The Salton Sea and its predecessor, Lake Cahuilla (one of the last and largest Pleistocene lakes, which evaporated approximately four hundred years ago), dominate this desert, and much of its substrate is fine-grained alluvium derived from the combination of runoff from the Colorado River and the Transverse Ranges.

Biogeography The smallest desert region in California is the Great Basin Desert, which occurs in two distinct subregions in California (see Figure 30.1). The southern subregion occurs east of the Sierra Nevada in the Mono Lake Basin and the Owens Valley. The northern section includes the Modoc Plateau east of the Cascade Range in northeastern California (Sawyer et al. 2009). Only the southwestern corner of the Great Basin Desert occurs within California, with the rest in Idaho, Nevada, Oregon, Utah, Washington, and Wyoming (see Table 30.1). It is dominated by big sage (Artemisia tridentata), single-leaf pinyon pine (Pinus monophylla), and one seed juniper (Juniperus monosperma). The Mojave Desert borders on Utah, Arizona, and Nevada and is the largest desert in California, occupying 30.5% of the state. The Mojave Desert is unique among the North American deserts for its extensive display of winter annual plants and its many endemic plants and animals. Extreme variations in climate and topography, coupled with active geologic activity, are the fundamental ecosystem drivers associated with this high diversity of plant species and ecosystem properties. Six distinct subregions of the Mojave Desert have been defined on the basis of seasonal precipitation patterns, landscape types, and floral diversity (Figure 30.2; Webb, Heaton et al. 2009). The northern Mojave Desert, with its climatic extremes, has the highest plant species richness in the Mojave Desert (1,025 species) (Rowlands et al. 1982). The eastern Mojave Desert contains most of the high elevations in the Mojave Desert. Combined with higher summer rainfall and exposed outcrops of rock strata like those of the Colorado PlaPhoto on previous page: A typical Mojave Desert setting, with the New York Mountains in the background. The foreground is covered with biological soil crusts. Photo: T. E. Esque. 636  Ecosystems

Drivers of Ecosystem Processes in Space and Time Desert ecosystems are complex entities (Figure 30.3). Desert ecosystem processes are controlled mainly by the abiotic factors of climate, geology, soils, geomorphic setting, and various disturbances. These ecosystem processes in turn set the stage for the plants and animals that inhabit these landscapes.

Climate The most defining characteristic of deserts is high spring and summer temperatures accompanied by very low and variable precipitation. These factors result in scant vegetative cover that is generally short in stature. Although climatic regimes are important in determining the structure and function of all ecosystems, they are an especially dominant force in deserts, where water availability is the ultimate determinant for plant and animal survival. The availability of soil moisture for plants and soil biota, which in turn provide food and habitat for desert animals, is determined by temperature, precipitation regimes (the overall amount of precipitation and the type and seasonality of events), and soil characteristics. Lower temperatures, which allow for longer retention of soil moisture, occur in fall and winter, at higher elevations, and on north and east slope aspects. The type of precipitation is also important. Compared to summer convective storms, winter frontal storms are generally less spatially variable with more gentle rainfall, allowing more water to infiltrate more deeply into the soil and over a larger area. If precipitation events are too short or too small, the resulting soil moisture stays at the surface and quickly evaporates. Seasonality, or the timing of precipitation relative

Non-desert

FIGURE 30.1 California’s desert regions. Source: Data from Cal Fire, Fire Resource and Assessment Program (FRAP). Map: P. Welch, Center for Integrated Spatial Research (CISR).

CA NE LI VA FO D RN A IA

115º

117º

NTTR UTAH ARIZONA

NTS NTTR

Northern Mojave

Eastern Mojave 15

DVNP

Las Vegas LMNRA

NAWSN

36º

GCNP

NAWSS NTC

Western Mojave

Central Mojave

EAFB

Barstow

South-Central Mojave

Kingman MNP

40

South-Eastern Mojave

MCAGCC

34º

N

0

MCAGCC-Marine Corps Air Ground Combat Center MNP-Mojave National Preserve JTNP Los Angeles NAWSN-Naval Air Weapons Station North DVNP-Death Valley National Park 10 NAWSS-Naval Air Weapons Station South EAFB-Edwards Air Force Base NTC-National Training Center (Ft. Irwin) GCNP-Grand Canyon National Park NTS-Nevada Test Site 40 80 JTNP-Joshua Tree National Park km km LMNRA-Lake Mead Natl. Rec. Area NTTR-Nevada Test and Training Range

FIGURE 30.2 Map of the Mojave Desert showing the six subregions. Source: Webb, Fenstermaker

et al. 2009.

to temperature, also matters because rain falling when temperatures are high evaporates more quickly than rain or snow falling when temperatures are low. In addition, some plants can only utilize precipitation that occurs in specific seasons. For example, many perennial grasses and succulent plants only occur in high abundance where significant summer rainfall occurs (Comstock and Ehleringer 1992, Ehleringer 2001). All three California deserts average less than 250 millimeters annual precipitation because of the rainshadow effect of the Sierra Nevada and Transverse Ranges. The Great Basin Desert has the lowest potential evapotranspiration (PET) of the three deserts due to its lower overall temperatures, which reflect its higher elevation and more northerly position in the state. Freezing temperatures occur there for four to five months a year. Because most precipitation falls during the cold winter, evaporation and transpiration are limited, allowing most of the water to infiltrate deeply into the soil. For example, at Bridgeport, northwest of Mono Lake, threequarters of the average precipitation of 238 millimeters falls between November and March (Western Regional Climate Center 2013a). The Mojave Desert is warmer than the Great Basin Desert and has four distinct precipitation zones: (1) low winter/low summer (average=113 mm yr-1, 70% winter, October–​April); (2) moderate winter/moderate summer (177 mm yr-1, 64% winter); (3) high winter/low summer (153 mm yr-1, 82% winter); and (4) high winter/high summer (271 mm yr–​1, 638  Ecosystems

61% winter) (Tagestad et al. in press). Interannual variability is high and has ranged from 47 to 587 mm yr–​1since records have been kept (Hereford et al. 2006). Like the Great Basin, the Mojave Desert is subject to freezing temperatures, which approach –​9 °C in many parts of the desert, but for a shorter time period. Thus the PET of the Mojave Desert is higher than the Great Basin’s. The Colorado Desert in southern California is the warmest and driest region in North America, reflecting its generally low elevation and southerly position. This desert seldom experiences freezing temperatures, and its PET is extremely high. Rainfall is strongly biseasonal, with average winter rainfall only slightly above average summer rainfall. For example, El Centro has a mean annual precipitation of 67 millimeters, 60% of which falls in the winter months. However, summerfall incursions of tropical moisture from the Pacific Ocean can make these averages meaningless because of extreme precipitation events associated with summer monsoonal rains. These monsoonal rains are highly intense events of short duration, resulting in high runoff and little contribution to soil moisture. The average maximum temperature in this region exceeds 38°C from June through September (Western Regional Climate Center 2013b), causing extremely high PET (greater than 2,500 mm yr–​1). Given that precipitation in desert regions is highly variable in both space and time, extreme drought and wet years are important ecosystem drivers because they can result in largescale and long-term effects on ecosystem structure through

SOIL SYSTEM Soil structure, soil texture, parent material Soil water

Soil biota, C, and nutrients

Water Pulse Storage

Depth Amount Retention

Moisture and nutrient cycling

VEGETATION AND BIOCRUSTS Vegetation

Biotic life zones

structure, cover, C, nutrients, species composition

Rooting, burrowing, decomposition

ANIMALS

Food, cover, structure

Forest Woodland

Shrubland

Biocrusts

C, N inputs; soil stability, moisture, fauna

Herbivory, pollination, seed dispersal

Climate change and air pollution

CLIMATE

Season, magnitude, frequency, duration of precipitation, temperature, wind events

Soil disruption and erosion

Destruction of vegetation and soil crusts, invasive plants, altered fire regime FIGURE 30.3 Conceptual model of the Mojave Desert ecosystem. Source: Modified from Chung-MacCoubrey et al. 2008.

episodes of plant death, establishment, or growth (Noy-Meir 1973, Ehleringer 2001). Therefore a great deal of effort has gone into developing a framework to understand the effects of precipitation events on various ecosystem processes. The pulse-dynamic and pulse-reserve are the most widely accepted conceptual models in use today (Noy-Meir 1973, Reynolds et al. 2004). In its simplest form, the pulse-dynamic model posits that most desert productivity (such as germination, growth, seed production) is a direct response to discrete rainfall events. The pulse-reserve model suggests that some of this productivity is carried over as reserves, such as seeds, litter, and nutrient or carbon (C) storage in roots. Recently, these models have included an emphasis on the role of nutrients as a driver of productivity, as it has been increasingly recognized that relatively infertile desert soils can quickly limit plant productivity when water is not limiting, such as

in years of high precipitation or frequent precipitation pulses (Ogle and Reynolds 2004). A pulse of productivity results in an increase in consumers (animals) and their consumption rates. Animals are not only important consumers of desert productivity; they also alter ecosystem processes by redistributing resources such as seeds around the landscape, disturbing the ground through digging, and preying on other animals (Whitford 2002). They also are much better than plants at amortizing the gains and losses experienced over a series of individual pulse events (Schmitz 2009). The pulse-dynamic and the pulse-reserve models continue to develop as other authors propose needed additions or alterations (Collins et al. 2014). For instance, Ogle and Reynolds (2004) noted that the effects of antecedent water and processes like the ability of plants to alter their rooting habits and to delay their response to rainfall need to be included in Deserts  639

these models. Schwinning et al. (2004) and Schwinning and Sala (2004) have shown there is a hierarchy of soil moisture pulses and ecological responses to these pulses that needs consideration. Specifically, because most precipitation events in deserts are less than 5 millimeters and only stimulate activity in organisms at the soil surface (Pointing and Belnap 2012), larger pulses of precipitation that penetrate deeper into the soil can trigger a greater variety of longer-lasting events, including stimulation of plant productivity. As a result, soil and plant processes in deserts may be temporally or spatially asynchronous or synchronous. For example, small wetting events result in nutrient input and accumulation in surface soils, but these small events are often too small to translocate nutrients to vascular plant roots (asynchrony). On the other hand, large wetting events stimulate nutrient inputs and transformations at the surface at the same time as translocations of these nutrients to vascular plants occur (synchrony).

Geology, Soils, and Geomorphic Setting After climate, other abiotic factors become important in determining the composition and function of desert ecosystems. Of these, geology, soils, and geomorphic setting are the most dominant, as they all have a strong influence on water and nutrient availability.

GEOLOGY

California’s deserts are all on the western margin of the Basin and Range Province. The geologic framework of these deserts consists of numerous mountains, generally north-south trending, created by fault-driven rock displacement upwards, and separated by fault-delineated, depressed blocks of land called grabens (Hall 2007, Cooke et al. 1993). This type of regional geology promotes internal drainage into valley lakes and playas rather than the ocean. This geology also provides the elevation gradients that are so influential on these desert ecosystems. Two of California’s most striking elevational gradients are found in Owens Valley of the Great Basin Desert and in Death Valley of the Mojave Desert, where desert floor to mountaintop elevation differences are the greatest in the continental United States, exceeding 3,500 meters. The northern Great Basin Desert is the least tectonically active of the California deserts. In contrast, the Mojave Desert is tectonically active with numerous active faults, and geology of the Colorado Desert is strongly controlled by the San Andreas fault system. Parent material matters greatly in determining what plant communities, and therefore what animal communities, can thrive in a location. This is because parent material controls nutrient content and the infiltration and retention of water in soils derived from them. The northern Great Basin Desert in California is dominated by volcanic rocks associated with the Cascade Mountains, but the southern Great Basin is mostly granitoid and associated metamorphic rocks from the Sierra Nevada and Inyo Mountains. In the Mojave Desert, approximately 55% of the bedrock is sedimentary, and most of these rocks are Paleozoic limestones, dolomites, and shales. Many mountains in the Mojave Desert have granitoid rocks at their cores, and 34% of the exposed bedrock in the Mojave Desert is plutonic. In contrast, the Colorado Desert contains only 38% consolidated bedrock, with the remainder consisting of Lake Cahuilla sediments and alluvial fans of Quaternary age. 640  Ecosystems

SOILS

As with climate and geology, desert soils have a dramatic effect on what plants and ecosystem processes occur in a given location. The wide variety of igneous, sedimentary, and metamorphic parent materials that are exposed in the California deserts results in a large range of soil types, which promotes high vegetative diversity. Desert soils possess various properties that are very different from soils in mesic regions and have a profound influence on ecosystem processes. One of the most important differences between desert soils and those in wetter regions is what is referred to as the “inverse texture hypothesis” (Sala et al. 1988). Most desert soils contain substantial amounts of sand and rocks, which when combined with low inputs of plant litter, soil biotic activity, and nutrient-holding capacity, result in coarser soils that are generally less fertile than soils containing more fine particles and organic material (Ward 2009). However, coarser soils allow water to infiltrate below the evaporative zone more rapidly than finer soil, allowing them to retain soil moisture longer. Because water is more limiting than nutrients in desert ecosystems, coarser textured soils therefore generally support higher plant and animal productivity than finer-textured soils (Sala et al. 1988). This is in contrast to wetter ecosystems, where nutrients are often more limiting than water, rendering finer, more fertile soils more productive than coarse soils. Other features distinguish desert soils from those found in wetter regions. The low weathering rates of rocks in deserts can result in soils with physical and chemical characteristics very similar to the parent material from which they were derived (Jenny 1941). Most desert soils are very old, and older soils in deserts generally support less plant productivity and species diversity than younger soils. Several features can develop with age in these settings that have a strong negative influence on plant productivity and ecosystem processes. These features include desert varnish, vesicular horizons, desert pavement, and subsurface clay and carbonate horizons (reviewed in Belnap et al. 2008). Dark desert varnish often covers the rocks found on the soil surfaces of hot deserts such as the Mojave and Colorado Deserts (Springer 1958). Its color comes from iron and manganese oxides. The origin of the varnish is still under debate, with the main theories being that the iron and manganese oxides are (1) deposited by wind onto wetted rocks, (2) leached out from the interior of the rocks, and/or (3) bioaccumulated by bacteria and fungi. All three processes can play a partial role in the formation of desert varnish. Vesicular horizons form in older soils when fine soil particles transported from distant sources by the wind and deposited over thousands of years create a thin layer of silt at the soil surface (McFadden and Knuepfer 1990, McDonald 1994). As air rises through moist soils, it can become trapped in the silt layer, creating layers of air pockets, or vesicles. Once formed, these vesicles restrict the ability of rainwater to infiltrate into the soil, thereby restricting the growth of vascular plants. Desert pavement also forms on very old soils, which are generally of Pleistocene age. After many studies, most researchers agree that desert pavement results from two main forces: the erosion of surface soils exposing subsurface rocks (e.g., Wainwright et al. 1995) and/or subsurface rocks rising to the surface due to wind-blown silts accumulating under them (e.g., McFadden et al. 1987). Silt particles move down in soils in two ways. First, after the silt is deposited, the infiltration of large rain events car-

FIGURE 30.4 Geomorphic surfaces in the Mojave Desert. At the top of the photo are steep hillslopes, cut by water channels. In the middle of the photo are mid- and low- elevation alluvial fans that receive materials from above. Fire scars can be seen on a lower- and mid-elevation alluvial fan. Note the patchy aspects of the fire, and the way geomorphic surfaces (higher terraces versus washes) influenced what areas were burned. Photo: Matthew L. Brooks.

ries silt downwards into the soil. Second, the shrinking and swelling of soils accompanied by wetting and drying cycles can rearrange the soil particles. In desert pavements the rocks at the surface are closely packed together, covering the soil surface almost entirely. Because these rocks are often covered with desert varnish and extremely close together, they create very high temperatures at the soil surface. Vesicular horizons also form under these rocks (Springer 1958). The combined presence of the hot rocks and vesicular horizons restrict the entry of water. Desert pavements are therefore generally devoid of plant cover, and their microbial activity is extremely restricted. Subsurface clay or carbonate horizons also are often present in hot deserts (McFadden et al. 1998). These horizons can effectively restrict water infiltration and inhibit root penetration. The same is true for subsurface cemented carbonate layers, which can accumulate over 200 kg C m-3, making them equal to peat bogs as carbon stores (Monger 2006). While carbonate layers can restrict root access to deep water, in some situations they may also store water that is then made available to plants during drought (Duniway et al. 2010). Episodic erosional events resulting from flooding or landslides often disrupt the process of soil formation. These episodic events can move “foreign” materials that occur upslope and deposit them onto of different surface types found below. New stream channels are often cut through different aged surfaces, exposing them. For instance, coarse material moved downslope can bury finer-textured deposits or desert pavements, creating an entirely new type of soil in which different plants can germinate and grow compared to what was originally found there. As a result, multiple ages and types of soils from different parent materials can occur side by side, creating the interesting mixture of plant and animal communities found in many settings throughout the California deserts.

Important Geomorphic Settings and Associated Vegetation In deserts, the interaction of climate, soil formation, and geomorphic processes (e.g., landslides, overland flow, eolian [wind-blown] deposition) creates a mosaic of heterogeneous geomorphic units, highly variable in space and often in time, that determine ecosystem processes and thus where plants

and animals live (Figure 30.4; Webb et al. 1988, McAuliffe and McDonald 1995, Hamerlynck et al. 2002). In the California deserts, the spatially dominant geomorphic units are playas, eolian features, alluvial fans, and hillslopes. All of these units are shrub-dominated, with plant cover ranging from less than 5% to approximately 35%. Water features, while not spatially extensive, are very important to particular communities of plants and animals. During the wet periods of the Pleistocene, extensive systems of lakes and rivers existed within the Mojave and Great Basin deserts. Most of these were internally drained, with no outlet to the sea, with the exception of the far southernmost part of the Mojave, which drained into the Colorado River. These waterways were extensive. For example, what was once Lake Manly in Death Valley was fed by the Owens River via Lakes Searles and Panamint, as well as the Amarogosa and Mojave Rivers. As the climate dried, the lakes became what are now playas. In very wet years runoff sometimes reaches these old Pleistocene salt-encrusted lake beds. There are two fundamental types of playas: wet and dry (Reynolds et al. 2007). Wet playas have shallow groundwater levels and perennial evaporation, resulting in high salt levels at or near the surface that often completely prevent plant growth; these playas tend to be the termini of larger regional rivers or Pleistocene drainage systems and are often large dust sources. Dry playas have deep water tables and are only periodically inundated, resulting in less salt accumulation in the soils. Dry playas generally also have a hard clay surface that produces a large amount of dust when disturbed. They also often have a local source of water and sediment and can sometimes support limited plant cover by salt-tolerant species. Small, local sand dunes often form around the playa edge and can also support salt-tolerant species. These species typically include four wing saltbush (Atriplex canescens), shadscale (Atriplex confertifolia), and pickleweed (Salicornia virginica). Various cholla cacti occur sporadically in this and other vegetation communities and provide infrequent, but important, nesting substrate for a variety of birds. Intermittent flooding of some playas can stimulate the hatching of an invertebrate community of fairy shrimp, tadpole shrimp, and amphipods that can be extremely abundant for short periods. These hatches may attract large accumulations of water birds and, as a result, can be important for migration stopovers (Carlisle et al. 2009). Large, eolian sand dunes are common where there is an Deserts  641

abundant sand supply, including areas downwind of major rivers, washes, old pluvial lake deposition, and some playas (e.g., Clarke and Rendell 1998). Active or destabilized dunes can undergo blowouts caused by rapidly erosive winds that can allow sand migration into areas without previous eolian deposits. Dunes store water and thereby support unique plants, including deep-rooted ones such as mesquite (Prosopis spp.), which can anchor dunes. Large dunes in the Mojave Desert include the Eureka, Death Valley, Panamint, and Kelso Dunes. In the Colorado Desert the main eolian features are the Coachella Valley and the Algodones Dunes. All these dune systems have endemic and rare plant and animal species (e.g., Skinner and Pavlik 1994). Many dune systems are now also important recreational sites. Furthermore, many of these dune systems are important archeological sites because they were occupied by Native Americans during the Holocene, when extensive lake systems filled many of the valleys. The mesquite trees’ pods were a staple food source for aboriginal people (Fowler 1999). The mesquite bosque (woodland) near Death Valley’s Furnace Creek has long been highly prized for its sweet pods and is currently used by the Timbisha Shoshone Tribe (Timbisha Shoshone Tribal Elders, pers. comm.). However, the Furnace Creek mesquite stand does not appear to be reproducing, possibly due to water diversion. Alluvial fans are by far the most common geomorphic surface in the California deserts. These units are composed of loose rock and coarse particles (e.g., sand) moved by fastmoving mountain streams and overland flow events. Loose rock and sand are transported downhill until a change in slope slows the flow, and this material is deposited in a fanlike shape. Soils farther from the mountain source are finer in texture than upslope soils because the finer, lighter particles stay suspended longer and are transported farther. Because most surface-water drainage channels and debris flows cross each other and disrupt soil formation on these fans, soils of different ages are exposed, supporting a patchwork of different plant and animal communities. Although most surfaces are of Holocene age (the last eleven thousand years), many older surfaces (e.g., Pleistocene desert pavements) are also present. Dominant shrubs and animals change with elevation and latitude (Benson and Darrow 1981, Turner 1994). In the Mojave and Colorado Deserts the majority of lower-elevation fans are dominated by creosote, with the understory vegetation changing as the age of the soil surface changes. The most common subdominant plant is white bursage (Ambrosia dumosa) in the Mojave Desert and triangle leaf bursage (Ambrosia deltoidea) in the Colorado Desert. In both deserts these dominant species can be surrounded by a relatively rich mixture of other shrubs and annual species (Turner 1994). Alluvial fans, because of their great spatial extent, support the greatest annual productivity in all three California deserts. Hillslopes and upper alluvial fans occur as elevations rise, and the dominance of creosote gives way to that of blackbrush (Coleogyne ramosissima). The lower end of the blackbrush belt is where the visually dominant Joshua trees are found. Above approximately1,250 meters, various juniper (Juniperus spp.) and pinyon (Pinus spp.) trees replace the Joshua trees. These trees increase in density until the shrubland vegetation gives way to woodlands, dominated by trees with a shrub understory of species like mountain mahogany (Cercocarpus spp.) and buckbrush (Ceanothus spp.). The most northerly areas of the California desert are occupied by Great Basin sage scrub, where some of the chenopod shrubs dominate in association with several sagebrush species (Artemisia spp.). 642  Ecosystems

Water features including lakes, springs, and rivers also occur in the California deserts. The largest is the Salton Sea, which formed in 1905 when the Colorado River was diverted into its ancestral drainage patterns that once filled Pleistocene Lake Cahuilla (see Chapter 32, “Lakes”). Once a wet playa, the Salton Sea now is sustained by agricultural drainage from the Imperial Valley. However, drought conditions threaten this supply. Water diversion to Los Angeles has changed the nature of the larger California desert lakes. Owens Lake occasionally is filled with runoff from the Owens River, but typically it is a playa due to water diversions. Mono Lake has decreased considerably in size because of similar human-made water diversions. In contrast, larger perennial lakes occur on the Modoc Plateau in the north, where water development has not significantly decreased inflows. Springs are common in the mountain ranges and upper parts of alluvial fans in the California deserts. Most springs have small catchment areas within mountain ranges; as a result, their flows increase and decrease with changes in precipitation and snowmelt. A few larger springs occur, often the end points of large groundwater flow systems, discharging water tens of thousands of years old and therefore not responsive to climate fluctuations. Unique riparian ecosystems support many endemic herbaceous plant and invertebrate species around these isolated springs. Groundwater development, combined with the establishment of non-native species including saltcedar (Tamarix spp.), threaten or have already irreversibly affected many of these small ecosystems (Hultine and Bush 2011). Several large rivers occur in the deserts of California, with the Colorado River forming the southeastern boundary of the California desert region. Before being diverted for human use, the Owens, Mojave, and Whitewater Rivers transported water and sediment into the desert, depositing fertile, finegrained sediments on alluvial plains; raising groundwater levels; and increasing ecosystem productivity. The fine-grained sediment deposited by these rivers is also one of the largest sources for the large sand dune systems found in the Colorado and Mojave Deserts. However, because dams trap sediment, water diversion threatens sand supplies needed to maintain and form the dunes. Surface-water dependent (riparian) vegetation and animals are concentrated along the rivers that cross the deserts, especially where groundwater levels are high as the result of perched water tables or subsurface geology that locally forces groundwater to the surface. Xero-riparian plant communities also occur along these rivers and along intermittently flowing washes. This habitat type supports more vegetation than upland desert communities, but less so than riparian areas, and contains a mixture of plants that both do and do not require surface water. Xero-riparian areas are fairly widespread in the California deserts. Because they modulate extreme environmental conditions and provide predator protection, nesting substrates, and food for many animal species, they are occupied by the greatest number and diversity of animal species in the deserts (Carlisle et al. 2009). They also provide important travel routes for migrant birds.

Ecosystem Components and Processes All ecosystems have basic processes in common, including production, decomposition, and carbon (C) and nitrogen (N) cycling. The drivers and rates of these processes, however,

are often substantially different in deserts than in mesic regions.

A

Biological Soil Crusts A unique feature of desert regions is the presence of biological soils crusts (biocrusts) (Figures 30.5 and 30.6; Belnap and Lange 2003). Biocrusts are communities of micro-organisms (cyanobacteria, bacteria, fungi, and green algae), macroscopic lichens and mosses, and microarthropods that occur within the top centimeter of the soil surface. As plants are often widely spaced in deserts, biocrusts can cover up to 70% of the soil surface. The biodiversity found in biocrusts often far exceeds that of the vascular plant community in which they are embedded. There are hundreds to thousands of species in biocrusts, whereas most dryland plant communities contain fewer than one hundred species. Biocrusts play many essential roles in deserts, and their influence increases with their biomass (Housman et al. 2006), which is positively related to precipitation and also controlled by parent material. For instance, in the California deserts, biocrust development is greatest on soils derived from gruss-weathered granites (those that weather directly into grain-sized particles and silt), followed by soils derived from sedimentary, other igneous, and finally metamorphic rocks (Belnap et. al 2014). This is likely due to the differences in the texture and nutrients found in the soils weathered from these rocks. Soils with higher levels of silt also have higher biocrust development (Belnap and Lange 2003), likely because silt can increase the water-holding capacity of the soil. Because biocrusts often completely cover the soil surface, they mediate most inputs to and from desert soils, including gases, light and heat, water, dust, plant litter, and seeds. They thus play a central role in the functioning of desert ecosystems (see Figure 30.6; Belnap et al. 2003). All biocrust organisms are integral in the formation and stabilization of soils and are believed to have played this role since they first appeared on land seven hundred million to one billion years ago. These organisms accelerate soil weathering by altering soil pH through secretion of acids, calcium, and hydroxide. Rock and soil weathering also increase because biocrust organisms retard evaporation of soil moisture, increasing the length of time these materials are wet. Biocrusts are very effective at reducing or eliminating erosion of soil particles, as cyanobacterial and fungal filaments bind soil particles together (see Figure 30.5, top photograph). This soil-binding action does not depend on the presence of living filaments; layers of abandoned sheaths build up over long periods of time and can still be found clinging to soil particles at depth in the soil. Lichens and mosses also protect the soil surface from wind and water, reducing soil erosion that in the absence of biocrusts is substantial because of the deserts’ low vegetative cover. The external morphology of biocrusts influences resource retention as well. In the Great Basin Desert and higher elevations of the Mojave Desert, biocrusts roughen the soil surface as a result of frost-heaving upwards and differential erosion downwards. This roughening enhances water infiltration and the capture of dust, seeds, and other materials crossing the soil surface. In hotter areas where soils do not freeze, however, the presence of biocrust can flatten the soil surface, leading to less water infiltration and an accelerated loss of local materials. Biocrusts also play other critical element-cycling roles in deserts. Biocrusts darken the soil surface, warming it through

B

FIGURE 30.5 Biological soil crusts. Photos: Jayne Belnap (top) and Todd Esque (bottom). A A scanning electron micrograph of a biocrust (x 90).

Cyanobacterial filaments wind among the sand grains, linking them together and conferring great soil surface stability. B A close-up of the lichen-dominated biocrusts found on grussy

granite soil surfaces. Note the tarantula in the photo.

Plant community composition distribution

Carbon fixation Nitrogen fixation Phosphorus availability Dust capture

BSC

Animal habitat

Exo-polysaccharides Soil aggregates Soil stability

Soil fertility Soil moisture Plant nutrition Plant biomass

Animal forage

FIGURE 30.6 A conceptual model showing the many crucial roles biological soil crusts (BSC) play in the desert ecosystems of California.

Deserts  643

increased absorption of sunlight, which increases soil nutrient transformation rates, plant growth, and soil biotic activity. Cyanobacteria, green algae, lichens, and mosses are all photosynthetic and thus contribute carbon to desert soils that are otherwise quite low in organic matter. Adequate carbon is required for microbial activity, and low levels slow decomposition and nutrient transformations. The carbon contributed by biocrusts can be substantial, often similar to the soil surface being covered by a vascular plant leaf. Biocrusts can also be the dominant source of nitrogen in deserts. As with carbon, the amount of nitrogen contributed can be large. Secretions by fungi and lichens free biologically unavailable phosphorus, another limiting nutrient, making it available to soil biota and vascular plants. Biocrust organisms can reduce the leaching of nutrients from soils, as their sticky sheaths are covered with clay particles to which essential nutrients cling. In addition, crust organisms secrete powerful metal chelators that maintain nutrients in plant-available forms—​a n important process in high pH desert soils (Belnap et al. 2003). As a result, vascular plants growing in biocrusted areas have higher levels of many essential nutrients than plants growing in areas without biocrusts. Although losses of carbon and nitrogen gases from soils are of great concern globally; the global significance of biocrust effects on these processes are not well understood. The composition of the biocrust community can influence the composition of the vascular plant community. Almost all seed types, regardless of size or shape, can easily penetrate or be buried in bare soil or a thin, early successional biocrust with cracks, unless the soil is covered with a hard mineral crust. However, seeds with large appendages (e.g., exotic cheatgrass, Bromus tectorum) have difficulty penetrating a well-developed, late-successional lichen or moss biocrust that completely covers the soil surface. Therefore a very different vascular flora occurs where early successional cyanobacterially dominated biocrusts are prevalent compared to sites where moss-lichen biocrusts dominate. Biocrusts are highly vulnerable to compressional forces (e.g., trampling, vehicles) and are easily broken by soil surface disturbance. When this occurs, all seed types again obtain easy access to underlying soils. Plant community composition therefore reflects the tension between the successional sequence of the biocrust and disturbance. This dynamic creates a mosaic of soil, nutrient, and plant patch types that move and change through space and time and across multiple scales. There are also many direct ecosystem feedbacks with biocrusts (Schlesinger et al. 1990). For example, soil characteristics and climate influence the distribution of biocrust and vegetation types (Aguiar and Sala 1999), whereas biocrusts and vegetation type affect soil resources via water and nutrient input/uptake, litter deposition, and microclimate alteration. Once disturbed, recovery of biocrusts can be very slow. As biocrusts are only metabolically active when wet, recovery rates depend on rainfall, which is low in deserts. For instance, in the Colorado Desert a study of tank tracks showed only 3–​6% recovery of lichens in the plant interspaces fifty years after disturbance (Belnap and Warren 2002). The concept of a “critical zone,” the area above and below the soil surface that is essential for supporting life, has become central in ecological thinking. In most ecosystems the belowground part of the critical zone is meters deep. In deserts, given that the majority of precipitation events are less than 5 millimeters and soil moisture is required for activity, biocrusts may define the critical zone for nutrients and many 644  Ecosystems

processes in deserts (Pointing and Belnap 2012). Because biocrusts influence almost all materials entering and leaving desert soils, they affect many aspects of ecosystem function and processes in these regions (Belnap et al. 2003). Thus their presence and composition influence trophic levels ranging from soil microbes to plants to the macrofauna that feed on the plants.

Decomposition Decomposition—​the breakdown of organic matter into its constituent parts—​is a basic ecosystem process because much of the fertility of an ecosystem is bound up in these materials. Decomposition in deserts, like many other processes, is controlled more by abiotic factors than biotic ones (Whitford 2002, Ward 2009). Up to 85% of decomposition in deserts results from photodegradation, the process whereby ultraviolet sunlight breaks down compounds in leaves; and by wind, which breaks larger plant parts into smaller ones. Biotic decomposers include micro-organisms (bacteria, fungi, and actinomycetes) and macro-organisms (e.g., mites, collembolans nematodes, protozoa, ants, termites, and beetles). Because soil organisms depend on moisture and organic matter to fuel their activities, and because both are generally low in desert soils, the number and activity rates of soil organisms often is also low. Decomposition was thus long thought to be slow in deserts. However, more recent studies have shown that rates can be surprisingly fast, especially during wet years and periods where heat and soil moisture coincide (Ward 2009, Weatherly et al. 2003). Soil fauna accelerate decomposition in several ways. First, macropores (or soil cavities) created by burrowing organisms allow more water to enter the soil, extending activity times for decomposers. Second, arthropods carry plant matter below the soil surface, preventing it from washing or blowing away and keeping it moist longer. Third, organic matter can be ingested, digested, and excreted by animals, both decomposing it and creating more easily degraded material. Finally, soil animals shred organic matter into smaller pieces, giving fungi and bacteria more surface area for enzymatic attack. Because fungi are able to be active at lower soil moisture, they are likely more important than bacteria in desert decomposition.

The Carbon Cycle The role of carbon in global climate change has garnered much attention recently. This has led to the question of how much carbon is being stored in vegetation and soils relative to how much is being released to the atmosphere in different terrestrial ecosystems. As deserts have sparse vegetation and low growth rates, they are expected to have relatively low net carbon uptake compared to other ecosystems. Most studies from California and other U.S. deserts support this finding. Uptake values from creosote shrublands in the Mojave Desert were found to be 10–​30 gC m-2 yr-1 (Lane et al. 1984, Rundel and Gibson 1996); in Arizona and New Mexico these values were 46–​72 gC m-2 yr-1 (Chew and Chew 1965, Whittaker and Niering 1975, Huenneke and Schlesinger 2006). The carbon content of the other major pools from Mojave sites include plant tissues (25–​65 gC m-2 [Schlesinger and Jones 1984]), with roughly similar amounts in roots. Small car-

bon pools are found in biocrusts (42 gC m-2). This makes a total pool of ~170 gC m-2 in desert plants, compared to 19,300 gC m-2 for dry tropical forest plants (Jaramillo et al. 2003). Whereas soil organic matter pools are small (~670 gC m-2 in the 0–​10 cm layer; Belnap, unpublished data), large pools of carbon are found in carbonates (CaCO3) in hot desert soils (~30,000 g C m-2; Schlesinger 1982). However, current soil carbonate accumulation is slow, ranging from 0.12 to 0.42 gC m-2 yr-1 (Schlesinger 1985, Marion et al. 2008), as it is largely constrained by the atmospheric deposition of calcium. Walvoord et al. (2005) found no evidence for substantial carbon accumulation in the deep unsaturated zone of soils in the Amargosa (Mojave) Desert and documented only a small net upward flux of CO2 from soils to the atmosphere at that site. Other studies from nearby U.S. deserts also show desert soils to be a small net sink in spring and a small net source in fall (Bowling et al. 2011). There have been two studies from the Mojave reporting net ecosystem carbon uptake as large, but major concerns about values reported in these studies have been raised (Schlesinger et al. 2009).

The Nitrogen Cycle Although in most years water is the limiting resource in deserts, nitrogen can be limiting in wet years. Indeed, after only a short time of above-average precipitation, nitrogen and phosphorus (P) limit plant growth. Nitrogen pools in desert soils are small compared to other ecosystems, as soils contain only ~0.02–​0.12% N, compared to 1.2% in the boreal forest (Lavoie et al. 2011). In the northern Mojave, shrub nitrogen was estimated at 33 kgN ha-1 and total nitrogen (plants and soils) at three sites ranged from ~965–​1,533 kgN ha-1. Because of the relatively low vegetative biomass in deserts, 70–​98% of the system’s nitrogen can be found in the soils (Rundel and Gibson 1996). Wetter ecosystems receive most of their nitrogen via precipitation, but deserts have low precipitation and generally receive only ~1–​2 kgN ha-1 yr-1 in nonurban areas. Nitrogen fixation by bacteria inside nodules on roots of legumes (e.g., mesquite) can be a good source at a very local scale, but the abundance of these plants is limited in California deserts, as are any contributions by free-living bacteria. Fixation by biocrusts can be a main source of nitrogen. They can contribute 4 kgN ha-1 yr-1 or more where they are exceptionally well developed; however, this value varies widely depending on lichen and cyanobacterial biomass. Dust inputs are important as well. Nitrogen losses in these deserts have the potential to be very high. For example, McCauley and Sparks (2009) showed abiotic gaseous losses of ~3 kgN ha-1 yr-1 in the Mojave Desert. Soil erosion also can result in high levels of nitrogen loss, depending on site disturbance. For instance, losses were estimated at 2 kgN ha-1 yr-1 for one site in the Mojave ungrazed by livestock, whereas losses at nearby grazed sites were up to 24 kgN ha-1 yr-1 (Rundel and Gibson 1996). Because availability is low in desert soils, plants attempt to conserve what nitrogen they have in their tissue. Accordingly, many plants can move up to 40% of the nitrogen found in their leaves to their stems or roots before they drop their leaves. Another loss pathway can be through the death of biocrust organisms. Although data are lacking for the California deserts, data from the nearby Colorado Plateau desert showed that loss of biocrust moss substantially altered nitrogen cycles (Reed et al. 2012). In soils under dead mosses nitrate (NO3-) was higher and ammonium (NH4+) was lower than in soils beneath live

mosses. Microbial nitrogen pools were also lower in soils with dead mosses. As nitrate is more easily lost from soils via leaching and gaseous losses, this switch from ammonium to nitrate could result in lowered total soil nitrogen. Also, type of nitrogen (NH4+ vs. NO3 -) may be more important than quantity in regulating ecosystem function (Austin et al. 2004). Organisms can readily uptake and use ammonium, but if they are confronted with nitrate, they have to use energy to convert it to ammonium before uptake.

Nutrient Transfers and Distribution It is generally believed that in all ecosystems, plants obtain the bulk of their nutrients from the soil via either plant roots or mycorrhizal fungi attached to their roots. Recent studies in deserts, however, have suggested that an additional mechanism may be operative in these regions. When labeled N and C were added to the top few millimeters of root-free interspace soil, both were detected in plants over 1 meter away within twenty-four hours (Green et al. 2008). The C label, but not the N label, from the plants was detected in biocrusts. Because the transfer occurred so quickly, these interspace compounds clearly bypassed bulk soils that would have resulted in a very slow transfer. As there were no roots in the soil, these compounds were most likely moved by a fungal network dominated by dark septate fungi, a type of endophytic fungi common in desert soils. Many nutrients are concentrated at the soil surface in deserts as the result of biocrust activity (e.g., C and N fixation, greater bioavailability of other nutrients such as P), dust deposition, and reduced loss via leaching because rainfall is low. This situation could favor the development of a fungal “highway” located at the soil surface that directly links biocrusts to nearby vascular plants (Collins et al. 2008, 2014). In deserts, nutrients and organic matter are often concentrated under shrubs. These zones have been termed “islands of fertility” (e.g., Schlesinger et al. 1990, Walker et al. 2001). Because carbon, some nutrients, and soil moisture can be higher where these islands occur, the abundance and activity of soil fauna, decomposition, nutrient transformation rates, and nutrient availability are higher there as well. This concentration of nutrients and organic matter influences plant germination and growth. A variety of processes can lead to these patterns. Plants extract nutrients from the interspace soils with their roots and use them to build leaf and stem tissue. This tissue eventually dies and is dropped, enriching the soil beneath the plant. Plant canopies also provide shade so that soils retain moisture longer, which allows greater microbial activity and nutrient transformations, increasing nutrient availability. In addition, soil and organic matter moved by wind and water from plant interspaces are intercepted by the plants (Soulard et al. 2013). This dynamic between “source” patches (where materials come from—​in this case, the plant interspace) and “sinks” (where materials are deposited—​in this case, under the plant canopy) can be a strong driver of ecosystem processes in deserts (Ludwig and Tongway 1997). Although the evidence is quite compelling that islands of fertility form in places where livestock grazing has heavily impacted plant interspaces (Schlesinger et al. 1990), fertile islands are less frequently observed where interspace soils are still covered by well-developed biocrusts and lack visible soil loss (Housman et al. 2007, Allington and Valone 2011). Therefore an alternative explanation for heterogeneous nutrient Deserts  645

distributions in deserts is that heavy land use has resulted in “oceans of depletion” rather than “islands of fertility.”

Primary Producers: The Vascular Plants Adaptations to Extreme Conditions The variable and extreme conditions found in deserts present a great challenge for plants, with the lack of water generally the most challenging condition plants face. Desert plants exhibit several strategies for surviving low levels of soil water: escape (e.g., annuals that escape as seeds), evasion (e.g., cacti that use tissue succulence or plants able to use perennial waters), endurance using deciduousness (e.g., bursage), or resistance (e.g., creosote). Some plants can combine these responses. For example, a drought resister can become drought-deciduous and drop its leaves when low soil water persists. Organisms can only work within the constraints of their evolutionary history, meaning all strategies are not possible for a given species. If, for instance, there is no genetic material orchestrating a leaf-drop response to drought, a plant will not be able to adopt this strategy. Many plants growing in deserts today did not originally evolve in a desert setting, so features seen in a given plant may be not adaptations to desert living but carryovers from some other set of conditions. Desert plants have many structural and physiological adaptations to help them tolerate extreme conditions (Rundel and Gibson 1996). Their leaves are often fewer, small, and narrow (less than 10 millimeters wide). These leaves have a high surface-to-volume ratio that helps them stay below lethal temperatures without having to use evaporative cooling, thus conserving water. In addition, maximal photosynthetic rates often acclimatize as leaf temperatures increase (Rundel and Gibson 1996). Some plants can maintain both wide and narrow leaves, then drop the wide leaves when water is scarce. Water loss and leaf temperature can be reduced with stems and leaves that are thick and/or waxy; covered with light-colored hairs, spines, or salt crystals to reflect light; and either oriented parallel to incoming light (creosote, prickly pear) or able to track solar rays (e.g., sunflower [Malvastrum rotundifolium]). Pores (stomata) that allow gas exchange can occur on the underside of the leaf, where they are shaded; this reduces water loss when pores are opened. Photosynthetic pathways can also help plants adapt to extremely dry conditions (Smith et al. 1997). All three pathways (C 3, C 4, and Crassulacean acid metabolism [CAM]) occur in desert plants. The C3 pathway is the most common. Plants using this pathway open their stomata during the day to obtain CO2 while ambient light allows for photosynthesis. However, these open stomata also release water from the plant’s leaves, reducing water efficiency. For this reason, C3 photosynthesis is especially abundant in winter-active plants, when water is less limiting. The C4 pathway is thought to have evolved from the C3 pathway, either because of, or allowing, a migration of grasses from shady forests to more open environments. The enzymes used in the C4 pathway are more efficient at high light and temperatures found in open habitats and are also thought to have high water-use efficiency. Thus C4 plants are most often summer annuals found in in the hottest desert locations or in very salty soils. In CAM photosynthesis, the leaf stomates remain shut during the day to reduce evapotranspiration and open at night for CO2 uptake, making these plants much more water-use efficient than C3 or C4 646  Ecosystems

plants. The CO2 is stored until daytime, when it is used for photosynthethis. CAM is common in plants with succulent leaves and/or stems (e.g., cactus) and is highly correlated with aridity. Roots of desert plants were long thought to go deeper than those of plants in other biomes to access deep water. While desert plants do not have the deepest roots of all biomes, their roots on average are deeper than most (Schwinning and Hooten 2009). However, desert plant root architecture is tremendously variable among both species and individuals within a species. Some plants have mostly surface, compact roots (e.g., grasses, prickly pear), some rely mostly on a deep taproot (e.g., mesquite), and others have roots that both spread widely and go deep (e.g., creosote). Root type is often a trade-off between the ability to transport smaller volumes of water in wetter soils (e.g., drought-deciduous shrubs) and the ability to operate in very dry soils (e.g., drought-tolerant evergreen shrubs). Multiple root types can occur in one plant (e.g., broom snakeweed [Gutierrezia sarothrae]). Root distributions often depend on geomorphic setting. For instance, plants along washes tend to lack lateral roots (e.g., Mormon tea [Ephedra spp.]), whereas many common shrubs in alluvial fan settings (e.g., creosote, bursage) have lateral roots as well as deep taproots (Rundel and Gibson 1996). Surface age, calcic horizons, soil texture, and desert pavements can all influence rooting patterns (Schwinning and Hooten 2009). On Holocene surfaces in one study, for example, creosote roots commonly penetrated to over 100 centimeters, whereas on Pleistocene surfaces, they were mostly restricted to the top 50 centimeters (Stevenson et al. 2009). Most roots are found at least 10 centimeters below the soil surface, as conditions are too hot and dry for root survival above that level. It was long believed that in the search for scarce water, desert plants should have far more roots relative to aboveground biomass than plants in wetter ecosystems. However, it turns out that the root-to-shoot ratio is very similar for perennial plants in both the Mojave Desert and temperate forests (0.5–​1.0) (Rundel and Gibson 1996). Desert roots also have a high capacity to extract water from dry soils through several strategies. First, some species like creosote can maintain turgor at much lower soil water levels than more mesic species and can therefore maintain a water potential gradient for water uptake from drying soils. Roots can appear within a day after rainfall events, follow soil moisture downward, and be shed after soils dry (Schwinning and Hooten 2009). Plants can also use hydraulic redistribution—​ a process whereby roots in moist soil layers absorb water and passively translocate it to roots in dry layers. This water is then exuded at night and reabsorbed the next morning, along with nutrients dissolved in the water. In this way plants can keep their entire root system active despite patchy water distribution in the soil (Caldwell et al. 1998). As released water can be taken by other plants, some plant roots such as creosote exude compounds to keep other plants’ roots away, a process termed allelopathy (Mahall and Callaway 1992). Another extreme condition desert plants often confront is salty soils. Rapid evaporation rates leave salts at the soil surface, and low rainfall prevents them from being leached downward into the soil. Salt levels can be toxic and difficult for plants to exclude. Plants have three options: (1) exclude salt-containing water at the root; (2) store salt in internal membrane-bound compartments, or vacuoles, making their leaves plump and liquid-filled, as seen in greasewood (Sarcobatus vermiculatus); or (3) exude salt onto their leaves (e.g., tamarisk) or into leaf hairs (e.g., saltbush). However, many

plants are unable to utilize any of these techniques, which is why very salty soils support few, if any, plants. Plants that can tolerate salty soils are called halophytes. As plants are a fundamental unit of ecosystems, both their three-dimensional shape and the pattern of their spacing are important aspects of ecosystem structure. These characteristics primarily define the habitat that animals “see” and food that is available to them. The architecture of a plant can be so important to some species of birds and spiders that they will inhabit a stick imitation of a shrub, ignoring the fact it is dead. Lizards that are “sit-and-wait” predators require shrubs with branches low enough to hide them, while sufficiently high to provide a clear line of sight and unencumbered access to prey. Many birds require perches of a certain height and select plant species accordingly. The shape of a plant is a cost-benefit game. For instance, the shape and size of plants and individual branches can enhance water harvesting. Large, flat, spread-out canopies and horizontal branches intercept the most rain. Where the plant’s shape delivers intercepted water is also important. The optimum is water delivery to where the plant’s roots are concentrated, which requires inwardly sloping rather than flat branches. Shaded soils retain moisture longer, and concentrated canopies are best for creating shade. Light is not limiting in deserts, as there are wide spaces between plants, so unlike trees in a forest, desert plants generally expend energy to grow out rather than up. With so much light, desert plants can also be bushy without shading their own leaves, unlike their forest counterparts. There are other design constraints on desert plants. For example, deserts often have extreme winds. In response, plants can be very stiff, such as blackbrush and cottonwood (Populus fremontii). This strategy can fail if winds are so strong the branches break. Alternatively, plants can be very limber and sway with the wind, like rabbitbrush (Chrysothamnus spp.), or can twist as they grow, like juniper and sagebrush. Some tall species, such as cottonwood, are able to regenerate vegetatively, and wind damage may create the possibility of new clones when fallen branches take root. Heat and water stress are also important in determining what a plant looks like: stems can act as water storage (cacti) or be arranged to maximize cooling (e.g., saltbush with vertical, reflective leaves).

Plant–Plant Interactions and Vegetation Dynamics There are several factors that shape the spatial patterning of plants at a given site. Although light is important in determining these patterns in mesic regions, light is seldom limiting in deserts. Abiotic factors can heavily influence where plants grow at a very local scale. Biotic interactions, including rooting patterns and nurse plant interactions, can also structure plant patterns. Plants that root at the same depth may be in direct competition for water (Cody 1986)—​a finding that is supported by the fact that the highest levels of plant diversity are found on coarse alluvial fans, where water can infiltrate to deeper levels. Conversely, surface-rooting plants (e.g., annuals) can intercept water before it infiltrates to depth, giving them an advantage over plants rooted below them. Allelopathic compounds secreted by creosote keep the roots of bursage and of other creosote plants at a distance, giving the creosote an evenly spaced pattern. Bursage roots, on the other hand, avoid contact with roots only of other

bursage plants and are often found clumped with other species. Some desert shrubs have roots that go straight down from the plant base to a deep soil layer, turn ninety degrees, grow laterally for some distance, and then go straight back up to the surface, discouraging plants from growing in the space between the root and plant base (Gile et al. 1995, 1997). Therefore, root architecture can shape plant distributions for some species. Nurse plants also influence plant distribution and composition patterns at a local scale. Nurse plants are larger established plants that shelter smaller plants, “nursing” them by providing shade and thus more soil moisture; cover to hide from predators; organic matter; and sometimes soil nutrients. In the Colorado Desert, ironwood may act as the nurse plant for a variety of plants. In the Mojave, Joshua trees often establish under a variety of shrubs and the desert agave (Agave deserti) under big galleta grass (Pleuraphis rigida). In many cases, the plant being nursed conveys a benefit to another species, and no harm is caused to either plant (Callaway 1995). However, in some cases the plant that is being nursed harms or eventually kills its benefactor (e.g., saguaros under paloverde; McAuliffe 1984). In an experimental study the nurse plants white bursage, creosote, and blackbrush potentially facilitated growth of other species, but their net effect was sometimes negative due to shading or root competition (Walker et al. 2001). It has been suggested that imitation of the nurse-plant phenomenon might be used to restore damaged ecosystems in harsh environments (Padilla and Pugnaire 2006). Precipitation also has a profound influence on the composition and productivity of desert plant communities. For example, Beatley (1980) concluded that most perennial plants recorded in eastern Mojave plots in 1963 were still present in 1975, but the number of plants had increased by 20% to 30% due to a wet period in the late 1960s. Hereford et al. (2006) found that drought had resulted in a large increase in dead biomass on living plants, and Miriti et al. (2007) and McAuliffe and Hamerlynck (2010) concluded that drought had culled substantial numbers of creosote and bursage in the southeastern and eastern Mojave Deserts, with up to a 100% loss in both white bursage and triangle leaf bursage. Because prolonged droughts and episodic wet periods characterize deserts, study results can depend on both what years are reported and the geographic location of the study area. Cody (2000) reported little change in a creosotewhite bursage plot near the border of the central and eastern Mojave Deserts (see Figure 30.2) measured at an interval of twenty-five years; however, the two measures of this single plot spanned a wet period. Similarly, plots in the Colton Hills, measured periodically from 1966 to 2001 (Smith and Smith 2002), showed little change with time. Although these studies would appear to refute claims of pronounced long-term change in Mojave Desert vegetation, they were mostly conducted in wet periods and show little drought effect. Two extensive and very long-term studies examining vegetative change have been performed in the California deserts. In the Great Basin Desert of northeastern California, sixtyeight transects measured between 1957 and 1998 showed a large increase in juniper and pinyon and a concomitant decrease in antelope bitterbrush (Purshia tridentata) and big sage (Schaefer et al. 2003). The transition from sage-bitterbrush shrublands to juniper-pinyon woodlands substantially reduced the shrub food base for ungulates (Schaefer et  al. Deserts  647

2003). Another long-term study was begun at the Nevada National Security Site (NNSS, formerly Nevada Test Site) in the eastern Mojave in 1963 (Beatley 1980) and remeasured in 1975–​2 011 (Webb et al. 2003, R. H. Webb, unpublished data). Changes observed in these plots illustrate the tension between vegetation increases (e.g., germination, establishment, and productivity) during sustained wet periods and decreases (e.g., branch pruning and mortality) during droughts. Low-elevation sites showed the greatest changes at NNSS, with drought resulting in a high mortality of chenopods (e.g., spiny hopsage [Grayia spinosa], saltbushes, and winterfat [Krascheninnikovia lanata]). Wet years, such as those in the early 1980s, resulted in dominance by Indian ricegrass (Achnatherum hymenoides) (Figure 30.7). Subsequent drought reduced these grasses, leaving Anderson thornbush (Lycium andersonii) dominant by 2011. Extreme cold affected plants as well over the monitoring period, decreasing cover of spiny menodora (Menodora spinescens), brittlebush (Encelia farinosa), barrel cactus (Ferocactus eastwoodii), and prickly pear cactus (Opuntia; Webb et al. 2003; Webb, DeFalco et al. 2009). Overall, creosote-dominated areas increased in cover, despite fluctuations in associated subshrubs (Figure 30.8). Exceptions occurred in some lower-elevation creosote assemblages with substantial chenopod die-off. At higher elevations, blackbrush and big sage assemblages generally lost cover but had low mortality; however, pinyon and juniper trees increased in cover. Thus long-term vegetation response strongly depends on elevational gradients and species composition. Changes observed over a half century at NNSS underscore the potential for climatically driven processes to significantly alter the structure, function, and species composition of perennial vegetation in the Mojave Desert. Shrub assemblages in this part of the Mojave Desert respond to climatic events as aggregations of individual species, not as a collective community, and directional changes can occur without immediate rebound. For some species climatically induced changes may rival the magnitude of some changes caused by landuse practices such as dispersed livestock grazing and vehicular recreation. Finally, succession in plant communities—​ a process whereby early colonizing plant are replaced over time by plants considered to be late colonizers—​is a central tenet in many ecosystem studies. However, succession appears less operative in deserts. Long-term transects and repeat photography indicates that a wide variety of plants can become established at a site after disturbance and, once there, can persist for extraordinary lengths of time. It may be that by garnering available nutrients and water, the initial colonizers simply keep other plants from establishing. Nevertheless, certain species that are most often found in disturbed sites, such as cheesebush (Hymenoclea), snakeweed, and tumbleweed (Salsola), and these often do eventually give way to longer-lived shrubs.

Key Consumers Food Webs and Trophic Pyramids Nutrients and energy are captured by plants and subsequently eaten by animals. These complex resource networks have long been described as food webs—​a concept introduced nearly one hundred years ago to simplify complicated biotic interactions (Elton 1927). Food webs can be immensely complicated 648  Ecosystems

and multidirectional, as illustrated by the numerous connections among just a few of the predaceous arthropods living in the Coachella Valley of the Colorado Desert (Polis 1991) (Figure 30.9). These organisms then fit within an even bigger food web, as they consume and are consumed by others. Another conceptual model used to describe the movement of resources through ecosystems, the trophic pyramid, not only describes the directional flow of nutrients through organisms but also groups organisms with similar ecosystem functions into trophic levels (e.g., producers, primary consumers, omnivores, predatory consumers). Higher trophic levels feed on those below them; the primary producers (plants) support all consumers in an ecosystem, either directly or indirectly. First-order consumers, animals that eat plants, include herbivorous insects and their larvae, some reptiles and birds, many small mammals, and ungulates. Granivores eat the seeds of plants. Many ants, birds, and small mammal species are particularly important granivores because they consume a large proportion of total seed production. Preferences for certain seed characteristics (e.g., size, shape, nutrient content) of granivores influences the composition of the seed bank, often determining the composition of the plant community and the recovery of desert landscapes after disturbance (Esque 2004). Pollinators often consume plant nectar or pollen. Whereas most pollinators are insects, some vertebrates perform this role (e.g., bats and hummingbirds). Folivores, animals that eat plant leaves and stems, include invertebrates, reptiles, amphibians, birds, and small and large mammals. Frugivores, animals that consume plant fruits, are also very common among the various animal groups. Second- and third-order consumers feed on organisms in trophic levels below them. Complex food web interactions often make it difficult to classify species as strictly second- or third-order consumers. Position in a food web also can vary among individuals within a species, depending on resource availability. Top predators can often feed at any consumer level of the pyramid. Many animals are omnivorous consumers, eating both plant and animal materials. Omnivores are dominated by birds and mammals that adapt to food shortages by accepting a variety of foods. Decomposers complete the loop, breaking down dead plants and animal materials and enabling the nutrients and carbon contained within these materials to reenter the ecosystem. Each time organisms are consumed, up to 90% of their resources are lost to the environment through inefficiencies such as heat transfer and physiological by-products, reducing resources available to support the abundance of individuals in each successively higher trophic level (Rundel and Gibson 1996). Therefore primary plant producers have the greatest mass, followed by primary consumers, and so on, such that top-order predators are quite rare, resulting in a pyramid of biomass, species diversity, and abundance. Limited numbers of species in the higher tropic levels especially occur in desert ecosystems, where prey are usually less abundant and more sporadically available than in mesic environments.

Animal Adaptations to Desert Life Desert animals have many behavioral and physiological adaptations to extreme desert conditions, especially high temperatures and lack of water. Maintaining sublethal body temperatures is achieved through a combination of body size, activity

FIGURE 30.7 Photographs of Plot 2, a mixed-shrub assemblage visually dominated by creosote bush (Larrea tridentata), on the Nevada National Security Site (NNSS), formerly Nevada Test Site. Source: R. H. Webb et al. unpublished data.

FIGURE 30.8 Photographs of Plot 50, Nevada National Security Site, formerly Nevada Test Site, which has been undisturbed for its fortyeight years of existence as a monitoring plot. Source: R. H. Webb et al. unpublished data.

A 1964: A mixture of shrubs is present, notably creosote bush, spiny

A 1964: At this time, the plot was dominated by chenopods,

hopsage (Grayia spinosa), and winterfat (Krasheninikovia lanata). White bursage (Ambrosia dumosa) is a minor component of this assemblage.

primarily spiny hopsage (Grayia spinosa) and winterfat (Krashinennikovia lanata).

B 2000: Creosote bush now dominates this plot in terms of cover;

white bursage is the co-dominant. Spiny hopsage and winterfat cover is greatly reduced because of drought effects from 1989 through 1991. C 2011: Although reduced in cover due to the early twenty-first-

century drought, creosote bush and white bursage are the co-dominants with Mormon tea (Ephedra nevadensis), which has increased in cover.

B 2001: Drought decimated the chenopods, resulting in an

assemblage dominated by Indian rice grass (Achnatherum hymenoides). C 2011: Anderson’s thornbush (Lycium andersonii), which tends to

co-occur with spiny hopsage, now dominates this plot.

Arthropodivores further into the food web Tastiotenia hyperparasitoid P. mesaensis scorpions Chalybian parasitoid wasp Steatoda Mimetus spiders spider Latrodectus spiders Pteromalid 10 other spiders parasitoid invading egg sac wasp Diguetia adults Salticid Phyllobaneus spiders Clerid beetle Diguetia eggs & spiderlings

Predacious insects (e.g., Asilids, Antlions, Mantids) (>48 families)

Parasitoid insects (e.g., Mutillid & Tiphiid wasps; Photopsis) (Bombyliid, >75 spp.)

Ants (16 spp.) Termites

Tenebrionid beetles (14 spp.)

Detritus

Herbivorous insects (>74 families)

Plants (174 spp.)

FIGURE 30.9 Trophic interactions above the soil surface involving a few of the predacious arthropods living within the Coachella Valley, Colorado Desert. This subweb is focused around the spiders Diguetia mojavea and Latrodectus hesperus. An arrow returning to a taxon indicates cannibalism. Source: Polis 1991.

patterns, microhabitat selection, coloration, metabolic rate regulation, hyperthermia (storing body heat and releasing it hours later), and countercurrent heat exchange systems (using blood). Water balance and osmoregulation are also closely regulated. As different animal groups utilize different mechanisms to regulate temperature, water, and solute concentrations, these various approaches are discussed by animal group below. There are, however, some aspects in common among all groups. Optimal body temperatures for desert animals range from 35°C to 39°C, several degrees higher than for nondesert animals (Schmidt-Nielson 1990, Rundel and Gibson 1996). Despite this, most desert animals still cannot be active on hot summer days. To avoid the heat, many animals (including invertebrates, reptiles, and small mammals) employ burrows. Soil temperatures just 30 centimeters below the surface can be a comfortable 33°C in summer, varying only 1–​2°C over a twenty-four-hour period. Burrow air is also highly humid, reducing the amount of water vapor lost dur650  Ecosystems

ing respiration. Coloration may play a role in regulating heat loss and gain. However, as animals range in color from very light to black (e.g., common raven [(Corvus coras], tenebrionid beetles), it is not clear whether coloration serves to avoid predation, to regulate temperature, or a combination of both. Body size is important in thermoregulation, as animals with larger bodies have a lower surface-to-volume ratio and therefore gain heat more slowly. Water loss through evaporation and respiration is a common problem for all animals but is especially acute for desert dwellers for two reasons. First, hot dry times when water losses through evaporation and respiration are the highest (e.g., summer) are also the times when water is least available as free surface water or in vegetation. Second, the potential for rapid water loss during these hot dry times, resulting in high fluid electrolyte concentrations, is extremely high in deserts and can quickly cause great stress (Schmidt-Nielson 1990). The water flux of an animal is the rate of water gain

and loss per day. Water is lost via excretion (urine and feces) and evaporation from respiration and body surfaces. Water gain occurs through drinking, eating food that contains water, or body surfaces. Daily loss of water is linearly and positively correlated with the mass of an animal (Schmidt-Nielson 1990, Rundel and Gibson 1996). For example, a 10 gram bird loses twice as much water each day than a 5 gram bird. This rule of thumb is true for reptiles, birds, mammals, and arthropods, though the amount of water lost differs among the animal groups, with reptiles and arthropods by far the most water efficient, followed by mammals, then birds. A 10 gram bird loses water three times faster than a 10 gram rodent; the rodent, in turn, loses six times the water of a 10 gram reptile. In general, desert vertebrates are more water efficient than nondesert vertebrates. However, organisms in the California deserts do not have specialized internal water storage and must instead conserve it. The trade-off with body mass places animals in a conundrum: larger body size prevents excess heat gain but accelerates water loss.

INVERTEBR ATE S

Invertebrates are the most abundant, diverse, and highest in total biomass of desert faunal groups, making them essential to many ecosystem processes (Crawford 1981). Most invertebrates are first-order consumers and include herbivores, pollinators (especially bees, wasps, moths, and butterflies), granivores, and/or root herbivores. Some invertebrates prey on other invertebrates. Other invertebrates decompose plant (detritivores) and animal materials (coprovores eat fecal material, necrovores eat carcasses). Invertebrates also are eaten by many second- and third-order consumers as well as by top predators. Outside of their role in food pyramids, invertebrates have many important influences on plant community composition through herbivory, seed predation, and seed dispersal. Invertebrates also influence soil structure and fertility as the result of burrowing and vertebrate community structure by providing food to these animals. Desert invertebrate communities generally involve larger body sizes, higher abundances of predatory insects, a larger predator-to-prey ratio, and increased sociality in groups with predicable food resources (e.g., ants, termites) when compared to nondesert groups (Crawford 1986). Most arthropods avoid temperature extremes to some degree (Whitford 2002), as their optimal temperature range is between 30°C and 39°C, with 46°C being lethal (Rundel and Gibson 1996). Behavioral adaptions include seeking lower temperatures in the day by burrowing into sand, leaf litter, or finding deep shade (e.g., centipedes, millipedes); being active at night, dawn, and/ or dusk (e.g., tenebrionid beetles, arachnids); or moving up and down in plant canopies (e.g., grasshoppers). Some species combine these strategies. For instance, termites use deep burrows and limit their activity to times of favorable temperatures. Invertebrates also have physiological adaptations to desert conditions. Their relatively small body size means they have a large surface-to-volume ratio that can result in faster and greater heat gain. On the other hand, that same small body size means that heat generated by activity or from the air can be quickly lost. To control water loss, many invertebrates have a cuticle, a hard outside coat. Water vapor is allowed to leave through small holes in the cuticle (spiracles), allowing some evaporative cooling. These cuticles are often covered by thick waxes

with high melting points (Rundel and Gibson 1996). In one study, desert arthropods were found to be more water efficient and have lower evaporation rates than similar species from nondesert regions (Crawford 1981). Many soil invertebrates can also use anhydrobiosis, the ability to desiccate and suspend all activity until wetted, to avoid extreme conditions (Crawford 1981). Lastly, some arthropods, such as the desert cockroach (Arenivaga investigata), can replenish body water by absorbing it from unsaturated air in their burrows (Rundel and Gibson 1996). Some of the more important, visible, and interesting invertebrates in deserts include ants, termites (discussed below), bees, moths, butterflies, grasshoppers, and scorpions. Ants are a vital component of most desert ecosystems, as they engineer many aspects of the environment. Ants are both first- and second-order consumers and provide food for many other animals. The diversity of ant species follows a climatic gradient in the California deserts (Figure 30.10). Harvester ants are especially important, as they move large quantities of nutrient-containing subsurface soil, seed coats, and insects (80–​2 80 kg soil ha-1 yr-1) to the surface when they create their mounds, increasing soil fertility (Parmenter et al. 1984, Taber 1998). Several ant species clear vegetation around their nests, affecting plant distribution patterns, and make mounds up to 5 meters wide and 1 meter high, affecting local hydrologic patterns. Colony densities range from twenty to fifty colonies per hectare. As a group, harvester ants eat more than one hundred species of seeds, but different species often show narrow seed preferences. Unlike rodents, ants usually collect seeds from the soil surface. Like rodents, ants play a role in seed dispersal because they drop many of the seeds they collect (MacMahon et al. 2000, DeFalco et al. 2010). Ant nests also provide refuge for other animal species, including beetles, orthopterans, termites, homopterans, collembollans, thysanurans, diplurans, millipedes, spiders, and mites. In a southeastern Arizona study, harvester ant mounds contained thirtyfold higher densities of microarthropods and fivefold more protozoans than plots without mounds (Wagner et al. 1997). Native bees are one of the most diverse groups in California deserts. As pollinators, they are first-order consumers. Solitary bees, which do not live in colonies, do most of the pollination in desert ecosystems (Crawford 1986). This may result from the low quantity and unpredictability of desert flowers, making this food source too unreliable to support large bee colonies. Desert bee sizes range from tiny to some of the largest bees in California. This range in size may help partition resources so that more species can be supported in a given setting. In addition, many bees are oligolectic, meaning they specialize on one plant species. Moths and butterflies are also important consumers of plant material. Adults deposit eggs on plant leaves; when the eggs hatch, the resulting caterpillars eat the plant leaves. Often, the adults pollinate the plant at the same time they deposit their eggs. Similar to bees, the larvae are often oligolectic. Thus, through herbivory, moths and butterflies can heavily influence plant community composition. Moths and butterflies are also important prey items for birds and reptiles. Grasshoppers consume vegetation and, as they are often quite numerous, can have a large impact on desert vegetation. While some species eat many different plant species (polyphagy), other grasshopper species eat only a specific plant (monophagy). For example, the creosotebush grasshopper (Bootettix argentatus) is found only on creosote. This grassDeserts  651

Species diversity

2.0

selves to the males and are transferred to female bees during subsequent sexual contact. The female bee then inadvertently transports the beetle larvae to the bee’s nest, where pollen, nectar, and the bee’s eggs provide nutrition for the larvae. Both the beetle and the bee utilize and are dependent at various times on the Borrego milkvetch, which provides the link in this unusual food web.

Ants Rodents

1.5

1.0

Ants

0.5

VERTEBR ATE S Rodents

0 0

50

100

150

200

250

300

Mean annual precipitation (mm)

FIGURE 30.10 The effect of mean annual precipitation on desert ant and rodent species diversity in the deserts of the southwestern United States. Source: Davidson 1977. A nts: Y=0.0054X+0.1774, r=0.93, p50 cm yr-1 precipitation); (C) coastal Mediterranean grassland; and (D) hot and cold desert range. The oak canopy effect (shaded area) indicates the difference exerted by the presence of oaks in (B) versus (A), where oaks suppress biomass during the wet season but extend forage productivity longer into the dry season. 1 kg ha-1 ≈ 1 lb. ac-1.

“Grasslands”). With little change in its herbaceous character, valley grassland and coastal prairie extend into the understory of oak savanna , oak woodland, chaparral, and other scrub vegetation types (Allen-Diaz et al. 2007). The northern coastal shrub dominant, the short-lived shrub coyote brush (Baccharis pilularis), invades and modifies coastal prairie but also is common quite far inland (Hobbs and Mooney 1986, Williams et al. 1987). Outside of California’s Mediterranean climate zone, grassland and woodland ecosystems are less widespread, replaced mainly by forests and shrublands. Grazing was formerly more important on transitory openings in forests but has greatly declined in recent decades. Use of middle- to high-elevation meadow systems is still common. Montane forage characteristics are quite different from that of the lowlands, with many native perennial grasses and sedges and a short summer grazing season. While the impenetrable and unpalatable chaparral of the Mediterranean zone is little grazed, cold and hot desert shrublands in northeastern and southern California furnish important perennial grass forage and ephemeral annual forage, respectively.

peratures decline, followed by rapid spring growth as soil temperatures increase while there is adequate soil moisture (Chiariello 1989). During spring, forage production usually exceeds the ability of grazing animals to consume it. Peak standing crop of the herbaceous vegetation generally occurs between April 1 and May 15, followed by the death of the annual plants. Standing dead biomass slowly decomposes as summer drought limits microbial activity until the ensuing autumn rains stimulate decomposition. Sufficient fall rainfall also brings about new annual plant germination (Jackson et al. 1988).

INL AND MEDITERR ANE AN ANNUAL GR ASSL AND WITH OAK CANOPY

Adding tree canopy cover to the annual grassland alters the local environment and affects both forage production and utilization. The type and amount of tree cover (evergreen or deciduous), geographic location (high or low rainfall), and local soil factors can be important variables (Frost and McDougald 1989, Frost et al. 1997, Rolo and Moreno 2012). Dahlgren et al. (1997) described soils beneath oak canopy as “islands of fertility” because of their greater carbon, nitrogen, and phosphorus stocks compared to adjacent open grassland sites. The patchy distribution of oak trees might be enhanced by the ability of oaks to garner water and nutrients from the open spaces between trees and then concentrate them by way of leaf fall to the area beneath the canopy (Schlesinger and Pilmanis 1998, Cross and Schlesinger 1999, Huenneke et al. 2002). An untested hypothesis is that herbivores provide a check on this effect by harvesting plant nutrients from beneath the canopy and redistributing them in a more homogeneous way across the landscape. Shade from trees usually inhibits herbaceous production in areas of California receiving more than 50 centimeters of annual precipitation (see Figure 37.4b; McClaran and Bartolome 1989b). The opposite generally holds for drier portions of the state, where shade reduces drought stress, moderates temperatures, and can increase forage growth. Frost et al. (1997) reported that an increase in tissue and litter nutritional quality under oaks in drier regions more than compensated for any possible reductions in forage amount. The differences in species found in shade as opposed to full sun explained differences of forage nutrient quality more than changes in the nutritional content of individual plants.

INL AND MEDITERR ANE AN ANNUAL GR ASSL AND COASTAL MEDITERR ANE AN GR ASSL AND

Grasslands of the inland Mediterranean climate zone (i.e., valley grassland) comprise largely exotic annual grass and forb species from the Mediterranean basin and surrounding areas (Bartolome, Barry et al. 2007). Factors at many spatial and temporal scales interact to control herbaceous productivity of California’s annual grasslands (Bartolome 1989). Aboveground biomass at the time of late spring seed set varies interannually as a function of the timing and amount of rainfall, temperature (Talbot et al. 1939, Bentley and Talbot 1948, Heady 1958, George et al. 1988), and edaphic and topographic characteristics (Jackson et al. 1988, Callaway et al. 1991). A typical yearly production curve for annual grassland starts with the onset of autumn germination following fall rains of more than 2.5 centimeters in a one-week period (see Figure 37.4a). Winter growth slowly progresses as tem844  Managed Systems

Coastal prairie grasslands occur in climate zones with significant coastal fog. Native perennial bunchgrasses are more prevalent there than in inland grasslands. More summer moisture allows perennial grasses to continue growing in the summer when exotic annuals have gone to seed (Corbin et al. 2007; see Chapter 23, “Grasslands”). The growth/production curve for coastal prairie (see Figure 37.4c) is quite different from more seasonally arid grasslands. The shorter period of water deficiency in summer, more moderate winter temperatures, and different growth patterns of perennial and annual grasses influence the growth curve. Typically, fall and peak biomass production are greater than for annual-dominated valley grasslands (George, Bartolome et al. 2001; Bartolome, Barry et al. 2007). The decline in forage amount and quality

after the spring peak is slower as perennials go dormant later. Coastal grasslands are among the most productive in North America both in net primary productivity and livestock production (Huntsinger et al. 2007).

HOT AND COLD DE SERT R ANGE

Outside of the Mediterranean zone, lower total rainfall and cold winters limit forage (see Figure 37.4d). On intermountain cold desert steppe where perennial grasses like blue-bunch wheatgrass (Elymus spicatus) are common, forage quality is high even during the dormant season in winter, and beef cattle are often raised without supplementation but at low stocking levels. Invasive annual grasses can increase production in some years but reduce forage quality and increase fire hazard. On warm desert range, annuals like red-stem filaree (Erodium cicutarium) form the basis for ephemeral grazing permits (Holechek et al. 2011). Forage growth in mountain meadows is rapid but short-lived and available only in summer.

Range Ecology and Management Grazing is an ecosystem process broadly defined as feeding on herbaceous plants, algae, fungi, or phytoplankton (Begon et al. 1996). While livestock grazing started in California less than 250 years ago, the rangelands have always been grazed by large and small animals, albeit in many different ways. Grazing (shown in Figure 37.3 as the interaction between primary production and primary consumption) includes three primary phenomena affecting ecosystems: defoliation, trampling, and nutrient redistribution (Jackson and Bartolome 2007). These can be precisely described using the terms “grazing pressure” (the amount of forage removed relative to availability); “grazing distribution” (where grazing occurs); “grazing period” (when grazing occurs); “selectivity” (what is grazed); and “kind and class” of animal (what kind of animal does the grazing and what is its phenological state). When described in these terms, it is apparent that for any given landscape, ecosystem, community, or management unit, grazing can be spatially and temporally complex. The needs of primary producers and consumers fundamentally conflict. Plants capture energy and support growth by growing leaves, but herbivores capture energy and protein by removing the leaves. This observation has led to the principle that the grazing manager cannot simultaneously optimize primary production and primary consumption (Heitschmidt and Stuth 1991). This principle has held up well despite observations that can be loosely grouped as describing compensatory plant growth (Bartolome 1993). Attempts to use highly complex specialized grazing systems to circumvent this principle have been many, costly, and mostly failures (Briske et al. 2008, Briske et al. 2011).

Range Livestock Production in the Mediterranean ­Climate Zone In the most typical range cattle production system (cowcalf), the animals present during the grazing year consist of a mix of older cows in various states of pregnancy and lactation, breeding bulls, and calves. Yearlings can also be present. These classes of cattle each have very different nutritional

requirements. In the inland zone the period between roughly March 1 and June 1 is referred to as the “adequate green feed period” (George, Bartolome et al. 2001) because animal needs for energy, protein, and vitamins can be met by range forage. During the rest of the year, range cattle need to either be supplemented or draw on reserves stored in their bodies. California ranchers have devised ways to adapt to the nutritional needs of livestock production. Their basic approach is to time calving in fall to allow the growing calves to take advantage of the later adequate green feed period in spring on annual range. Young nursing calves in fall utilize milk produced by the cow, which is generally allowed to lose weight and use fat reserves. Often the pregnant cows are fed a protein supplement to overcome lower forage quality during the late summer and fall dry season. Historically, ranchers with access to higher-elevation range could extend the green feed period by moving animals to mountain pastures (Huntsinger, Forero et al. 2010; Brownsey et al. 2013). Another way to effectively extend the quality of feed is to use irrigated pastures (Huntsinger et al. 2007). In the inevitable years when forage production is low, hay may be fed to maintain adequate nutritional levels in cows. Similar production schedules to adapt nutritional needs to range forage quality are seen in sheep operations and on coastal prairie, where a longer green feed season is typical.

Models for Range Community Dynamics To understand grazing, ecological models must incorporate the dynamics of vegetation change. Understanding how changes in grazing might influence the ecosystem requires understanding the many, interacting factors that shape plant communities. As the influences of relatively unpredictable and unmanageable abiotic forces like rainfall have come to be better understood, vegetation change models have shifted away from the idea that there is a single, equilibrial, climax community that represents the potential, or most pristine, community for a given site. Instead, newer models emphasize that unpredictability in vegetation change means there is no single “equilibrial” state and that several “stable” vegetation states are possible on a given site with different potentials to produce ecosystem services. While equilibrium models posit that biotic interactions such as competition and herbivory are key drivers of plant community structure, nonequilibrium models posit that plant-plant and plant-animal interactions are of minimal importance relative to abiotic constraints (Wiens 1984, Ellis and Swift 1988, Briske et al. 2003). The choice of model is critical to describing, understanding, and predicting range dynamics. (We discuss equilibrium and nonequilibrium models with more detail in Box 37.2.)

Ecological Site Classifications, State-and-Transition Models, and Adaptive Management: Conceptual Tools for Range Management Just as different definitions of the term “rangeland” can lead to vastly different estimates of how much rangeland there is, overgeneralization of ecological knowledge can lead to incorrect assumptions about possible management outcomes. Site specificity is important because rangelands are so widespread and so diverse. Over the past century, rangeland managers in the western U.S. have worked to identify and define “sites” R ange Ecosystems   845

that are similar enough to have the same potential ecological dynamics. A site is typically conceptualized as an area of homogenous soil and topography that spans 10 4 to 105 m 2 within a zone of a relatively singular climate (Fuhlendorf and Smeins 1996; Bestelmeyer, Brown et al. 2011). The goal of this work is to allow managers to predict with some certainty that similar sites will react similarly to management treatment (Brown 2010). Rangeland ecologists have promoted various elements of range ecosystems as predictors of potential vegetation and response to management on a given site. By 1919, Arthur Sampson, one of the first Forest Service range scientists, had adapted the equilibrial succession model developed by grassland ecologist Frederick Clements (Clements 1916) to rangelands as the “range condition model” (Sampson 1919). This model represented the idea that there is a single climax or equilibrium state for grasslands of a particular climate region. Livestock grazing was the cause of “range retrogression” away from the climax state, and the removal of livestock would allow for competition and other processes internal to the system to drive succession to the climax state. The status of a zone within the climatic region relative to the general, regional climax was used to predict the potential of that zone. Differences among soils across the landscape were mostly attributed to soil depletion (a result of grazing) or soil succession (a result of the plant community on the soil). Later, Dyksterhuis (1949) emphasized the importance of delineating sites by their inherent edaphic and topographic characteristics. He retained but improved climax theory by positing that each range landscape had a “polyclimax,” with each site class (known then as a “range site”) having its own climax. Grazing remained the primary posited cause for a range site’s departure from its climax. Through his “quantitative range condition model,” Dyksterhuis recommended measuring the abundance of plant species considered indicators of range condition on each site. Subsequently, advances in observation and manipulative experimentation have verified the basic idea that geology, topography, and soils are the primary governing agents of a site’s potential within a particular climate region (Grigal et al. 1999, Bestelmeyer et al. 2009, Brown 2010) and that rangeland sites have the potential to support alternative stable states (Briske et al. 2003). Consequently, using the current vegetation of a site to define that site’s potential, and the vegetation’s departure from the climatic climax or polyclimax as an indicator of the effect of livestock grazing, has been largely abandoned. Today, defining site classes using soils and topography is increasingly accepted for the rangelands of the American West (Caudle et al. 2013). This shift is reflected by the change in name of site classifications from “range sites,” used in the 1950s through the 1990s, to ecological sites, used today (Booker et al. 2012). With the state-and-transition model, Westoby et al. (1989) provided a flexible approach to describing the dynamics of managed ecological sites. These are box-and-arrow diagrams in which boxes represent theoretical or observed ecosystem states and arrows represent the theoretical or observed transitions among these states. The transitions usually describe changes through time instead of space (Bestelmeyer, Goolsby et al. 2011). A state-and-transition model can help the range manager describe, understand, and predict potential states as well as potential management impacts on those states. The models are flexible in that they can be used to describe succession resulting from both nonequilibrium and equilibrium 846  Managed Systems

dynamics; as explained in Box 37.2, nonequilibrium dynamics are generally more prevalent on rangelands. The three major U.S. land management agencies are working toward creating ecological site descriptions, and pairing each with a state-and-transition model, for the rangelands of the West (Caudle et al. 2013). Currently, federal agencies emphasize the concepts of stable states and thresholds and use recent advances in soil information and Geographic Information System technology (Brown 2010). Each federal ecological site description is linked to a particular soil type. Problematically, each soil map unit in a given soil survey contains multiple soil types that are not differentiated (Soil Survey Staff 1999), resulting in a situation in which the size of the map unit is almost always larger than the size of the ecological sites in that map unit. The main factors limiting management of California range ecosystems are patchy availability and applicability of published ecological site descriptions, largely untested management practices with poor economic justification, and unknown responses to rapid environmental change. The more traditional goals of sustainable grazing management and enhanced forage production have been joined by the need to evaluate and anticipate response of rangelands to global change and the potential for enhancing carbon sequestration and other ecosystem services. This enlarged set of goals requires an approach that incorporates realistic predictive models, improved monitoring, and well-structured adaptive management. Monitoring and assessment of range condition, using the ecological site as the basic spatial unit and stateand-transition models to understand the dynamics of the system, is recommended for semiarid and arid rangelands of the western U.S. (Herrick et al. 2006). Incorporating best-available information about how to define ecological sites, understanding their potential states through state-and-transition models, and monitoring and assessment linking the models to an adaptive management framework is an approach also gaining prominence, especially on public lands (Herrick et al. 2012, Spiegal et al. 2014). This approach recognizes the synergies and trade-offs inherent to natural resource management and emphasizes “learning by doing.”

Ecosystem Services and Livestock Production Range production depends on many ecosystem services, often with relatively little direct human manipulation, including supporting and regulating services like water flow regulation and soil nutrient cycling. The “seminatural” appearance of range is responsible for many of the values that conservation and environmental policies seek to enhance or maintain. Rangeland owners and the public both consume ecosystem services from rangelands, benefiting from the market and nonmarket benefits of seeing and living in rangeland landscapes. In California the relationship of livestock grazing to ecosystem services such as biodiversity and oak regeneration is increasingly studied, and there is strong recent interest in understanding the dynamics of rangeland carbon storage and sequestration (Booker et al. 2012). A growing body of research and management experience shows that livestock grazing can enhance biodiversity (Barry 2011, Huntsinger and Oviedo 2014). Some species can benefit from grazing that alters grassland structure to create shorter vegetation, more openings, or more structural heterogeneity in general than when livestock are excluded. Those species

include burrowing owls (Athene cunicularia; Nuzum 2005), a variety of beetles (Dennis et al. 1997), kit fox (Vulpes macrotis mutica; USFWS 2010), kangaroo rats (Dipodomys stephensi; USFWS 1997, Kelt et al. 2005, Germano et al. 2012), blunt-nosed leopard lizards (Gambelia sila), and San Joaquin antelope squirrel (Ammospermophilus nelsoni; Germano et al. 2012). To a surprising degree, this understanding stems from cases where, as part of conservation efforts, livestock grazing was removed and species or habitats of interest subsequently disappeared. Trade-offs, however, are integral to managing range for livestock production. While livestock production can provide and enhance many ecosystem services, it can also be the source of “disservices,” including adverse impacts to particular habitats, soil erosion, and nutrient and pathogen runoff. The responses of a site and its organisms to grazing are influenced by complex interaction of the abiotic environment, regional species pool, and land management of the past and present (Heady 1984). Untangling background siteand time-specific processes from the effects of management can be expensive and time-consuming, and such untangling is unfortunately often not achieved in grazing studies (but see Langstroth 1991, Dyer et al. 1996, Jackson and Bartolome 2002). The complexity of isolating the effects of grazing is exacerbated by the heterogeneity of range landscapes: grazing on one site might provide services for one suite of organisms, while grazing another site in the same landscape can result in disservices to a different suite of organisms. Moreover, the grazing process occurs at multiple spatial and temporal scales. The process affects organisms that operate at these varied scales differently. While such complexity prohibits broad generalizations, below we offer the “state of the knowledge” about the links between production of livestock and ecosystem services. We organize this discussion around the components of range and rangeland ecosystems conceptualized in Figure 37.3, including plant communities (primary producers), herbivores (primary consumers), carnivores (secondary consumers), soils, water, and the atmosphere. We highlight the relationship of livestock production to maintaining and enhancing ecosystem services as applied to different range regions and, at times, particular ecological sites within those regions.

Native Plants and Livestock Production MEDITERR ANE AN CLIMATE INL AND AND COASTAL GR ASSL ANDS

Widespread cultivation by eager homesteaders, livestock, and drought in the late 1800s contributed to the rapid spread of highly competitive species from the Mediterranean Basin into Californian grasslands. Conversion from the original vegetation to the exotics that now characterize the majority of Mediterranean climate grasslands in California has been nearly total (see Chapter 23, “Grasslands”). In many areas, native species form only a small percentage of the herbaceous cover (Biswell 1956, Heady et al. 1991, Hamilton et al. 2002). Many rangeland ecologists view this conversion as irreversible (Heady 1977, Heady et al. 1991); however, persistent native species richness, the dominant cover of natives in some areas, and the success of native species restoration on some sites is tempering that view (Bartolome, Barry et al. 2007). A statewide assessment of the changes in the Califor-

nia grassland since colonization is impossible, because the spread of livestock largely coincided with dramatic nineteenth-century species invasion, and community structure prior to these events was not scientifically recorded (see Chapter 23, “Grasslands”; Schiffman 2007b). Also lacking for California grasslands is research directly linking grazing to plant responses at the individual, population, and site levels of ecological organization. Our understanding relies largely on results from more general research conducted by range scientists studying productivity and community composition responses to grazing management (Jackson and Bartolome 2007). Early ecological research in the grasslands of California’s Mediterranean climate zone was directed toward understanding the forage base for livestock grazing (Bentley and Talbot 1948, Sampson et al. 1951, Biswell 1956). Much of this work established the primacy of location and weather as factors controlling herbaceous production and composition (Talbot et al. 1939, Heady 1958). Later studies and management practice have shown that effects of grazing are strongly related to the abundance of litter or residual dry matter remaining at the time of autumn germination (Hedrick 1948, Heady 1956 and 1965, Bartolome et al. 1980, Bartolome et al. 2006). That body of work has quantified the experience shared by California ranchers and grassland managers: management activities are successful only within the constraints of the California climate. Forage productivity can be effectively managed for livestock production via manipulation of the amount of fall residual dry matter through grazing intensity (Bartolome et al. 1980), but species composition is more or less entrained by intra- and interannual weather. A working hypothesis is that grasslands so constrained by highly variable weather are best described using nonequilibrium models (see Box 37.2, Jackson and Bartolome 2002). However, at smaller-scale sites within the grassland, such as serpentine soils and vernal pools, biotic interactions like grazing can impose stronger controls on species composition than weather (see Box 37.3). Despite the challenges presented by erratic climate and weather, California grassland managers are increasingly using livestock as a management tool for native species restoration (Stahlheber and D’Antonio 2013). Here we describe responses of native plants to grazing as identified in studies that did not explicitly compare the effects of rainfall fluctuations over time with the effects of grazing over time (but see Box 37.3). Stahlheber and D’Antonio (2013) used meta-analysis to identify impacts of livestock grazing on diversity and cover of natives in the grasslands of California’s Mediterranean climate zone. They chose to analyze valley grassland and coastal prairie together despite the differences between the two grassland types because both have been dramatically invaded by exotic species, and as a result, managers of both have comparable restoration goals. Grazing increased forb (broad-leaved herb) cover, most reliably exotic low-growing forbs like filaree and cat’s ear (Hypochaeris spp.), especially at the more arid study sites. Native forb cover responded more variably than exotic forb cover, increasing more in response to wetand dry-season grazing as opposed to continuous/year-round grazing, and more in inland sites than in coastal sites. The authors hypothesize that the difference between coastal and inland site responses stems from the presence of native perennial forbs on the coast and their absence inland. Although the effect of grazing on grass cover was variable, grazing increased cover of native perennial grasses more reliably than R ange Ecosystems   847

cover of exotic annual grasses. Wet-season grazing produced a strong response, diminishing exotic annual grass cover and increasing native perennial grass cover. The authors conclude that site specificity is a critical consideration but that grazing can be compatible with efforts on some sites to enhance and restore native forb and grass abundance, if an increase in exotic forbs is acceptable. The extent and richness of native species growing in particular types of sites within the grasslands of the Mediterranean climate zone appear to increase with grazing. For instance, serpentine-derived soils are refugia for native species (Harrison and Viers 2007), and grazing appears to enhance native forb species richness in these soils (Harrison et al. 2003, Pasari et al. 2014). Vernal pools, specialized seasonal wetlands nested within the grasslands of the Central Valley, are another specific site type in which grazing enhances the native community. Vernal pools are depressions underlain by soils with an impermeable layer of claypan, hardpan, cemented mudflow, and/or rock that fill with fall and winter rains and runoff (see Chapter 31, “Wetlands”; Rains et al. 2008). As temperatures warm in the spring, the standing water evaporates and concentric rings dry individually over time, exhibiting remarkable wildflower shows by plants specifically adapted to the water regime of particular bands. By summer, the pools are dry and the soils are at permanent wilting point until the rains come again in the fall (O’Geen et al. 2008). Vernal pools are highly valued as habitat for endemic, rare, and endangered plant and animal species specifically adapted to the inundation regimes. They also prevent regional flooding, regulate groundwater recharge, contribute to the phosphorus and nitrogen cycles, and provide prey for migratory birds (Hobson and Dahlgren 1998). Marty (2005) investigated the effects of cattle grazing on the native flora and fauna of pools in valley grassland in eastern Sacramento County. Removal of livestock grazing for three years increased cover of exotic annual grasses. Endemic plants, largely annuals, were “choked out” by the exotic grasses, and native species cover and richness declined. Reintroducing grazing resulted in increased native vegetation species richness and cover. Grassland patches in the uplands between the pools also supported higher richness and cover of native species when grazed. Within-pool aquatic invertebrates species benefited from grazing as well, because exotic annual grasses were removed and inundation could occur. The grazing regime with the strongest effect was continuous October–​ June grazing rather than wet-season or dry-season grazing. In an example of an unexpected synergistic relationship, rancher-created stock ponds provide habitat to replace the many lost vernal pools. Half of the remaining habitat for the endangered California tiger salamander (Ambystoma californiense) in the San Francisco Bay area is found in stock ponds (USFWS 2004), and a variety of “payment for ecosystem services” and mitigation initiatives help ranchers to maintain the ponds and support the species. Some amphibian species, including the endangered California red-legged frog (Rana draytonii) and the California tiger salamander, also seem to benefit from grazing that reduces vegetation near the ponds (DiDonato 2007).

MOJAVE DE SERT GR ASSL ANDS

The Californian portion of the Mojave Desert (see Chapter 30, “Deserts”) has undergone drastic land use change over the 848  Managed Systems

last fifty years. The region’s human population increased by 350% between 1970 and 1990 (Hunter et al. 2003, Berry et al. 2006). Human use, often characterized as a “disturbance,” is associated with dramatic invasions of annual grasses and forbs and their attendant altered fire regimes (Brooks and Matchett 2006), dangerous dust storms (Grantz et al. 1998), and loss of habitat for valued species (Inman et al. 2013). Livestock grazing is often considered a disturbance in line with fast-paced urbanization, off-highway vehicle use, wind and solar energy developments, and crop agriculture (Brooks et al. 2006, Lovich and Bainbridge 1999). The effects of grazing, however, have not been investigated nearly as thoroughly as they have been in the grasslands of the Mediterranean climate zone to the west. Grasslands cover approximately 85,000 hectares of the Californian Mojave, located mostly in the western portion known as the Antelope Valley (Menke et al. 2013). Like in their westerly counterparts, different scales of observation reveal different responses to grazing. For instance, Brooks et al. (2006) found significant effects of cattle grazing on species composition within 200 meters of artificial livestock watering sites in the west-central Californian Mojave, whereas a multiyear observational study across 20,000 hectares of grasslands on Tejon Ranch in the western Mojave Desert indicated that grazing did not affect species composition nearly as much as did interannual rainfall timing and amount (Spiegal and Bartolome unpublished data). A pressing need exists to describe and understand the effects of grazing on Mojave Desert grasslands before changes intensify.

MONTANE ME ADOWS

Grazing impacts in California’s montane meadows have been a concern since the late 1800s, when John Muir famously described sheep in the Sierra Nevada as “hoofed locusts” (Muir 1894). Teasing apart the specific impacts of grazing has proven difficult, however, as montane meadow systems are immensely complex (Ratliffe 1985) and grazing introductions coincided with widespread water diversion, water application for mining, a huge population boom as mining camps populated the high country, and burning for clearing by miners, farmers, and graziers. Abiotic factors, especially hydrology, elevation, and weather, control the systems to a large degree, and vegetation varies from year to year and over small spatial scales even within a single meadow (see Chapter 29, “Alpine Ecosystems”; Fites-Kaufman et al. 2007). Despite this complexity, studies have illustrated trends in grazing effects on montane meadow vegetation. Livestock grazing can change plant species composition and production both directly, through grazing and trampling, and indirectly, through stream incision and lowered water tables. Important factors affecting whether species composition or productivity will more strongly respond to grazing in Sierra Nevada montane meadows appear to be location (northern, central, or southern) and grazing intensity. Grazing-induced changes in species composition have been detected over both long and short time scales. Stratigraphic pollen records from a montane meadow complex in the upper reaches of the South Fork Kern River, on the Kern Plateau, show that species composition changed in the region during the last 150 years following the widespread introduction of livestock. Willows (Salix spp.) and liverwort (Riccia sp.) declined significantly coincident with the introduction

of livestock during the Gold Rush period. Pollen from sedges (Cyperaceae) and silver sagebrush (Artemisia cana) increased over this period (Dull 1999). In the same area, on a shorter time scale, Stevenson (2004) examined the impacts of seasonal cattle grazing on montane meadow hydrology and flora in nine pairs of grazed and ungrazed meadows. Her observational study found that livestock grazing was associated with changes in channel morphology and soil-moisture class distribution. Although plant species richness was not affected by grazing, species composition differed between ungrazed and grazed meadows, with tall forbs less common in the latter. A study further north, in Yosemite National Park of the central Sierra Nevada, found that whether grazing affects productivity or species composition can be most related to grazing intensity. Using a five-year recreational packstock grazing experiment, Cole et al. (2004) concluded that when grazing impact is light, productivity and ground cover are more strongly affected than species composition. In three characteristic meadow types, they found that productivity was reduced by about 20%, vegetative cover was also reduced, and bare ground increased in grazed as compared to ungrazed plots.

SHRUBL ANDS

California’s chaparral and coastal sage scrub communities are managed as range to a lesser degree than other rangeland types such as grasslands and oak savanna. Shrublands of California’s southern coast were converted to grassland to improve forage by European settlers, but the climate and short growing season proved to be less than ideal for livestock production (see Chapter 22, “Coastal Sage Scrub”). Some evidence indicates that these converted areas have largely changed back to coastal sage scrub. Like the coastal sage scrub, conversion of chaparral to grassland has been attempted, especially by seeding grasses after fires (Mooney et al. 1986; see Chapter 24, “Chaparral”). Compared with other Mediterranean climate regions globally, such conversions have been largely unsuccessful (Mooney and Parsons 1973). Because livestock grazing is somewhat uncommon on chaparral itself, understanding the impacts of grazing on chaparral vegetation might require expansion to a landscape view. Again, disentangling the effects of grazing from other factors can complicate this understanding. For instance, in a 2003 study Keeley and others posited that blue oak (Quercus douglasii) savannas, which grow in close proximity to chaparral in the southern Sierra Nevada, provide an important source of alien annual grass seeds that colonize chaparral sites postfire (Keeley et al. 2003). They were not able to find differences between grazed and ungrazed savanna, however; alien species richness and cover do not differ much between oak woodlands that are grazed and those in which grazing was removed a century ago. Accordingly, the effects of grazing on postfire chaparral could not be isolated.

OAK SAVANNA

Grazing might influence oak recruitment and survival in oak woodlands, a concern because replacement of oaks lost to age, disease, and harvest is suspected to be inadequate in some areas. In the Sierra Nevada foothills, for example, blue oak–​ foothill pine (Pinus sabiniana) woodland/savanna is the char-

acteristic landscape. Blue oak woodland/savanna is also widespread through the Coast Range and Cascade foothills. Its herbaceous understory is for the most part non-native, annual grasses and forbs. Some surveys over the last few decades have noted a lack of sapling-sized blue oaks (Allen-Diaz et al. 2007, Zavaleta et al. 2007); if inadequate sapling recruitment persists, blue oak stands could begin to thin and disappear as adult trees die but are not replaced. Research has shown that grazing by both livestock and wildlife reduces growth and survival of blue oaks. Especially when range is grazed during the summer, livestock may browse on seedlings (McCreary and George 2005). Protection of seedlings as they move into the sapling stage might be necessary for successful maintenance of blue oak stands (Allen-Diaz et al. 2007); use of “treeshelters”—​individual, translucent plastic protectors that fit over oak seedlings and are secured with a metal fence post—​have proved successful (McCreary 2001). Grazing might indirectly help blue oak seedlings by reducing competition with annual grasses and forbs (Tyler et al. 2006). Coast live oak (Quercus agrifolia) woodland/savanna occurs within 100 kilometers of the coast (see Chapter 25, “Oak Woodlands”). Canopy cover can vary from open to dense, and the herbaceous understory, if present, comprises largely non-native grasses and forbs. The evergreen coast live oak appears fairly resistant to livestock grazing and could replace less resistant deciduous oaks in areas with high grazing pressure. However, the sudden oak death pathogen (Phytopthora ramorum) (see Box 13.2 in Chapter 13, “Biological Invasions”) has caused widespread mortality of coast live oak, which will likely affect this woodland type in the future (Allen-Diaz et al. 2007). In the past, oaks were removed to increase forage production. The University of California and other range management advisers encouraged this. As described earlier, studies have subsequently shown that forage can remain green longer under an open oak canopy and production is just as high as in the open. As part of the Integrated Hardwood Range Management Program, from 1985 to 2010 University of California outreach promoted the benefits of oaks. These include the contribution of oaks to maintaining property values, increasing wildlife habitat, and in some cases, extending the green forage season for grazing. Outreach efforts were linked to an understanding of rancher needs and values derived from survey research. Over the period of the program’s duration, oak planting by landowners there increased and cutting declined (Huntsinger, Johnson et al. 2010).

R ANGE RIPARIAN ZONE S

Use by livestock of riparian areas is more often spatially concentrated than their use of upland or dryland range. Riparian areas can include the only sources of water on the range and offer shady or cooler places to rest. Cattle in particular also resist climbing steep slopes and may congregate in low or flat areas. On the other hand, they will usually avoid muddy areas if possible. Widespread concern about the impacts of grazing on riparian ecosystems in the American West (Belsky et al. 1999) has inspired much study. Care should be taken when extrapolating study results from climates that differ from the arid and/ or semiarid Mediterranean climates of Californian rangelands (Jackson and Bartolome 2007). In addition, landscape heteroR ange Ecosystems   849

geneity is a critical consideration when assessing the impacts of livestock on riparian zones. In the grassland landscapes of California’s inland Mediterranean climate zone, trampling effects are stronger in riparian zones than in grassy uplands because livestock preferentially use areas near shade and water sources (Tate et al. 2003). This effect can be patchy, however, because oak tree canopy can mitigate trampling effects by increasing organic matter through litterfall (Tate, Dudley et al. 2004). In addition, at spatial scales nested within landscapes, for instance within riparian zones, abiotic factors like flooding—​not biotic interactions like livestock grazing—​a re primary controls of vegetation patterns (Stringham and Repp 2010).

INVASIVE GR ASSE S AND FORBS

Controlling invasive plants has proven to be one of the greatest challenges facing California range managers (Stromberg et al. 2007). In valley grasslands and coastal and foothill oak savannas, cover of non-native grasses and forbs commonly exceeds 90% (Bartolome, Jackson et al. 2007). The effects of plant invasions on California’s range ecosystems have been profound and continue today. Obviously, species composition and dominance have changed, but so have many ecosystem processes, including hydrologic and nutrient cycles and fire regimes (D’Antonio et al. 2007). However, most common, non-native grasses and forbs are no longer invasive; rather, they have completely taken over and are naturalized species. Of greater concern to range managers is the expansion and increasing abundance of newer species that continue to invade California rangelands. The three most troublesome invasive rangeland plants in California at present are yellow starthistle (Centaurea solstitialis; see Box 13.1 in Chapter 13, “Biological Invasions”), meadusahead (Elymus caput-medusae), and barbed goatgrass (Aegilops triuncialis). Range dominated by these species often has greatly reduced forage value for both livestock and wildlife. In addition, native plant populations, wildlife habitat, recreation, and other ecosystem services are often negatively affected (D’Antonio et al. 2007). Livestock have both positive and negative impacts on invasive species control. Livestock production has been implicated in the introduction of weeds into noninvaded areas, as invaders either hitch rides on livestock or contaminate hay fed to livestock. Concentrated livestock use can increase the cover of bare ground, which can provide favorable germination sites for weeds. On the positive side, livestock grazing is an invasive plant management tool available to range managers; however, a single weed management tool often does not result in successful control (DiTomaso et al. 2007). To increase likelihood of successful long-term control, experts recommend combining several weed management methods tailored to situation-specific goals, constraints, and opportunities (DiTomaso et al. 2007). Using livestock to control invasive plants often requires prescription grazing, which is the application of specific livestock grazing actions to accomplish specific vegetation management goals. Grazing intensity, animal distribution, and grazing period often differ from standard grazing practice, and livestock performance can be significantly reduced (Germano et al. 2012). Consequently, finding a livestock operator willing to implement a grazing prescription can prove difficult and might require reduced grazing fees or even payment 850  Managed Systems

to the operator. Intensive grazing, sometimes necessary for successful weed control, can have undesirable consequences. For example, concentrated hoof impacts and greatly reduced vegetative cover could result in increased soil erosion; as noted earlier, greater bare ground might also allow other weed species to thrive. In addition, intensive grazing can significantly impact desirable species in the weed-infested area. Those caveats noted, prescription grazing can work well in controlling some invasive species (DiTomaso et al. 2007). An essential planning factor is that prescription grazing needs to be timed to the target weed species’ phenology. Grazing must occur when weeds are most vulnerable to defoliation; poorly timed grazing can actually benefit target species (Huntsinger et al. 2007). Correctly timed livestock grazing can be used to control yellow starthistle effectively (DiTomaso et al. 2006). Reducing residual dry matter through livestock grazing (or fire) creates unfavorable conditions for medusahead, providing some measure of control. Goatgrass is only palatable to livestock in early spring, so it typically must be managed with other methods such as fire or herbicide application. Grazing can also be used to manage velvet grass (Holcus lanatus), one of the most troublesome perennial grass weeds in the coastal prairie grassland type (Hayes and Holl 2003).

FUEL MANAGEMENT AND WOODY PL ANT ENCROACHMENT

Woody plant encroachment and the fine fuels created by dry grasses can exacerbate fire hazard. Fine fuels left on range in summer and fall promote a higher probability of ignition and rapid fire spread. Cattle, sheep, goats, and horses can all be used to reduce fine-fuel loads. The effectiveness of grazing on fire behavior has not at this point been quantified but is inferred from the removal and alteration of fuels (Stechman 1983). Grazing is an alternative to prescribed fire for fire hazard reduction at wildland-urban interfaces and in other situations where risk of fire escape cannot be tolerated. Goat herders usually charge to graze for vegetation management, and goats will concentrate on shrubs, while cattle and sheep graziers typically pay for grazing leases, and cattle and sheep concentrate on grasses and forbs. Management emphasizes control of shrub encroachment and fine fuels while protecting other resources. Control of woody vegetation can be desirable from a conservation or biodiversity perspective as well. Coastal prairie grasslands are highly subject to woody plant invasion, reducing the extent of native perennial grasslands. Grazing can help prevent encroachment of shrubs and trees. Coastal prairie intergrades with several shrub and forest community types that tend to encroach on open grassland in the absence of grazing and fire (Ford and Hayes 2007). The native shrub coyote brush is a primary offender in this regard, and cattle (McBride and Heady 1968, McBride 1974) as well as the native grazer tule elk (Johnson and Cushman 2007) significantly reduce cover of coyote brush in open grasslands if grazing takes place when shrubs are small.

R ANGE MANAGEMENT IMPLICATIONS

The wildlife and plants present on Californian range today have coexisted with more than two hundred years of livestock grazing that has at some times been extreme. These

species have also persisted despite increases and decreases in fire frequencies and extensive water diversions. The most unprecedented ecological change on rangelands today is that the range is subject to reduced intensity and even removal of livestock grazing from areas for various purposes, while at the same time these areas rarely are able to duplicate the levels or types of grazing that occurred before the introduction of livestock. No real knowledge base is available for predicting the long-term impact of reductions in herbivory and fire over large areas. The primary message could be that changing traditional grazing practices, whether to increase, reduce, change the season of or eliminate grazing, should be approached with caution and carried out incrementally. The secondary message is that grazing effects, desirable or undesirable, depend on multiple, site-specific factors—​ another reason for incremental, adaptive management and for avoiding generalization. The interannual variability of rainfall is also a constant and very powerful factor modifying the effects of grazing on the California range, and desirable outcomes in one year might very well be followed by undesirable outcomes the next. Clear opportunities for livestock-related ecosystem services in California involve restoration in Mediterranean climate grasslands, conservation of oak savanna, weed management, and rancher husbandry and management. Researchers and managers must document results and link outcomes to specific ecological sites for them to be useful in future management. Managers need to draw on this information and use adaptive, incremental approaches. Finally, there are ways to manipulate grazing impacts on vegetation. Keeping many animals in a small area—​for example, in a rotational grazing plan with high stocking rates—​ can result in more even grazing. Because each animal has less choice about what to eat, animals will eat less preferred plants. This could be a useful strategy if a manager wants animals to graze weedy plants that are not preferred or when an even level of utilization is desired. On the other hand, when managing for diverse vegetation structure, or when animal selectivity or choice is beneficial in meeting management goals, allowing fewer animals to graze longer on larger areas can be useful, as in a year-round or season-long grazing plan. For example, in grazing vernal pools, giving animals time and space to make the choice of when to use the plants in the pools and to select grass rather than rare forbs is important. High-intensity, short-duration grazing that crowds animals in small areas for short periods of time would be inappropriate. In general, the more the animals prefer or select a target species, the easier it is to use grazing to manage that species. Knowledge of the preferences of different types of animals is very useful in developing these kinds of prescriptions.

Native Herbivores and Livestock Production Research has shown that livestock grazing can compete with, facilitate, or not affect native herbivores. Burrowing mammals, as well larger grazing animals, are perhaps most directly influenced. Whether livestock grazing has beneficial or harmful effects depends on the wild herbivore species under consideration. Small burrowing mammals are important players in the function of California’s grassland ecosystems. While gophers do not increase with livestock grazing, ground squirrels and kangaroo rats do. Among the larger wildlife species, a comparison of wild ungulates with their domestic counterparts indicates that they share preferences for forage. In range

operations in which wildlife conservation is a goal, these interactions should inform management planning.

BURROWING MAMMALS

In 1923, Grinnell estimated that 1 billion burrowing mammals lived in California (Grinnell 1923). These small mammals are considered ecosystem engineers because of their contributions to soil disturbance, seed dispersal, granivory, and herbivory (Schiffman 2007a). Prevalent species include the California Beechey ground squirrel (Spermophilus beecheyi), pocket gopher (Thomomys bottae), and kangaroo rat (Dipodomys spp.). Responses to livestock grazing tend to be speciesand site-specific. Ground squirrel populations increase with livestock grazing in inland grasslands (Fitch and Bentley 1949, Howard et al. 1959), coastal grasslands (Linsdale 1946), and blue oak savanna with annual grass understory (Bartolome 1997). Howard (1953) hypothesized that livestock-induced litter reduction favors the germination of broadleaved plants that are desirable to squirrels. A reevaluation could be appropriate, incorporating knowledge about nonequilibrium dynamics prevalent on drier rangelands (see Box 37.2). Concern that ground squirrels compete with livestock for forage (Fitch and Bentley 1949, Howard et al. 1959) and create leg-breaking burrows dangerous to livestock and humans (Marsh 1998) resulted in eradication efforts (Howard 1953). Ground squirrel eradication has been controversial because of the ecological services they provide, including soil formation; burrow habitat for rare wildlife species like burrowing owls and California tiger salamanders; and prey for raptors, coyotes, and rattlesnakes (Davidson et al. 2012). Pocket gopher activity is spatially and temporally complex but pervasive in California grasslands. In nonserpentine, annual-dominated grasslands in Monterey County, gophers occur at densities of 26–​125 ha-1 (Lidicker 1989, Stromberg and Griffin 1996). Hobbs and Mooney (1995) found that no part of the soil was left undisturbed for more than five years during their eleven-year study of gopher disturbance in coastal grasslands on serpentine-derived soils. Gophers affect the composition and arrangement of vegetation in grasslands (Hobbs et al. 2007). Native bunchgrass seeds have low success rates on gopher tailings (Stromberg and Griffin 1996); soil disturbance by burrowing mammals in general tends to favor annuals (Schiffman 2000). In a study on interactions of cultivation, gopher disturbance, and cattle grazing on grassland composition and structure in Monterey County coastal range, Stromberg and Griffin (1996) found gophers inhabit a wide variety of soil textures, produce the same amount of tailings in cultivated as they do in uncultivated areas, and produce more tailings in ungrazed areas than in grazed areas. An increase in gopher activity with livestock removal also occurred in Vina Plains (Hunter 1991) and Jasper Ridge (Hobbs and Mooney 1991). Population densities of kangaroo rats tend to increase in grazed areas (Reynolds and Trost 1980, Bock et al. 1984, Jones and Longland 1999, Germano et al. 2012). Research conducted at the Carrizo Plain National Monument indicates that the giant kangaroo rat (Dipodomys ingens) decreases the abundance of exotic annual grasses due to preferential granivory on their seeds. Depending on plant-plant interactions and the importance of these interactions at a particular site, the suppression of these exotics could enhance native species R ange Ecosystems   851

communities (Gurney 2012). Intensive livestock grazing benefits kangaroo rats and an associated suite of special-status vertebrates that inhabit grasslands of the San Joaquin Valley, including the San Joaquin kit fox and blunt-nosed leopard lizard (Germano et al. 2012).

L ARGE NATIVE HERBIVORE S

Ungulates and other range herbivores graze selectively, meaning they choose some plants over others. Livestock and large native herbivores overlap in their dietary preferences. Cattle, horses, and tule elk prefer grasses (McCullough 1969, Heady and Child 1994); sheep prefer forbs (Bartolome and McClaran 1992); deer (Odocoileus hemionus) prefer browse (Gogan and Barrett 1995); and pronghorn can overlap with all the other species because they prefer grasses, forbs, or shrubs depending on season (Yoakum and O’Gara 2000). Interactions between livestock and native ungulates vary widely but are not addressed here due to space constraints. An extensive review of studies on impacts of livestock grazing and other range management practices on wildlife in the West was recently conducted by Krausman et al. (2011).

Native Carnivores and Livestock Production During the late Pleistocene (19,000–​17,000 ya), mammalian carnivores were diverse, and at least eleven species were as large or larger than the coyote (Canis latrans; Edwards 1996; see Chapter 9, “Paleovertebrate Communities”). In large part, the carnivores’ food base comprised small burrowing mammals (Schiffman 2000) and large mammalian herbivores, which were also diverse with at least eighteen species (Edwards 1996). A massive extinction occurred during the dramatic climatic changes of the early Holocene approximately 10,000 years ago (Edwards 1992, 1996). A handful of carnivore species, most of the small burrowing mammal species, as well as three (elk, deer, and pronghorn) of the eighteen large herbivores, survived the extinction (Bartolome, Barry et al. 2007). Throughout the Holocene, Native Californians hunted these animals, possibly intensively enough to have kept their populations down (see Chapter 10, “Indigenous California”). Archeological evidence suggests that populations of large herbivores and predators were larger at the time of European contact than the long-term norm, possibly due to the concurrent decimation of indigenous peoples and their hunting practices (Schiffman 2007a). The earliest historical accounts of California’s natural resources convey a high abundance and widespread distribution of burrowing mammals, large herbivores, and carnivores (Minnich 2008, Schiffman 2007a). Concerns about predation on livestock have led livestock producers and government predator control agents to use lethal methods to suppress predators. In 2010 in California, about 9,600 head of cattle (including calves) were lost to predators, equaling about 1% of the total head of cattle in California. Coyotes were responsible for 57% of the deaths, while mountain lions (Puma concolor) and bobcats (Lynx rufus) were responsible for 33% (NASS 2011b). During the nineteenth and twentieth centuries, regulations on control methods were few, and poison, trapping, and shooting were employed. Many lament the resulting extirpation of the grizzly bear (Ursus arctos), which was likely a keystone species before Euro852  Managed Systems

pean settlement (Schiffman 2007a). With the 1998 passage of Proposition 4, poison and leg traps were banned (Timm et al. 2007). Today the law stipulates that a person may kill a coyote on sight, using approved methods, but a depredation permit must be obtained to kill a mountain lion. The predators that survived the extinction event between the Pleistocene and Holocene can cause problems for people making their living from California’s natural resources. As a result, predator control efforts, both off the record and official, have been extensive over the past two centuries. However, lethal predator control is controversial in California. Along with the ecological services that predator species provide, Californians also tend toward conserving predators for their intrinsic value (Wolch et al. 1997, Fox 2006). Ranchers and other rangeland managers are mitigating this conflict by employing innovative, nonlethal control methods (Andelt 2004, NASS 2011b, Shivik 2006).

The Soil System and Livestock Production Livestock grazing can influence physical, biological, and chemical components of the soil. One way is through the amount of dry plant matter left on the soil after the grazing season. Research has shown that residual dry matter can protect the soil from erosive forces (Bartolome et al. 2006, Tate et al. 2006), return organic matter to the soil, and affect the next year’s plant germination (Heady 1956). By managing grazing to leave particular amounts and patterns of ungrazed plant matter behind at the conclusion of an annual grazing cycle, the manager has the best opportunity, within the confines of weather conditions and other abiotic factors, to influence the next year’s germination and to protect the soil. The amount of residual dry matter recommended for protecting soils and forage quality varies with rainfall, slope, soil characteristics, and other factors (Bartolome et al. 1980). Prescriptions are necessarily unique to the ecosystem and the goals of the manager.

BULK DENSIT Y

The weight of a beef cow is approximately 450 kilograms; this can compact the soil, reducing porosity and increasing bulk density (mass of soil particles/volume of soil). Undesirable effects of increased bulk density on rangelands include reduced root growth (Brady and Weil 2002) and infiltration rates (Daniel et al. 2002, Pietola et al. 2005), both of which can increase surface runoff (with and without pollutants) and erosion (Blackburn 1984). Site specificity strongly influences livestock impacts on bulk density. Soil texture, which varies widely over short distances in California, is a major control on bulk density and its response to grazing. For example, clayrich soils naturally tend toward lower bulk densities because clay-sized particles decrease average pore size, increasing the “volume of soil” factor in the bulk density equation (Brady and Weil 2002). No meta-analysis quantifying the impacts of livestock trampling on soils of different textures has been conducted, but increases in bulk density in grazed compared to ungrazed sites on Mediterranean climate range have been found on coarse sandy loams of the San Joaquin Valley (Tate, Dudley et al. 2004), sandy loams of Salinas and Carmel Valleys (Steenwerth et al. 2002), and clay loams east of Berkeley (Liacos 1962). Changes in bulk density can persist for cen-

turies in some soils (Sharratt et al. 1998), but bulk density decreased after just six years of grazing removal in the coarse sandy loams of the San Joaquin Valley.

SOIL BIOTA

The impacts of livestock grazing on soil biota vary. The microbial community appears largely unaffected, at least in some sites, while cryptogamic crusts are generally negatively affected. An investigation in the Salinas and Carmel Valleys of the Central Coast of land-use effects on soil microbial biomass and composition indicated that total microbial biomass and composition vary more as a function of grassland vegetation growth habit (annual versus perennial) and cultivation history than of grazing history (Steenwerth et al. 2002). Cryptogamic soil crusts—​assemblages of algae, fungi, mosses, bacteria, and/or liverworts coexisting mutualistically—​provide nutrient cycling, nitrogen fixing, water conservation, and primary productivity (Brady and Weil 2002). Though often associated with arid rangelands and deserts (see Chapter 30, “Deserts”; Dunne 1989), they are also found on Californian Mediterranean climate (Fierer and Gabet 2002) and montane range sites. For instance, in his study of the pollen record of the Kern Plateau, Dull (1999) found that a liverwort decreased dramatically in the period after cattle and sheep grazing was initiated and proposed that cryptogamic crusts likely disappeared from his site due to direct (e.g., grazing, trampling) and indirect (e.g., stream incision and lowered water table) effects of livestock.

BIOGEOCHEMICAL CYCLE S

Grazing alters biogeochemical cycles because herbivores mineralize organic matter and return it to the environment in solid, liquid, and gaseous forms (Hack-Ten Broeke and Van der Putten 1997). In general, grazing in grasslands accelerates carbon and nutrient cycling by effectively bypassing the microbial decomposition pathway (see Figure 37.3a; Singer and Schoenecker 2003). This acceleration happens in a spatially heterogeneous manner because livestock use some areas preferentially and because their excreta is deposited in patches that make up a small fraction of the grazed landscape (Tate et al. 2003).

Water Quality and Livestock Production Protecting water quality is an important range management concern. The location of significant areas of California’s rangelands between the Sierra Nevada snowpack and the state’s major river systems means that almost all surface water in California passes through rangeland; in addition, twothirds of the state’s major reservoirs are located on rangeland (Harper et al. 1996a). Nonpoint source pollution, the diffuse discharge of pollutants throughout the environment, is the primary pollution problem on rangelands. Four main nonpoint source pollutants occur on rangelands: sediment, nutrients, heat that causes elevated water temperatures, and pathogenic organisms. All four can degrade water quality for fish, wildlife, and human uses (Harper et al. 1996b). In general, concentrations of livestock in and near water should be minimized unless such concentrations meet specific management

goals. Herding and fencing can both be used to better distribute livestock. Buffer strips can be used to filter out pollutants before they reach water. Research shows that normal grazing management practices may be adequate to protect water quality and aquatic habitat in some cases. Sediment is the pollutant most common on rangelands (George, Larson-Praplan et al. 2011) and often results from improperly constructed roads or animal confinement areas located too close to water bodies, rather than directly from livestock grazing (Weaver and Hagans 1994). Existing road building and corral placement methods can significantly reduce rangeland sediment pollution if applied (George and Jolley 1995). Direct livestock impacts on water quality are a concern, but generalizations should be avoided. While direct inputs of animal excrement in small ponds in areas with intensive grazing can increase ammonia and nitrite (Clausnitzer and Huddleston 2002, Knutson et al. 2004), grazing can actually enhance the ability of riparian vegetation to filter nitrate out of surface waters. In a grazing removal experiment in the Sierran foothills grassland–​oak savanna, soil water nitrate was five times higher in plots from which grazing had been removed for two years than in grazed plots, perhaps because of reduced nitrate use by plants (Jackson et al. 2006). Livestock impacts on overhanging riparian vegetation can raise water temperature, but properly managed livestock grazing practices should be able to minimize such impacts. A study of montane meadows at Yosemite National Park compared grazed pools to pools where grazing was excluded to evaluate effects on Yosemite toads (Anaxyrus canorus). Overall water quality was high and unaffected by fencing, and conditions for toad breeding pool habitat did not improve following fencing and cattle exclusion compared to standard U.S. Forest Service grazing management. Toads were more likely to use warmer, shallower, and more nitrogen-enriched pools, contrary to the expectations of the researchers (Roche et al. 2012). Studies in the same area found no detectable effects of grazing treatments on tadpole, young of the year, or pool occupancy (Allen-Diaz et al. 2010, Lind et al. 2011). As with wildlife and human feces, livestock feces can contaminate drinking water with pathogenic organisms such as E. coli, Giardia, Salmonella, and Cryptosporidium, all of which can cause serious disease in humans. Leaving several meters of ungrazed vegetation around water bodies greatly reduces any movement of pathogens into water, so a fenced buffer strip is an effective management practice (Tate, Pereira et al. 2004; Tate et al. 2006); buffer strips can also reduce nutrient and sediment transport into water bodies. In general, practices that reduce livestock concentration in riparian areas will also help minimize nonpoint source pollution in areas where it is a concern (George, Jackson et al. 2011). To comply with the 1972 Clean Water Act, the Environmental Protection Agency and state water resources boards developed total maximum daily load (TMDL) standards, the maximum amount of a pollutant that a water body can receive and still meet water quality standards. In response, the California livestock industry and public agencies worked with the State Water Resources Control Board to address water quality impacts on rangelands, resulting in the 1995 California Rangeland Water Quality Management Plan. The plan detailed voluntary rangeland livestock production “Best Management Practices” to protect water quality from nonpoint source pollution. In 2004 the State Water Resources Control Board adopted new nonpoint source pollution policies, which replaced voluntary compliance with regulatory R ange Ecosystems   853

programs that implemented TMDL requirements on agricultural lands including rangelands (George, Larson-Praplan et al. 2011). Watershed and agricultural groups, with assistance from state and federal agencies, are working with Regional Water Quality Control Boards to meet these TMDL standards.

Atmosphere and Livestock Production Worldwide, livestock meat and milk production processes represent approximately 7.1 gigatonnes CO2 -equivalent yr-1, or 14.5% of all anthropogenic greenhouse gas emissions (Gerber et al. 2013). Generally, systems that include more grass-based, extensive production (range) emit less total greenhouse gas than systems with more intensive feeding (feedlots) (Subak 1999, Casey and Holden 2006). Primarily because they are so extensive, rangelands contain significant carbon stocks. Grasslands alone have the potential to partially offset the emissions of meat and milk production by sequestering 0.6 gigatonnes CO2 -equivalent yr-1. Factors currently hindering the realization of the potential include lack of reliable measurement techniques and questions about economic viability (Gerber et al. 2013). In terms of long-term carbon storage, rangelands can be superior to forests because relatively more of total carbon is stored in the soil (White et al. 2000, Paruelo et al. 2010) where it is usually better protected from atmospheric release (for example, by wildfire) than carbon stored in vegetation. However, carbon inputs in rangelands—​i.e., net carbon flows from the atmosphere—​are comparatively small, and soil carbon is more difficult to measure than carbon in trees. The nonequilibrium characteristics of arid ranges mean that yearto-year sequestration is difficult to predict and that in many years carbon may actually be emitted rather than stored, as plants decompose on the surface (Booker et al. 2012). Timing and amount of rainfall, temperature variations, and soil type—​the most important factors influencing carbon sequestration on such ranges—​a re not amenable to management (Westoby et al. 1989, Parton et al. 1994, Briske et al. 2003 and 2005, Booker et al. 2012). Opportunities to increase carbon sequestration on rangelands are highly variable and best predicted at a coarse scale by the position of an ecological site along an aridity gradient. The magnitude of carbon sequestration and management influences diminishes with decreasing rainfall (Booker et al. 2012). Nevertheless, proposals for managing rangelands for climate change mitigation are gaining attention at state and federal levels in the United States. Rangeland livestock producers, generally operating with low and variable financial returns, continue to express considerable interest in diversifying income streams to include payments related to carbon sequestration (Diaz et al. 2009). Range managers are in a unique position to offset emissions from livestock production through carbon sequestration. More work is needed to develop reliable measuring techniques and an economic system that appropriately compensates those enhancing carbon sequestration. In the meantime, manipulating the amount of woody vegetation through grazing, where feasible, remains an intervention opportunity that is manageable, tractable, and likely has a significant effect on carbon stocks. However, more needs to be known about the effects of these vegetation state changes on carbon, especially soil carbon, in different ecological sites, and how to balance increases in aboveground carbon stores with possibly higher fire probabilities. The con854  Managed Systems

sequences of altered disturbance regimes also must be evaluated, or short-term gains may result in long-term loss (Booker et al. 2012). As carbon is stored in rangeland soils, management that protects soils is crucial.

Livestock Operators and Ecosystem Services Livestock producers own the majority of the Mediterranean rangelands of California. While many of the ecosystem services that ranchers seek from their own land are consumed only by them—​t he ability to leave land to their heirs, host people at the ranch, enjoy a rural lifestyle, and work with animals—​many of the ecosystem services produced on private land are shared with and valued by the public. These include wildlife habitat; beautiful, seminatural landscapes; watersheds; livestock products; and recreation. Through the decisions they make, ranchers shape the characteristics and rate of ecosystem service production from rangelands (Huntsinger and Oviedo 2014). The construction of stock ponds can benefit tiger salamanders and red-legged frogs, and control of weeds and managed grazing can offer ecosystem benefits. Irrigation canals and pasture runoff create favorable habitat conditions for some species, including the California black rail (Laterallus jamaicensis coturniculus; Richmond et al. 2010, 2012). Ranchers can offer ecosystem services like bird-watching, hunting, and other recreational opportunities on the market. In California, production of organic and grass-fed livestock products from rangelands is increasing, providing a way for ranchers to increase the market value of range products. Considerable interest exists among ranchers in finding ways to diversify income streams by marketing additional ecosystem services or participating in payment-for-services programs (Cheatum et al. 2011). As studies have evinced the capacity of grazing to improve certain kinds of habitat, interest in using grazing for conservation benefits has grown. However, even when convinced that livestock grazing is essential for achieving conservation goals, managers and planners often fail to consider the ranches or people that produce livestock. As in many parts of the world, government and nongovernmental organizations often hire “professional land managers” or “environmental consultants” who do not know the rancher perspective, respect rancher knowledge, or understand the imperatives of the pastoralist operation. They might ask ranchers to graze for a month here or a month there at limited times, making an economic enterprise infeasible. The following section explores some of the challenges and creative solutions being implemented as range management moves into the future.

Future of Range Management in California California ranchers face serious challenges including those associated with inheritance, increasing property taxes, worsening industry economics, loss of infrastructure, increasing conflicts with urban neighbors, fragmentation, development, and agricultural intensification of grazing lands. Livestock prices do not necessarily reflect conditions affecting the range industry, as ranchers are at the bottom of a multilayered industry where livestock prices are determined by largescale feedlot, processing, and retail enterprises. Drought and the prospect of its increased likelihood also confront rangeland producers. Animals permitted on public rangelands have

generally declined, as have livestock numbers overall (see Figure 37.2). The likelihood that a ranch will persist is higher if the rancher is able to conduct a profitable business, and this has a direct impact on the conservation of rangeland landscapes. Average returns of only 2% to 3% from livestock production on western rangelands (Workman 1986, Torell et al. 2001) mean that ranchers often must depend on returns from land appreciation to recoup their long-term investment and to obtain capital, which means selling the land to development interests. In some areas conversion to more intensive, highvalue agriculture, such as grains, viticulture or horticulture, is an option. With agricultural intensification, carbon emissions increase and wildlife habitat is lost. Further challenges for ranching in California include a critical shortage of packing and processing facilities within the state, federal food safety regulations that forbid on-farm slaughter for sale, and competition for grazing lands from California’s large dairy industry, where young females may be raised to milking age on rangelands. Finally, variability in forage production remains an annual challenge and adds an element of risk to livestock operations (Brownsey et al. 2013). While years of abundance interspersed with years of belowaverage rainfall are expected and incorporated into the management of California ranches, long-term drought can create huge costs through feed purchases made in an attempt to maintain the herd. The multi-year drought beginning in 2012, for example, caused liquidation of many herds as ranchers found themselves without forage or affordable feed. Interest in acquiring and feeding agricultural by-products like waste squash, corn stalks, tomatoes, and crop aftermath skyrocketed. Cow-calf operators maintain a base cow herd to produce calves, which limits their ability to reduce numbers in drought. Standard practice is to reduce brood herds by selling older, less productive animals first. Brood herds are often the result of decades of breeding and selection, with cows that do best on a ranch’s unique configuration of forage resources persisting in the herd. The genetic resources lost from these “adapted” herds take a long time to recover. For those in the business of grazing yearlings for a few months annually, it becomes a matter of weighing how long the animals can be kept in a drought year versus opportunities to maximize price. If range has been contracted for, the operator might still have to pay grazing rents even when the animals are gone. Depending on concurrent livestock prices, operators might have to sell at or below cost. Beyond the loss of income from having fewer or no cows, ranchers need capital to purchase replacement animals when drought ends. The amount of land a California rancher leases to complete the annual forage requirements for a herd of cows is substantial. Leasing of public and private land is common, with public land usually about a third to half of the rangeland portfolio (Liffmann et al. 2000, Sulak and Huntsinger 2007; Huntsinger, Johnson et al. 2010; Lubell et al. 2013). Studies have shown that without such leases, a large proportion of operators believe they cannot sustain their ranches (Sulak and Huntsinger 2007). In an interview, one central Sierra permittee put his family’s dilemma eloquently into words: “Public lease versus private lease? Where is the opportunity? How will we pass on this ranching operation to the next generation? These questions will be resolved over the next ten years—​w ithout public lands as an option the answers may be harder to come by for the next generation.” Agency lessors,

on the other hand, find themselves in the position of selecting the surviving generation of ranchers by way of their leasing decisions. The California Rangeland Conservation Coalition has developed a “rangeland resolution” with more than one hundred public agency, ranching groups, and environmental organization signatories that commits them to working for rangeland conservation in concert with livestock production in the state (CRCC 2014). These and other similar efforts in the western United States argue for sustaining “working landscapes” that combine livestock production and ecosystem service production (Huntsinger and Sayre 2007, Huntsinger et al. 2014). There are initiatives at multiple scales, from mitigation markets and planning that support landscape conservation through maintaining ranching as a widespread land use, to environmental payment-for-services programs and outreach efforts that encourage specific management practices and help sustain extensive rural land uses.

Ecosystem Service Initiatives Public acquisition is one way to protect grasslands from development, but it is costly and controversial (Merenlender et al. 2004). Regulations protecting endangered species, water quality, air quality and wildlife in general can also protect ecosystem services but can have negative impacts on ranchers and therefore on extensive rangelands. They may increase costs and imply social disapproval of ranching. While ranchers overwhelmingly responded that they valued the natural beauty of California oak woodlands, the majority in a survey of two California counties responded that “being over-regulated” was a good reason to quit ranching altogether (Liffmann et al. 2000). Regulatory efforts can also have unfortunate unintended consequences. Researchers in North Carolina documented landowner destruction of forest habitat to prevent occupation by a rare species in an effort to avoid the restrictions of the Endangered Species Act (Lueck and Michael 2003). Overall, the Endangered Species Act and regulations like it, while a valuable component of environmental protection, do little to increase habitat extent or quality from private lands (Bean and Wilcove 1997) or to stimulate ecosystem service flows in general. As an alternative, in recent years efforts have been made by the conservation community to offer incentives to private landowners for maintaining the natural characteristics of their land as a way to conserve open space and wildlife habitat (Barry and Huntsinger 2002, Huntsinger and Hopkinson 1996). The rancher provides on-site management for these “working landscapes,” and the land remains agriculturally productive. Efforts to define and measure range ecosystem services are part of this effort. Ranchers have considerable interest in payment for ecosystem services, though they may not recognize it by name (Cheatum et al. 2011). Interest has been expressed in habitat improvement, carbon sequestration, and other possibilities that can complement livestock grazing. Fee hunting, in that it encourages landowners to manage habitat for wildlife, is an existing market for ecosystem services. Examples of payment for ecosystem services programs available for privately owned range include the federal Environmental Quality Incentives Program (EQIP) and Conservation Stewardship Program (CSP). These programs, administrated by the U.S. Department of Agriculture’s Natural R ange Ecosystems   855

Resources Conservation Service (NRCS), offer cost-sharing and payments for certain conservation projects and practices, including managing ponds and water developments to conserve aquatic species, improving wildlife habitat, and protecting water quality. These federal benefits are founded on the idea that the benefit received by the public from the conservation activities of farmers and ranchers makes public investment in these activities worthwhile. In 2012, EQIP paid $117 million, CSP paid $8 million, the Grassland Reserve Program (incorporated into the Agricultural Conservation Easement Program [ACEP] in 2014) paid $0.3 million, and the Wildlife Habitat Incentive Program, now folded into EQIP, paid $0.6 million for conservation projects on California farms and ranches. In addition, the NRCS spent $42.6 million on technical assistance in California in 2012 (USDA 2013). Markets, payments, and cost-shares for ecosystem services help support enterprises financially. Payment for ecosystem services may also transmit a sense of social approval to the landowner, feeding back to greater interest in ecosystem service production. However, equally important is building partnerships with ranchers and landowners based on a mutual interest in maintaining a healthy, pleasant environment. In fact, ranchers and rangeland landowners in numerous studies emerge as strongly motivated by environmental factors; building on that affection for the land and appreciation for responsible stewardship is perhaps one of the most effective ways of conserving rangelands (Smith and Martin 1972; Huntsinger, Johnson et al. 2010). Conservation easements are also a market for ecosystem services, with willing purchasers and sellers. They are now the most widely used, private-sector land conservation method in the United States (Gustanski and Squires 2000). In exchange for tax benefits or outright payment, a landowner voluntarily agrees to a permanent deed restriction on the property title that prohibits development. This right is then held by a third party, sometimes a public agency, but often a nongovernmental organization known as a land trust. Although far from perfect as a conservation strategy (Merenlender et al. 2004, Reiner and Craig 2011), easements allow ranchers to continue providing ecosystem services from the property, while extracting some of the capital value of the land by donating or selling the right to develop (Sulak et al. 2004). This market is subsidized through tax benefits and the U.S. Farm Bill, with the NRCS, for example, allocating $4 million in California for purchase of easements on farms and ranches in 2012 through the Farm and Ranch Lands Protection Program incorporated into the ACEP in 2014 (USDA 2013). A 2005 survey of oak woodland landowners found that approximately 6% of properties had a conservation easement on them (Huntsinger, Johnson et al. 2010). Ferranto and colleagues found in 2011 that 6% of California forest and rangeland owners in ten representative California counties had a conservation easement in place in 2008 (Ferranto et al. 2011). Mitigation easements are similar but are purchased using the funds of property developers to preserve specific types of habitat lost as a result of development. A landowner might have one part of a property designated as a “mitigation easement” for a particular threatened or endangered species, for example. To meet the requirements of the law, easements must be regularly monitored to make sure that the landowner is complying with the terms of the easement. A challenge for conservationists is how to keep the cost of monitoring down. Costs are partly related to how complex the easement requirements are. In some cases, simple site visits or aerial monitoring to 856  Managed Systems

check for construction have been effective. Many organizations now emphasize that they use the monitoring process to build a partnership with the landowner.

Niche Markets for Ecosystem Services Advertising or certifying that agricultural products also offer ecosystem service benefits can raise their value on the market. While the main motivation for purchasers of meat, milk, and leather probably lies in the quality and characteristics of the product itself, heritage values, belief in “sustainable” uses of land, and the appeal of woodland and local landscapes undoubtedly also have a positive impact on many consumers, and interest in various product designations is on the increase. Some producers market to the public by stating that they manage for ecosystem services and “sustainability,” incorporating this into the price. Nongovernmental certification programs play a growing role in informing consumers of the ecosystem services associated with buying various products or brands. California livestock operators have been fairly innovative in developing alternative livestock production systems, especially as alternative production may become an increasingly important part of the industry. Producers are known to market products over the Internet, shipping frozen products or meeting up with purchasers in designated places in urban areas to complete transactions. Farmers’ markets have also become an important outlet for local, organic, and grass-fed products. Finally, grazing as an ecosystem service has also found a market. Grazing for control of fire and invasive weeds can be lucrative. Companies have sprung up offering to provide goats specifically for vegetation management, and they charge as much as $1,300 per hectare for this service. Cattle are also used for fire hazard management, and in fact this is one rationale for grazing on some public lands (Byrd et al. 2009).

Climate Change The implications of climate change for California range are likely to be profound (see Chapter 14, “Climate Change Impacts”). Changes in temperature, precipitation, and carbon dioxide will likely alter forage production and timing, plant species composition and phenology, non-native plant invasions, frequency of drought and wildfire, livestock performance, and many other range attributes (Polley et al. 2013). A good deal of uncertainty still surrounds the magnitude and even direction of change in range attributes at state and local scales, but it appears highly likely that California will experience increased ecosystem variability and more frequent extreme weather events (Polley et al. 2013, ChaplinKramer and George 2013). Although California range managers have long experience working in variable and uncertain ecosystems, the potential for adverse consequences presents a significant challenge for everyone interested in California’s range landscapes. For example, Shaw et al. (2011) projected the impact of several climate change scenarios on forage production in California. They found that under most scenarios, statewide forage production would decline significantly by the end of the century, with a corresponding reduction in profits for California ranchers. Another model that incorporated the effects of temperature as well as precipitation changes on forage pro-

duction and that focused on the San Francisco Bay Area suggested climate change effects on Bay Area forage production might be less grim (Chaplin-Kramer and George 2013). This model projected that in most years, annual forage production could actually increase, although this news was tempered by the likelihood that the growing season could be shorter and that the frequency of extremely dry years would increase (both changes necessitating the potentially costly provision of additional food for livestock). These findings suggest that climate change could bring range managers some opportunities, such as the increase in Bay Area forage, as well as anticipated difficulties. Whatever changes are wrought by climate change, those interested in the maintenance of California’s range ecosystems and the services they provide must start planning and implementing effective adaptation and transformation strategies. Alternative livestock breeds/species and production methods, diversified businesses, innovative pest management, geographic relocation, alternative ecosystem services, and new policies, incentives, and social networks have been proposed as potential strategies (Joyce et al. 2013). Technical outreach to ranchers to encourage the development of drought plans and strategies is important, as is exploration of alternative feed sources such as agricultural by-products. The creation of grass banks—​areas set aside for grazing only during drought—​could also increase flexibility in response to growing uncertainty. Because temporal and spatial changes in California range ecosystems interact at the scale of the range site (Bartolome et al. 2009), it is very difficult to predict how projected climate change will unfold. Practices will undoubtedly need to adapt, and they will need to be developed within a realistic adaptive management framework.

Summary Range forms a diverse class of managed ecosystems covering about a third of California, primarily in natural and seminatural grasslands, savannas, and shrublands. Characteristic of range ecosystems are aridity, nonarable soils, and the important ecological and economic roles of grazing and browsing animals. After more than two hundred years, grazing remains California’s most extensive land use, with goals and management strongly affected by patterns of land ownership. After peaking in 1970, beef cattle numbers have slowly declined along with sheep numbers, leading to social and economic stresses within the producer community. Grazing management differs considerably among the Mediterranean, desert, and montane regions of the state. Models for the principal functional components of range ecosystems (primary producers, primary consumers, secondary consumers, and detritivores) provide a foundation for understanding the basic system and how it can best be managed and sustained. Forage growth patterns drive seasonal livestock production practices in different range areas. Cattle are the most numerous kind of range livestock at present, producing calves in fall in the Mediterranean climate zone and in spring in the intermountain zone. Warm desert grazing is less linked to season. Forage production is heavily influenced by local climate and soils. Models predicting range community responses to grazing currently emphasize nonequilibrium dynamics and integrated management goals as opposed to older, equilibrium-type models focusing on livestock production or grazing impacts. Livestock man-

agement goals today include maintaining biodiversity, controlling invasive plants, managing fuels, and protecting soil, water, and air quality. Management of public range must now consider impacts on private range, as forage resources have become limited statewide. Partnerships with ranchers and rangeland landowners are essential to conserving the extensive landscapes cherished in California.

Acknowledgments Thanks to Rebecca C. Wenk (1979–​2 011) for the livestock and wildflowers photograph, one of many beautiful images she created while working with the University of California–​ Berkeley Range Group.

Recommended Reading Briske, D. D., editor. 2011. Conservation benefits of rangeland practices: Assessment, recommendations, and knowledge gaps. U.S. Department of Agriculture Natural Resources Conservation Service. Bush, L. 2006. Grazing handbook: A guide for resource managers in coastal California. Sotoyome Resource Conservation District, Santa Rosa, California. California Rangeland Conservation Coalition. . Campos, P., L. Huntsinger, J. L. Oviedo, P. F. Starrs, M. Diaz, R. B. Standiford, and G. Montero, editors. 2013. Mediterranean oak woodland working landscapes: Dehesas of Spain and ranchlands of California. Landscape Series 16. Springer Science+Business Media, Dordrecht, Netherlands. Holechek, J. L., R. D. Pieper, and C. H. Herbel. 2011. Range management: Principles and practices. Sixth edition. Prentice Hall, Upper Saddle River, New Jersey. Stromberg, M. R., J. D. Corbin, and C. M. D’Antonio, editors. 2007. California grasslands: Ecology and management. University of California Press, Berkeley, California. University of California Division of Agriculture and Natural Resources. 2014. Annual rangeland handbook. .

Glossary Clementsian equilibrium ecology  Theory developed by Frederic Clements (1874–​1945) and followers describing the predictable development of vegetation over time, based on equilibrium concepts. Climax community  The cornerstone of Clementsian equilibrium ecology. The stable vegetation endpoint in a successional series as determined by climate. Compensatory plant growth  The increase in net primary productivity as a result of herbivory. Ecological site  An assemblage of basic land units, defined by topographic, soil, and climatic characteristics, that share potential vegetation and react similarly to management. Forage  Vegetation available for consumption by grazing or browsing animals. Naturalized species  A well-established, common, and widespread non-native species that survives and reproduces in the wild without human assistance. Nonequilibrium ecosystem  An ecosystem that displays biotic decoupling, nonlinear succession, and multiple stable states. In contrast, in an equilibrium ecosystem, succession is controlled by biotic interactions, passes from stage to stage in a linear fashion, and has a predictable endpoint. R ange Ecosystems   857

Phenology  The stages of seasonal plant growth, development, and reproduction. Range  Refers to land grazed by livestock Rangeland  According to the Society for Range Management (1998), this refers to land on which the indigenous vegetation (climax or natural potential) is predominantly grasses, grasslike plants, forbs, or shrubs, and is managed as a natural ecosystem. If plants are introduced, they are managed similarly. Rangelands include natural grasslands, savannas, shrublands, many deserts, tundras, alpine communities, marshes, and meadows. Residual dry matter  Refers to old, aboveground herbaceous plant material left standing or on the ground at the beginning of a new growing season. Savanna  This is grassland with scattered trees (usually less than 30% tree cover). Specialized grazing systems  Grazing management employing recurring periods of grazing and removal of grazing in at least two pastures or management units. Transhumance  The practice of seasonally moving animals to take advantage of forage, often along an elevational gradient.

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864  Managed Systems

climate change, and the terrestrial carbon sink. Global Change Biology 6:817–​833. Wiens, J. A. 1989. Spatial scaling in ecology. Functional Ecology 3:385–​397. ———. 1984. On understanding a non-equilibrium world: Myth and reality in community patterns and processes. Pages 439–​458 in D. R. Strong, D. Simberloff, L. G. Abele, and A. B. Thistle, editors. Ecological communities: Conceptual issues and the evidence. Princeton University Press, Princeton, New Jersey. Williams, K., R. J. Hobbs, and S. P. Hamburg. 1987. Invasion of an annual grassland in Northern California by Baccharis pilularis ssp. consanguinea. Oecologia 72:461–​465. Wolch, J. R., A. Gullo, and U. Lassiter. 1997. Changing attitudes toward California’s cougars. Society and Animals 5:95–​116. Workman, J. P. 1986. Range economics. Macmillan, New York, New York. Yoakum, J. D., and B. W. O’Gara. 2000. Pronghorn. Pages 559–​577 in S. Demarais and Paul R. Krausman. Ecology and management of large mammals in North America. Prentice Hall, Upper Saddle River, New Jersey. Zavaleta, E. S., K. Hulvey, and B. Fulfrost. 2007. Regional patterns of recruitment success and failure in two endemic California oaks. Diversity and Distributions 13:735–​745.

THIRT Y-EIGHT

Agriculture ALE X MCCALL A and R ICHARD HOWIT T

Introduction The task of this chapter is to provide a comprehensive overview of California agriculture as a managed ecosystem. It is a highly productive, complex, ever-changing system that is fragile and, in the judgment of many, unsustainable. Whole books have been written to cover the same terrain (Starrs and Goin 2010), as it comprises the largest and most diverse agriculture in the U.S. We begin with a description of the broad dimensions of California agriculture in terms of land and water use, value and diversity of production, scale and distribution of production, and the relative importance of crop and livestock production. We focus on irrigated and rain-fed crop production and confined livestock production operations. The second section of this chapter provides a history of the rather short life of California agriculture. In less than 250 years, the state’s agricultural system has grown to be a modern, large-scale, technologically intense commercial enterprise. Commercial agriculture is slightly over 150 years old, but most recently it has been transformed from an extensive, dryland, livestock and cereals economy to a diverse, intensively farmed irrigated agriculture. The section closes by describing how Cal-

ifornia agriculture is distributed on one-tenth of California’s land mass. Most of it is in river valleys, large (the Great Central Valley) and small (the Coachella Valley), running 1,125 kilometers southeast to northwest from Mexico to the Oregon border. The third section deals with the functioning of California’s agricultural system in its current form as a managed ecosystem, looking at components such as fertility, salinity, urbanization, groundwater depletion, water quality, and expansion and contraction of the irrigated area. Section four explores how this managed ecosystem has impacted natural ecosystems by looking at three major cases: loss of wetlands to agriculture, restrictions on salmon habitat, and changes in the food base for birds in the Pacific flyway. Section five looks at synergies and trade-offs with other parts of California’s ecosystems and points out that not all interactions are negative. We discuss two emerging, positive interactions: flooded rice fields and expanded bird food availability in the Sacramento Valley, and agriculture as fish habitat in the flood plains of the Yolo Bypass. The chapter closes with our thoughts about the future of the California agricultural ecosystem. 865

Overview of California Agriculture General Parameters L AND

The majority of California’s 40 million hectares is mountains and desert. About 17.5 million hectares are classified as having some possible use for agriculture: 6.5 million hectares potentially available for grazing and 11 million hectares for cropland. Actual land in farms in 2007 (according to the 2014 U.S. Census of Agriculture) was identified as 10.25 million hectares, down from a high of over 14 million hectares in the 1950s. In 2007 slightly less than 4 million hectares were identified as cropland, of which 3.25 million hectares were irrigated. Because of urban encroachment, conversion to conservation uses, and land degradation, both of these numbers have been declining. Cropland is down from its peak in 1950 of 5.5 million hectares and irrigated area down from its peak in 1997 of 3.6 million hectares. Crop and irrigated land is located primarily in river valleys, with most of it in the Great Central Valley (Figure 38.1).

tor expansion of groundwater use, with the application of new centrifugal pump technologies in the early 1900s. This expansion occurred primarily in 1900–​1930 and in the 1940s. Irrigated acreage in 1940 was just over 1.6 million hectares and rose to exceed 2.4 million hectares by 1950. Large-scale, interbasin transfers of water, funded by the public sector, rapidly expanded as the Central Valley Project (CVP) came fully on board in the 1950s, 1960s, and early 1970s when 0.8 million additional hectares were added. The completion of the Tehama-Colusa Canal and the California Water Project (CWP) in the 1970s and early 1980s brought irrigated area to its peak of almost 3.6 million hectares (Figure 38.2). California’s rise to become the top agricultural producing state after World War II paralleled public sector irrigation development, including the rapid expansion interbasin transfers of surface water. California agriculture consumes the vast majority of human water used in California, though agriculture’s share has steadily eroded from about 90% in the 1960s to less than 80% in recent years. This declining share, coupled with a steady decline in total water available (from 53 billion cubic meters in 1995 to 44 billion cubic meters in 2005), means agricultural use has fallen from an annual peak of over 43 billion cubic meters in 1980 to 33 billion cubic meters in 2005.

CLIMATE AND AGROECOLOGY

California is blessed with a rich and diverse agroecological endowment. Its climatic conditions range from cool temperate to subtropical hot, and its elevations range from below sea level to high plateaus. These features give it the potential to produce a diverse menu of products (claimed to exceed four hundred in 2013 by the California Department of Food and Agriculture). The critical constraint on fulfilling that potential is water availability—​in sufficient quantities, at the right places, at the right times. Average annual precipitation varies from 5 centimeters in the Imperial Valley to 173 centimeters at Blue Canyon in the Sierras near Lake Tahoe. Almost all falls in November through March. Sixty percent of the water used in California comes from Sierra snowmelt.

WATER AND IRRIGATION

The development of California agriculture therefore is inextricably linked to water control, capture, management, and transport. Early water challenges included controlling flooding of the Sacramento River and managing hydraulic mining debris in the Sacramento Valley 1860–​1920 (see Kelley 1989), control of the rivers feeding Tulare Lake, and ultimately converting the Tulare lake bed into farm land (Arax and Wartzman 2005). Reclamation of wetlands also played a major role. Of the wetlands that existed in California in 1900, only 4.9% remain today (see Chapter 31, “Wetlands”). For example, a total of 179,000 hectares were drained in the Sacramento River Delta alone between the 1850s and the 1930s. California’s irrigation expansion occurred in three waves: small, within-basin development; then groundwater extraction; then large-scale, interbasin surface water transfers. Early expansion was primarily small-scale, private sector, withinbasin development on the western slopes of the Sierras (1870s through the 1930s). The second wave was again private secPhoto on previous page: Intensive cropping around the Salinas River, Monterey County, California. Photo: Aerial Archives / Alamy. 866  Managed Systems

AGRICULTUR AL PRODUCTION

The state of California is by far the largest producer of agricultural products in the U.S., with a farm gate value of $43.5 billion in 2011. This exceeded the next two largest producing states, Iowa ($29.9 billion) and Texas ($22.7 billion), by significant margins. California has been the leading state since the late 1940s, with its share of U.S. production rising from 8% in 1950 to a high of 13.3% in 2005. High grain and oilseed prices since 2007, which dominate farm value of production in the Midwest, lowered California’s share to 11.6% in 2011. Crop production increasingly dominates California agriculture, with the share of output coming from the crop sector increasing from 61% in 1950 to 75% in the first decade of the twenty-first century. Running counter to this shift towards crop production, California’s dairy industry has grown to become the largest commodity sector in the state’s agriculture, accounting for more than 17% of the value of California farm output, more than twice the value of each of the next two most important commodities: almonds and grapes. In recent years, California has produced about 15% of the value of U.S. crop production and 7.4% of U.S. livestock production.

FARM SIZE

California agriculture is characterized by a highly skewed distribution of farm size as measured by farm sales. The 2007 Census of Agriculture counted just over 81,000 farms in California (for census purposes a farm is an enterprise that sells over $1,000 yr-1 of product); 69% of those farms sell less than $50,000 yr-1 of product and account for only 1.5% of the total value of California output, while 8,580 farms (10.6%) who sell over $500,000 yr-1 account for 90.1% of the value of California output ($33.8 billion on 2007). This is even more skewed than overall U.S. figures, where farms selling over $500,000 make up 5% of U.S. farms and sell 73% of U.S. output. Finally, California agriculture has become increasingly foreign-trade

Important farmland in California, 2010 Irrigated farmland Dryland farming and grazing land Other land Urban and built-up land Water Local, state, and federal owned land Out of survey area

FIGURE 3 8.1 Cropland and irrigated agricultural areas in California. Map: California Department of Conservation, Farmland Mapping and Monitoring Program, 1984–​2 013.

oriented, exporting now nearly 40% of the value of farm output (whereas the U.S. value is approximately 33%). For example, more than two-thirds of California almond, walnut, and pistachio production is exported.

The Scope of California’s Agriculture

Total cropland

Irrigated land

5

Million ha

This chapter focuses on irrigated and dry land crops as well as confined animal operations—​principally dairy, beef, and poultry production. The area utilized as rain-fed range pasture for cattle and sheep is dealt with in Chapter 37, “Range Ecosystems.” The vast majority of the value of California crop production is derived from irrigated acreage. Irrigated area varies by year depending on water availability, but total area has been declining in recent years. It peaked at 8.9 million hectares in 1997 and declined to 3.6 million hectares in 2007. Precisely separating the value of irrigated production from the value of total crop production is difficult in California because county reports are based on volumes marketed, not on the source of water used in production. Some field crop production of barley, wheat, and hay, and perhaps still a few acres of almonds and grapes, are produced under rain-fed

6

4 3 2 1 0

69 879 889 899 909 919 929 939 949 959 969 978 987 997 1 1 1 1 1 1 1 1 1 1 1 1 1

18

Year FIGURE 3 8.2 Total cropland and irrigated land in farms, California, 1869–​2 002. Sources: Data from Olmstead and Rhode 1997 (1869–​ 1919), U.S. Department of Commerce Bureau of Census 1967 (1929–​ 1959), and U.S. Census of Agriculture (1969–​2 007).

Agriculture  867

TA B L E 3 8 .1 Proportion of total value of agricultural output in California, 2011

Share included

Included value ($1,000)

4,927,714

93%

4,582,774

15,322,511

97%

14,862,835

Nursery, greenhouse, and floriculture

3,687,630

100%

3,687,630

Vegetable crops

7,241,252

100%

7,241,252

12,357,994

92%

11,336,483

Milk and cream

7,680,566

95%

7,296,357

Chickens, eggs, and turkeys

1,381,092

100%

1,381,092

39,196

100%

39,196

Sheep and lambs

287,463

50%

143,731

Cattle and calves

2,825,125

80%

2,260,100

432,015

50%

216,007

43,544,001

95%

41,710,974

Category Field crops Fruit and nut crops

Livestock, poultry, and products

Hogs

Other Totals

2011 Total value ($1,000)

Source: California Department of Food & Agriculture 2013, and authors' calculations.

conditions; and some irrigated pasture is used in the dairy industry. However, it is highly unlikely that rain-fed agriculture produces even 5–​7% of the value of plant production (less than $2 billion in 2011). We can only conclude that the vast majority of the value of field crop, fruit and nut, and vegetable production is produced under irrigation of some form. In 2011 this totaled $27.5 billion (63% of the value of California agriculture sales). The nursery sector (value almost $3.7 billion in 2011), which is not counted as a crop per se, also uses irrigation. Beyond plant products produced, the livestock sector, especially the dairy, poultry, and beef feed lot subsectors, also depend highly on the irrigated sector for feed stocks. It is also difficult to sort out what share of the value of livestock, poultry, and products comes from confined animal operations. Virtually all of California’s dairy operations are now large-scale, confinement operations. There are still a few nonconfinement operations on the North Coast, but we estimate the share of confinement dairy production at 95% of the value of milk and cream. Essentially all of poultry production is confinement, as is what little hog production there is. This leaves the problem of segregating the value of rain-fed range cow-calf operations from the aggregate 2011 value of $2.825 billion reported for cattle and calves. NASS statistics do segregate out sales from feed lots of 630,000 head with a value of $780 million. But these numbers exclude cows culled from dairy herds, which likely number as many as 500,000 head per year out of a dairy herd of 1.8 million milking cows. The sum of those two would be 1.13 million head out of 1.7 million slaughtered in 2011, yielding a minimum of 67% of value that comes from confinement operations. The 2007 census data also show that 80% of cattle and calf sales come from operations of 5,000 head or more. As these are highly 868  Managed Systems

likely to be confinement operations, we can estimate that up to 80% of the cattle and calves value comes from confinement operations (Table 38.1). The vast majority (95%) of the real value of California agriculture is thus included in our analysis.

A Brief History of California Agriculture California, until well into the eighteenth century, was one of the few remaining major “hunter-gatherer” societies left in the world (Adams 1946, Diamond 1999, and Smith 1998). The origins of sedentary California agriculture came with the development of Spanish missions over the period 1769–​1823. Over its brief history of 250 years, the character of California agriculture has been in a perpetual state of transition and adjustment: from early mission attempts to raise livestock, grow grains, and develop horticulture; to the era of ruminants (i.e., cattle and sheep); to the development of largescale, extensive wheat and barley production; to the beginnings of intensive fruit, nut, and vegetable agriculture based on ditch irrigation and groundwater; to pioneering largescale beef feedlot and dairy production; to the intensified and expanded production of an increasingly diverse portfolio of crops resulting from massive public irrigation schemes; to today’s highly sophisticated, technologically advanced, management-intensive agricultural industry, which is embedded in a rich, urban state of thirty-eight million people. It is a history of perpetual, profound, and often painful change. California agriculture has always battled economic adversity. While blessing California with good weather and fertile soils, nature did not provide adequate rainfall in the right places or times. Substantial investments are therefore needed

to bring water to the soil to grow crops. The upside is that irrigation potentially allows watering of crops at the precise time of need and in the correct amounts, greatly increasing the range of production options. Thus water management is a critical additional dimension of complexity for California agriculture. California is a long distance from everywhere; therefore, importing and exporting have always been expensive in terms of both money and time. Finally, California, because of its mild, Mediterranean climate, has different and more complex problems with pests and diseases than does the rest of mainland agriculture. Despite continuous change, Johnston and McCalla (2004) argue that at least seven constants have driven California agriculture. 1. California agriculture has always been “demand driven.” It was never subsistence, family-farm agriculture like that which characterized much of early U.S. agriculture. 2. California agriculture is resource-dependent (land and water). Its history includes aggressive development of new land and water resources along with cases of soil and groundwater exploitation. 3. California agriculture has been shaped by the absence of water in the right place. It has always been in search of more water and has been an aggressive participant in water disputes with both internal and external competing interests. 4. California agriculture has always depended on a large supply of agricultural labor for planting, cultivating, and harvesting its abundant produce from both relatively large-scale operations and specialty-crop farms. The source of a stable supply of field labor has varied over time from countries of Asia and the Americas. 5. California agriculture has grown rapidly and almost continuously, although it has been periodically buffeted by natural catastrophes (e.g., floods, droughts) and adverse economic shocks (e.g., the Great Depression, various recessions). 6. California agriculture, at least since the Gold Rush, has required very high levels of technical and economic management skills. It has always been dominated by large-scale operations that have grown in complexity and sophistication. 7. It has always been on the technological frontier in developing, modifying, or stealing new technologies, such as large-scale mechanical technology, irrigation equipment, plant varieties, pest control, food processing, and wine making. If we are to understand where California agriculture might go in the twenty-first century, we must understand the forces that have shaped California agriculture to date. Therefore we trace that evolution in more detail in terms of eight episodes grouped into three clusters—​pre-twentieth century and the first and the second halves of the twentieth century. This review focuses on the impacts of five critical drivers of California agricultural development:

. Changes in the market demand for California agri.

cultural products, driven by population and income growth in and outside of California. The constant search for ways to control, manage, and access more water.

. The seeking of biological, mechanical, and engineer. .

ing technology to better manage the complexities and opportunities of the natural ecosystem. People pressure: competition for resources, especially water and land as rapid economic growth and urbanization forced agriculture out of preferred locations. Public investments in infrastructure, education, research, and development; and public policy interventions that foster agricultural development such as the Wright Act, marketing boards, cooperatives, and the University of California agricultural sciences.

Pre-twentieth Century SPANISH-ME XICAN PERIOD, 1760–​1 848

The Franciscan Order extended its missionary activities to Alta California in the 1760s. Led by the pioneering efforts of Fathers Portola, Serra, and others, the order developed a string of twenty-one Spanish Missions from San Diego to Sonoma over the period 1769–​1823. The Spanish Church missionary strategy included, in addition to its missions (spiritual), the presidio (military) and the pueblo (commercial) as components of early development. Livestock, field crops, and horticulture were introduced to feed the settlers and to provide economic activity for the converted natives. But Alta California was never much more than self-sufficient, as total acreage of all cultivated field crops in all missions never exceeded 2,000–​4,100 hectares and livestock numbers varied between 285,000 and 400,000 head in the period 1807–​1834 (Adams 1946). Mission agriculture was small, and much of it disappeared in the subsequent Mexican period. Until Mexican independence in 1821, land was vested in the church, and few land grants were given out in the Spanish-California period. Mexican independence was followed by a period of uncertainty as to the role of the church in secular affairs. This was not settled until the secularization of all missions in 1834, which stripped the church of land ownership and established the principle (unrealized) of the division of land between settlers and natives. Prior to 1822, there had been some thirty large “Rancho” land grants, and that number had risen to fifty by 1834. But after secularization (1834) and before the Bear Flag Rebellion (1846), 813 additional land grants were issued, totaling 5 million to 6 million hectares (Jelinek 1982). This was the period of the California Rancho—​large spreads of land acquired by grants to Mexican citizens where cattle ranged largely untended (Jelinek 1982). Periodically cattle were slaughtered on the range to meet an East Coast and international demand for tallow and hides.

GOLD, STATEHOOD, CAT TLE , AND GROW TH, 1848–​1 860S

The discovery of gold in 1848 and the Gold Rush of 1849 shaped the new state of California (which entered the Union in 1850) and fundamentally altered California agriculture. The European population of California was estimated at 7,000 in 1845. In January 1849 it was estimated to be 26,000. By December 1849 it was 92,000, and it multiplied to 255,000 by 1852 and to 380,000 by 1860 (Jelinek 1982). This explosive growth increased population tenfold between early 1849 and 1860. The numerous gold miners, and an even larger numbers of people who came to profit by serving the miners, Agriculture  869

needed to eat, and a strong demand for food (especially meat) emerged. Rancheros reaped first advantage from the population surge: “Hides hides gave way to beef as the price of cattle rose from under $4 a head before the rush to $500 a head at one point in 1849, leveling off at $50–​$150 a head during the 1850’s” (Jelinek 1982, pp. 23–​24). Southern herds were driven up the Central Valley or along the coast and sold to Americans who drove them into northern and Mother Lode towns for processing. Competition soon came from American cattlemen from the Midwest and Texas, who in the 1850s drove herds west to California. By mid-decade, “up to 40,000 head entered annually” (Jelinek 1982). Large numbers of sheep were also driven in from the Southwest. In the peak year of 1856, 200,000 head of sheep entered California. But the rancheros lost out to American entrepreneurs who better understood the nature of the demand for meat and the need for improved herds. By the end of the decade, an American-owned cattle and sheep business had flourished in California. Estimates of the number of cattle vary from 1,234,000 head recorded in the 1860 census (Hart et al. 1946) to 3 million head estimated by Jelinek (1982). Weather, the ever-threatening wild card of California agriculture, dealt a near-death blow to the cattle industry in the first half of the 1860s. In 1861–​1862 a huge flood in the Central Valley created a lake 400–​485 kilometers long and 32–​ 100 kilometers wide that drowned perhaps 200,000 head of cattle (Jelinek 1982, McClurg 2000). Immediately following the flood was a two-year drought in 1863–​1864. Durrenberger (1999) claimed that droughts of the 1860s “resulted in the death of millions of head of cattle.” Thus, even with some recovery in population, there were only 630,000 head of cattle left in California in 1870.

SHEEP, WHE AT, AND E ARLY HORTICULTURE (1860s–​1 8 90s)

The cattle industry was briefly overtaken in the 1860s by the sheep industry as California’s major agricultural enterprise. The first census in 1850 identified 17,514 head of sheep. By 1860 it had climbed to a million head, and the industry peaked in 1876 at 6,406,465 head (Hart et al. 1946). But even before the sheep population peaked, wheat acreage was growing rapidly on extensive ranches (Scheuring 2010). California already had significant wheat production in 1859 and had begun to export wheat. The combined acreage of wheat and barley soared in the 1860s, exceeding 0.5 million hectares in 1867 and peaking at nearly two million hectares in the late 1880s (Olmstead and Rhode 1997). However, as quickly as wheat (and barley) had grown to dominate valley agriculture, it crashed to the point that “by the end [of the first decade] of the 1900s only about 0.5 million acres of wheat were cut, and the state became a net importer of wheat” (Olmstead and Rhode 1997, p 2.). Three causes are often postulated to explain the demise of the wheat industry, though there is some disagreement on the third. The first was soil exhaustion. Yields were declining as largescale wheat growers simply mined natural soil fertility and moved on (Stoll 1998). Second, a severe depression in agricultural prices occurred in the 1880s, and it was acute in wheat. California’s distance from European markets resulted in very low farm prices. Third, development of a small but diversified fruit, nut, and vegetable industry provided an alternative 870  Managed Systems

land use. In the same period (the 1890s) irrigated acreage was increasing rapidly. It is tempting to argue that horticulture replaced, if not displaced, wheat. But some scholars argue that wheat declined and fell on its own (Stoll 1998). It is safe to say that the expansion of horticulture occurred simultaneously with the decline of wheat, thus staving off a severe depression in California agriculture.

CALIFORNIA AGRICULTURE AT THE TURN OF THE CENTURY

California agriculture at the end of the nineteenth century had already experienced several transformations of phenomenal magnitude and speed—​f rom mission agriculture to cattle, to sheep, to wheat (Figure 38.3). Another transformation was under way that would forever shift California agriculture from extensive, dryland agriculture to intensive, irrigated agriculture. The cattle and sheep booms each lasted less than twenty years, and the wheat (and barley) peak had passed before the beginning of the 1900s. The rapid growth of irrigated land started in 1880 and tripled between 1900 and 1930. This is a proxy for the growth of irrigated horticulture and vegetable production. Thus the major outlines of the state’s agriculture were fairly well established by the beginning of the twentieth century (Benedict 1946).

The First Half of the Twentieth Century E XTENSIVE TO INTENSIVE AGRICULTURE (18 90–​1 930)

According to Olmstead and Rhode (1997, p. 5), “The share of intensive crops in the value of total output climbed from less than 4 percent in 1879 to over 20 percent in 1889. By 1909 the intensive share reached nearly one-half, and by 1929, it was almost four-fifths of the total.” The growth in fruit shipments was rapid, increasing fivefold between 1890 and 1910. The most phenomenal growth was in oranges, from two navel orange trees planted in 1873 to 5.5 million orange-bearing trees in 1900. California, aided by the transcontinental railroad and new cooling technology, soon expanded from servicing local needs to shipping products to eastern U.S. markets and abroad. Those who analyzed the phenomenal transformation of California agriculture between 1890 and 1930 postulated a particular set of drivers. Jelinek (1982) argued the process was stimulated by prominent individuals, but success really depended on four critical factors:

. Available agricultural labor from a succession of inter. . .

national sources—​China, Japan, the Philippines, India, Mexico. Irrigation development—​about 0.5 million irrigated hectares in 1890 and almost 2 million by 1930. Improved transportation services—​refrigerated rail shipping, trucking and rural roads, and improved handling, storage, and transportation technology. The development of marketing cooperatives that provided innovation in selling, rapidly increasing production outside of California and to the world.

Rhode (1995) argues that two dominant factors, usually not discussed, were (1) rapid decreases in credit costs (interest

5 Peak cattle acreage 1860 Peak wheat (Jelenek) acreage 1885

4 3 2

ttle

Ca

1 0

1850 1860

1870

1880 1890

d gate age

Sheep

Wheat

1900 1910 1920

2

Hectares harvested

2.5

Peak sheep acreage 1880 Ri sin ac g irr re i

Livestock numbers (millions)

6

1.5 1

0.5 0 1930

Year FIGURE 3 8.3 Selected waves of agricultural development in California, 1850–​1930. Source: Johnston & McCalla (2004).

rates declined significantly around 1890) and (2) horticultural productivity that was substantially improved by “biological learning.” Rhode (1995) also identified four other drivers of importance, three of which are the same as Jelinek’s:

. Mechanization (California was always a leader).

.

. .

Among other innovations, California was home to the first commercial combine harvester, track-laying tractor, orchard sprayers, and mechanical fruit and nut harvesters. Irrigation. Small-ditch irrigation schemes and newly formed irrigation districts, under the Wright Act of 1897, fostered expansion of surface-water irrigated acreage, and the 1890 invention of the centrifugal pump allowed greatly expanded groundwater use in the early 1900s. Labor. California had access to large supplies of quality labor at low cost. Cooperatives. California had innovative forms of collective grower action.

The major point is that the transformation from extensive grain growing and livestock grazing occurred relatively quickly and resulted from a complex interaction of many factors. From 1890 to 1930 the population of California increased fivefold (1 million to 5 million). Incomes rose rapidly from 1910 to 1929, which drove consumer demand toward fruits, vegetables, and livestock products and away from grains and field crops. A world-class agricultural research and extension system was established. Californians continued to import biological technology, to test and modify it, and to apply it quickly. The 1920s, according to Benedict (1946), was a period of relative optimism and rapid development. In the period 1919–​1929, grape acreage expanded 94%; subtropical fruit and nut acreage 82%; vegetable acreage 91%; and temperatezone fruit acreage 63%. In contrast, acreage of cereals, hay, and other field crops fell (Benedict 1946). In part, this transition was responsive to changes in relative agricultural prices in the 1920s. All agricultural prices fell sharply in 1919 at the end of World War I, and grain prices stayed relatively low throughout the 1920s before plunging again in 1930. Prices of fruits, vegetables, nuts, and cotton, however, recovered substantially in the 1920s, fueled, no doubt, by rising incomes and a growing California population. Immigration in the 1920s amounted to 1.25 million people, who came for wellpaying jobs in growing cities.

Thus, by 1930, California seemed on the way to agricultural riches, but ominous events were beginning to cause worry. Much of the expansion of irrigation in the period 1900–​1930 came from groundwater sources; in 1902 less than 10% of irrigation water came from groundwater sources. The fastest expansion in groundwater exploitation occurred in 1910–​1930, driven, in part, by widespread adoption of the centrifugal pump. The number of pumping units increased from approximately 10,000 in 1910 to almost 50,000 in 1930. Groundwater use again expanded in the 1940s, rising to 75,000 units in 1950 (Olmstead and Rhode 1997). Groundwater had been perceived as an unlimited resource, but by 1930 problems with falling water tables, subsidence, and salinization were steadily approaching levels that Riesner (1993) called an ecological time bomb.

DEPRE SSION AND WAR (1930–​1 94 9)

The threat of water shortages was only one of the pending shocks facing California agriculture. The 1920s was a period of rapid expansion in many perennial crops, as perennialcrop prices had fared better than grain prices. Therefore the crash into the Depression was even more precipitous and shocking. Failing prices, exacerbated by significant droughts in 1929, 1931, 1933, and 1934, led to sharp contractions of farm income. Irrigated area dropped by 0.5 million hectares between 1929 and 1935. The index of farmland values, which had been at 160 in 1930, plunged to 109 in 1933 (Benedict 1946). Unemployment rose rapidly, and job-induced in-migration, which virtually stopped in the early 1930s, was soon replaced by an influx of poor farmers displaced by the Dust Bowl and the Depression. These new migrants, poor and unemployed, settled mainly in rural California, adding to an already volatile and sometimes violent labor situation. Contractions in demand hammered farm prices, drought reduced farm production (and income), surplus labor put downward pressure on wages, and poverty rates soared among both farmers and farm workers. The crisis of the 1930s fundamentally altered the policy environment in which California agriculture operated. Prior to 1930, California farmers opposed federal participation in agricultural affairs and were little affected by it, but the Depression changed that. Federal intervention came on several fronts, including new forms of credit under the Farm Credit Administration. The Agricultural Adjustment Administration (AAA) was a comprehensive program designed to reduce production and provide floor prices. Relatively generous support prices, coupled with California’s efficiency in producing rice, cotton, and milk, no doubt contributed to rapid expansion of production of these commodities in the 1940s. The 1930s also saw major efforts by California agriculture to enlist state assistance in constructing a major water scheme to capture and transport northern Sierra water south to Central Valley agriculture. The original Central Valley Water Project was proposed as a state operation to be financed by a voluntary bond sale. It quickly received legislative approval, but the deepening Depression prevented the bond sale from being initiated. Attention turned to the federal government, where the idea of spending to help agriculture while creating public works employment appealed to New Dealers in the early 1930s. In 1935 the Central Valley Project (CVP) became a federal Bureau of Reclamation project, and after 1937 a massive Agriculture  871

dam and conveyance system began to be constructed. The major impact on California agriculture would not, however, occur until after World War II. Prior to the 1930s, irrigation development in California was almost exclusively financed privately. Less than 1% of irrigated acreage in California had been developed through federal action (Benedict 1946). The takeover of the CVP by the Bureau of Reclamation in 1935; subsequent bureau projects such Monticello, New Melones, and San Luis; and the state’s subsequent development of the California State Water Project (SWP) in the 1960s meant that a very large share of all subsequent surface-water development was publicly financed. The 1940s saw a rapid return to prosperity and growth. This recovery was mainly a product of developments in the California and U.S. economies that were driven by the war effort. Durrenberger (1999) argued that World War II transformed California from a rural, natural resource–​based economy to a leading industrial and military state in just five years. The population in California almost doubled, from 5.6 million in 1930 to 10.6 million in 1950—​w ith nearly 4 million of the increase occurring in the 1940s. In-migration resumed and, at its peak in the early 1940s, amounted to about 0.5 million people per year. California led all states in receipt of federal wartime expenditures. According to Durrenberger (1999, p. 101), “Over 90% of federal expenditure to promulgate the war in the Southwest Pacific was allocated to California.” Unlike after World War I, no sharp fall in agricultural prices and incomes followed World War II. California agriculture continued to grow and diversify due to a combination of federal policy and a rapidly growing economy. The value of farm output in California grew 24% in 1945–​1950. Expansion and its labor requirements made labor the dominant issue in California agriculture in the postwar 1940s and 1950s. The wartime boom had siphoned excess labor out of agriculture. California agriculture, facing rising wage rates, pressed for and received a program that allowed importing of Mexican labor. The federal Bracero Program, initiated in 1942, supplied significant quantities of farm labor through the boom years of the war and the postwar expansion of California agriculture. Water then replaced labor as the dominant issue in California agriculture. Expansion of production caused groundwater overdrafts to resume in the 1940s. However, construction on the CVP was suspended during the war years (1942–​1944), delaying the availability of new surfacewater supplies to production areas with overdrafted groundwater supplies. In 1948, California permanently took over as the largest agricultural state in the Union in terms of value of production (Bradley 1997).

The Second Half of the Twentieth Century to the Present California emerged from the first half of the twentieth century as the leading state in the U.S. military/industrial complex. Its agriculture had weathered the Depression, had regained health during World War II, and was poised to expand as the CVP came online. At midcentury, the future must have been seen as a time of great promise for the state. The second half of the century, at least until the 1990s, met that promise. California’s population grew in fifty years post–​ World War II from 10 million to 35 million people. California gross domestic product (GDP) generally grew faster than that of the United States, meaning per capita California GDP 872  Managed Systems

exceeded the U.S. GDP in most years. The growth was fueled by rapid expansion, first in the aerospace industry and then in electronics and computers. California led the nation in both fields. Military expenditures also remained high through the 1980s. Accordingly, when defense cutbacks came in the 1990s, California suffered a disproportionately high share of defense reduction. Immigration slowed substantially, a severe recession struck the state in the early 1990s, and the state continued to suffer through a prolonged and severe drought. A rapid recovery in the second half of the 1990s, fueled in part by the “dot com” boom, quickly collapsed into a recession in the first years of the twenty-first century, bringing with it severe financial difficulties for the state. A modest recovery marked the middle of the first decade of the twenty-first century, but it then collapsed—​beginning with the housing sector in 2008 and then spreading rapidly into a prolonged, deep recession. Agricultural prices spiked upward in 2007 and 2008 but fell when the general economy collapsed in late 2008. They rose somewhat in 2010 and have been volatile since, but as of 2014 they remain at levels much above the early years of this century.

BIG WATER , GROW TH, RELOCATION, AND DIVERSIFICATION (1950–​1 970)

The decades of the 1950s and 1960s were boom periods in California. Massive investments in infrastructure continued in water projects, highways, airports, ports, higher education, and urban development. Virtually all of the increase in population was in burgeoning urban areas on the south coast, particularly in the Los Angeles Basin and the San Francisco Bay Area. Rapidly expanding housing growth, mostly in sprawling single-home subdivisions, accelerated urban takeover of agricultural land. In just twenty years, Los Angeles County went from generating the highest value of agricultural production in the state (and nation) to falling out of the “top ten” California counties in 1970 and falling below twenty-eighth after 2000. Vast stretches of Orange and San Diego Counties, longtime major producers of citrus and subtropical fruits and vegetables, were developed quickly. In the north, rapid urbanization quickly consumed much of Santa Clara County’s agriculture, pushing fresh- and dried-fruit production into the Sacramento and northern San Joaquin Valleys. This rapid relocation of production was abetted, in part, because the state’s stock of irrigated land increased from less than 2 million hectares in 1945 to more than 3 million hectares in 1970, peaking at approximately 3.6 million hectares in the 1990s. The cumulative impacts of population and income growth and urbanization, coupled with new production opportunities opened by new water transfers, led to rapid and significant changes in California agriculture. These changes included expansion in the suite of crops produced and changes in the location of production. Three examples illustrate the process. First, southern California’s dairy industry moved from southern Los Angeles and northern Orange Counties to eastern Los Angeles County (Chino and Pomona) and then to western San Bernardino and Riverside Counties in the 1950s and 1960s. The dairy industry eventually migrated north into the southern San Joaquin Valley, where it is now concentrated in Tulare and Merced Counties. Each time farmers moved, they recapitalized using the proceeds of urban land sales prices to

% of statewide harvested area of oranges

90 80

1950

70

1975

60

2000

50 40 30 20 10 0

Sacramento Valley

San Joaquin Valley

Southern California

Region

% of statewide harvested area of almonds

FIGURE 3 8.4 Oranges, share of harvested area by major agricultural production region, in 1950, 1975, 2000. Source: Johnston & McCalla 2014.

90 80

1950

70

1975

60

2000

50 40 30 20 10 0 Central Coast

Sacramento Valley

San Joaquin Valley

Southern California

Region FIGURE 3 8.5 Almonds, share of harvested area by major agricultural production region, in 1950, 1975, 2000. Source: Johnston & McCalla 2014.

0.4 1950

% Value production

expand and modernize their operations. Second, the citrus industry experienced a similar migration, first east to Riverside and San Bernardino, then north. Today, more than 50% of the state’s production is in Tulare County, compared to nearly 45% of production in Los Angeles and Orange Counties in 1950 (Figure 38.4). Third, rapid urban development in the South San Francisco Bay Area pushed deciduous fruit production out of the Santa Clara Valley into the Sacramento and northern San Joaquin Valleys. For example, in 1950 nearly 80% of the 40,000 bearing hectares of prunes were on the Central Coast. Prunes in the Sacramento Valley increased from 8,000 hectares to 20,000 hectares in the 1960s. By the end of the century, nearly all prunes were grown in the upper Sacramento Valley. This massive relocation resulted in high-priced buyouts of farm land at urban prices, allowing relocating farmers to recapitalize with larger operations that experienced substantial yield increases because of new trees, better varieties, higher planting densities, and new cultural practices. Prune yield was 3,212 kg ha-1 in 1950, 4,695 kg ha-1 in 1970, and over 6,672 kg ha-1 by 1987. Crops also moved and expanded as new water became available. One significant example is almonds. In 1950 half of the state’s almonds were grown in the Sacramento Valley, 25% in the San Joaquin Valley, and the remainder in coastal counties. Bearing hectares totaled 36,000 and yields averaged 940 kg ha-1. Almonds were aggressively planted in the San Joaquin Valley beginning in the late 1960s, and by 1970 the state had 60,000 bearing hectares. This area increased to 160,000 hectares by the mid-1980s, to 214,000 hectares by 2001, and to 310,000 by 2011. Yields approached 1,800 kg ha-1 by 2000 and have exceeded 2,200 kg ha-1 in five of the six years between 2005 and 2011. Average production has doubled since 2000. Exports expanded as rapidly as supplies increased, accounting for about two-thirds of the crop in recent years. By 2000, 80% of production was in the San Joaquin Valley, 20% in the Sacramento Valley, and virtually none on the coast (Figure 38.5). The 1950s and 1960s saw the beginning of a second fundamental transformation of California crop agriculture in terms of expansion, changing composition, relocation, and greatly enhanced yields. The dominant driver of this transformation was productivity growth funded in part by land profits from relocation to lower land-price areas. Traditional field crops, as a share of production, declined steadily, to be replaced by higher-valued, income-sensitive crops. Higher incomes plus urbanization accounted for the rising importance of fresh vegetables and horticulture products in California agriculture. These trends have contributed to significant changes in the relative shares of crop production over the last sixty years (Figure 38.6). Rising incomes after World War II also fueled a rapid expansion in consumer demand for beef. U.S. consumption rose from somewhat more than 20 kilograms per capita in 1950 to almost 40 kilograms in the mid-1970s. California’s livestock sector responded strongly to that demand expansion. One of the most phenomenal growth patterns observed was the practice of fattening slaughter beef in confined feedlots. Cattle numbers in California had been flat from 1900 to 1940, at approximately 1.4 million head. Numbers increased to 3.9 million head in 1969—​a 250% increase (Olmstead and Rhode 1997). Overall, the state’s feedlot industry exploded after World War II, increasing from 125,000 head in 1945 to 1 million head in 1965 (Scheuring 2010). Again, California had led the nation in new approaches to large-scale agricul-

1970 1990

0.3

2010 0.2

0.1

0

Field crops

Fruit and nut

Vegetable

Nursery and greenhouse

FIGURE 3 8.6 Relative shares of crop production, California, 1950–​ 2010. Source: Johnston and McCalla 2014.

tural production. However, by the 1970s large-scale feedlots were established in Arizona, Colorado, Texas, and the Midwest, areas generally more proximate to Great Plains and Midwestern feed supplies. California feedlot numbers declined. Per capita beef consumption also steadily declined after the 1970s, stabilizing around 30 kilogram per capita in the 1990s and early 2000s. Agriculture  873

TA B L E 3 8 . 2 Transformation of California’s dairy industry, 1950–2011

Average number of cows per farm

Year

Number of cows

Number of farms

1950

780,000

19,428

40

3,500

1970

750,000

5,000

150

6,000

2011

1,769,000

1,668

1,101

10,600

kg / cow

Source: California Department of Food & Agriculture 2013.

%of total production

70

Meat animals Dairy products

60

Poultry and eggs

50 40 30 20 10 0

1950

1970

1990

2010

Year FIGURE 3 8.7 Relative shares of livestock and livestock products, California, 1950–​2 010. Source: Johnston and McCalla 2010, California Department of Food & Agriculture 2013.

California then underwent a phenomenal transformation of the dairy industry (Table 38.2, Figure 38.7). In just over sixty years cow numbers have more than doubled and milk yields per cow have tripled. The number of farms has dropped tenfold as average herd size has grown from 40 to 1,101 animals. The dairy industry emerged as the dominant commodity in the agricultural portfolio of California. In 1993, California overtook Wisconsin as the number one milk producer in the nation.

UPS AND DOWNS, INTENSIFICATION AND INTERNATIONALIZ ATION (1970–​2 010)

Despite this record of rapid growth, the next four decades were to be even more explosive and also more unstable. Whereas the 1950s and 1960s were characterized by relatively stable prices, increased price volatility in the next four decades would lead to substantial swings in the profitability and economic sustainability of firms in California agriculture. The early 1970s can be characterized as a period of aggressive expansion fueled by improving world markets and concern about “feeding a hungry world.” Product prices were strong for food commodities. With strong prices came a rapid run-up in U.S. farm asset values. Worldwide market demands slowed later in the 1970s, but U.S. farmland values continued to rise into the early 1980s. This period witnessed an eroding of California’s heavy reli874  Managed Systems

ance on production of undifferentiated commodities toward a more diverse, specialized agriculture involving higher-valued, more capital-intensive crops. The mass of production for many products shifted into the San Joaquin Valley from both the south and the north, and the number of commodities produced grew from two hundred to four hundred. The period included two contrasting water-resource trends that greatly influenced by population growth. First, California agriculture seemed flush with new surface-water supplies at the end of the 1970s. Increased surface-water deliveries occurred following completion of Oroville Dam and San Luis Reservoir in 1967 and 1968, respectively, and with extensions of the California Aqueduct serving westside and southern San Joaquin agriculture in the 1970s. The Kern County Intertie Canal, which connected the east side of the valley with the aqueduct, was completed in 1977. A second significant increment in surface-water availability followed the extension of the CVP’s Tehama-Colusa Canal, enabling intensification of production on the west side of the Sacramento Valley (e.g., of tomatoes, almonds, and vegetable seeds). These completions signaled the end of the decades-long expansion of major surface-water delivery systems. Second, two of the century’s more severe droughts occurred during this period—​the first in 1976–​1977 and the second over the period 1987–​1992. The former was more severe, but the latter, longer drought had a far greater impact on agriculture. Both droughts sharply reduced water deliveries from the north to meet the growing needs of San Joaquin Valley agriculture. Average runoff in the Sacramento and San Joaquin hydrological areas fell to half of normal levels in the 1987–​ 1992 drought. At the turn of the century, agriculture, in the face of resource competition from urban and environmental demands, was confronted with increasing water-resource scarcity and uncertainty.

Into a New Century By the century’s end, California agriculture was more diverse in production and less dependent on field-crops and livestock (except for dairy) production than in 1970. Contractual marketing arrangements for agricultural production had become the norm in this new, higher-valued production system, changing marketing channels and risk exposures of producers and contracting firms. Yet in the early years of the twenty-first century, prices were low and California agriculture was again concerned about its future. Some price recovery occurred in 2004 before basic commodity prices spiked wildly in 2007. Wheat, rice, corn, and soybeans prices nearly tripled. They began to fall in 2008, then collapsed with the severe economic recession that began in 2008 and lingered for over five years. As of 2014, agricultural prices in general have been very volatile but have not returned to pre-2006 levels. California’s agriculture, especially the dairy industry, suffered through this instability. Federally mandated expansion of biofuel production has pushed 2014 corn and soybean prices well above 2007–​2008 peaks, impacting heavily the costs of feeding livestock. Competitive pressures have increased for water resources throughout the state and for land in some areas, particularly in the northern San Joaquin and southern Sacramento Valleys. Environmental issues continue to command attention, with more emphasis on instream water use, dairy-waste management, new chemical standards, water quality, and particulate matter (air-pollution) concerns. The two dominant

underlying forces affecting regional shifts in the location of agricultural production remain population growth and watersupply conditions.

Defining the Agricultural and Ecosystem Regions of California In many ways, the development of California agriculture has mirrored its highly variable ecosystems. This is hardly surprising, as both natural and managed ecosystem performance are driven by the resource base and soils, topography, water resources, and microclimates. We define nine regions of California agriculture in terms of the agricultural ecology that occurs in each region rather than geographic location (see Figure 38.1). The North Coast region was one of the earliest settled but was never a major force in agricultural production since its principal products were originally fur, fisheries, and lumber. More recently, the region developed sporadically with the concentration of livestock production in beef herds, dairy cows, and, for a while, chickens on the coastal plain. Agricultural value in the North coast more recently has become dominated by wine production and, unofficially, marijuana production, which one government official correctly listed as the most valuable crop in Mendocino County for one year. The Central Coast was first developed with large, extensive cattle ranches that took advantage of the differences in growing season microclimate among the coastal, Coast Range, and Central Valley environments to move cattle seasonally from west to east and back again. However, these cattle operations depended heavily on rainfall, and many did not survive the climatic fluctuations that led to inevitable booms and busts of stocking rates. Nowhere is this more aptly described in John Steinbeck’s East of Eden (1952). During the early 1900s, more intensive agriculture emerged in the fertile, alluvial Salinas and Santa Maria Valleys, largely based on groundwater extraction. The Salinas Valley became renowned for vegetable production—​in particular, lettuce. Watsonville was the center for artichoke production given its unique microclimate. Over the last thirty years, grazing has given way to wine grape production in many of the Central Coastal valleys, which are increasingly recognized for the quality of many of their wines. We divide the Sacramento Valley lengthwise into eastside and westside ecological regions on the basis of soils, crops and, in particular, water sources and development. The eastside region includes both sides of the Sacramento River, which were the first to be developed in the early parts of the previous century as local surface irrigation districts used tributary streams and river water. These irrigation districts were formed by groups of farmers to mutually finance and develop countybased, surface-water distribution systems. A wide range of perennial crops was grown, from peaches, plums, and nectarines to olives and some field crops. Agricultural development and ecological conversion of the westside was initially dominated by extensive dryland wheat production, mechanized as much as animal power would allow. Photographs of combine harvesters drawn by forty mules show the lengths to which these large extensive farms had to go to obtain timely wheat harvests. Irrigated agriculture in the westside valley region occurred thirty years later than on the eastside and was stimulated by both the development of heavy agricultural machinery for rice production and the Tehema Colusa and Glenn Colusa canal systems. Heavy soils and ample water supplies

made the northern part of this region very well suited for rice production, much of which was exported to Japan and Korea. Further south on the westside of the Sacramento Valley, irrigation districts based on Clear Lake and Lake Berryessa, developed in the 1940s, expanded grain and row crop acreage in Yolo and Solano Counties. The Sacramento River Delta is approximately 0.25 million hectares and occupies a unique natural and agricultural ecosystem. Originally, the Delta was a fertile, tideland marsh. In the 1840s the explorer Kit Carson remarked on the extraordinary number of birds and beavers he encountered in his journey up the San Joaquin River and through the Sacramento Delta. In the latter part of the 1800s, levee-building began in the Delta to improve riverboat transportation to the gold fields. The levees quickly became necessary for water control, which allowed the agricultural development of the Delta’s fertile peat soils. This drainage of the Delta completely transformed it ecologically but also allowed extensive cropping of field crops, pears, asparagus, and other specialty crops. The new agricultural ecology in the Delta was not stable. Drainage and cultivation of the soils induces oxidization, which removes topsoil at a rate of up to 0.5 cm yr-1. As a result, Delta soils have subsided by up to 4.5 meters in some places. This soil loss both reduces agricultural potential and further weakens the levees that retain the water from fields and are now below sea level. The ecology of the Delta channels has been further modified over the past fifty years through use as part of the conveyance infrastructure moving water from northern California to the San Joaquin Valley in southern California. This use means that most of the Delta has to be kept at the lowest possible salinity levels. Thus the Delta’s historical, seasonal salinity fluctuations have been replaced by a constant mediumsalinity regime. This shift has favored invasive species, so much so that more than 80% of its marine biomass now comprises exotic species (Hanak et al. 2011). In short, the Sacramento River Delta’s natural and managed ecosystems are neither stable nor reliable for the support of either the remaining ecosystem or its use for water conveyance. Like the Sacramento Valley, we divide the San Joaquin Valley into eastside and westside, each with distinct development history, microclimates, water sources, and cropping patterns. Irrigation districts on the eastside developed mostly around the turn of the past century. The three main river systems supplying this area—​the Merced, Tuolumne, and Mokelumne—​have catchments on the western slopes of the Sierras. All three flow into the San Joaquin River, containing headwaters in the southern part of the valley and flowing north to the Sacramento River Delta. With the formation of irrigation districts on the eastside, oak savanna gave way to irrigation canals, perennial crops, dairies, grapes for table consumption, and raisins. Further east in this region, the foothills retain many features of the original oak savanna, but with a lower oak density due to cattle grazing and firewood harvesting. Because the historical settlement of this area occurred about eighty years before the westside, farm sizes and degree of mechanization were much smaller. Many crops that lend themselves to smaller-scale production persist in this region. The west side of the San Joaquin Valley was very unproductive until the introduction of the centrifugal groundwater pump in the 1920s. Given the low productivity and extensive nature of the region, landholdings were extremely large and concentrated in few hands. The introduction of groundwater for irrigation led to the classic Agriculture  875

common property problem of excess pumping. Over a period of forty years, groundwater levels on the west side of the San Joaquin Valley fell by 60 meters. By the late 1930s the unreliability of groundwater supplies made westside land owners more than willing to step in and take contracts in the newly proposed state water project, when established irrigators on the eastside backed out of the project for fear of indebtedness under Depression conditions. As state and federal water projects developed in the 1950s and 1960s, new large areas of westside irrigated land came into production. Despite discussion and planning, no formal drainage system was established. This rapidly led to significant environmental drainage problems. With expansion of the federal government’s role in irrigation projects under the National Reclamation Act, several districts on the westside came under federal jurisdiction. However, its attendant restrictions on irrigated land ownership were blatantly ignored for many years. Because of the almost complete absence of natural water supplies on the westside, growers are always acutely aware of the scarcity value of water, particularly during drought periods. This scarcity has generated both efficient distribution systems and highly efficient, large-scale farming in this region—​both criticized at times for their management. The remaining two regions include south of the Tehachapi Mountains along the South Coast and the southern desert area near the Mexican border. The South Coast region, stretching from Ventura to San Diego, was the original agricultural development in California associated with the missions. These areas rapidly developed crop specializations—​for example, Orange County was a center of citrus production until uprooted by urban development. Ventura specialized in lemons, and San Diego and San Clemente in avocados and ornamental plants. Water sources included both local surface water and groundwater until the 1930s, when the introduction of water from the east side of the Sierras into the San Fernando Valley expanded both irrigation and urban development. Once the most agriculturally productive region in the state, the South Coast counties’ agriculture now struggles with urbanization pressures and extraordinarily high water costs that threaten even high-value crops such as avocados. In contrast to the South Coast, the southern desert regions have strong water supply endowments from the Colorado River water irrigation. In particular, the Imperial Valley, with 172,000 hectares of irrigated land, arose in the early part of the century and dominates agricultural production in this region despite its large area of relatively low-value fodder crops. The Coachella Valley, known for dates, vegetables, and other high-value produce, is a smaller but highly productive area, while the Palo Verde area’s production is dominated by fodder crops, mainly alfalfa.

The Balance of Water Use and Supply in California Sources of water supply in California fluctuate strongly with rainfall and climate (see Chapter 2, “Climate”). The years 1998–​2 006 represent the typical range of California’s water years (Figure 38.8, right side). The 2001 drought significantly reduced supplies from streamflows, which were partially compensated by increased groundwater pumping and releases from storage in state, federal, and local projects. In normal water years the ratio of supply between ground­water and recycled water is approximately equal. However, in drought 876  Managed Systems

years like 2001 this ratio changes significantly with escalated pumping and much less recycled water available. Groundwater is the main stabilizing source in dry years. However, the current institution of correlative rights puts few restrictions on pumping by overlying landowners, resulting in excessive extraction. Over the last twenty years, the annual rate of overdraft (pumping excess over natural and artificial recharge rate) averaged 1.5 billion m3 yr-1. This is clearly unsustainable and a source of significant social costs in the forms of water-quality degradation and the threat of supply loss for shallow wells. Water diverted from the Colorado River also makes up an essential part of the overall California supply, particularly to the predominantly urban area in southern California managed by the Metropolitan Water District. Use of California’s water resources is dominated by irrigated agriculture (Figure 38.8, left side). The next largest use is allocation to flows in federally designated Wild and Scenic Rivers; however, this water use fluctuates much more than irrigated agriculture use or water allocated to maintain Delta outflow standards and instream fish requirements (e.g., see Figure 38.8, 1998 versus 2001 uses). Two dominant problems with the current imbalance between demand and supply of California water resources are, first, continuing overdraft of scarce groundwater reserves; and, second, the continued declines of many fish species despite significant recent increases in instream water allocations.

Principal Functions of the California Agricultural Ecosystem The California agricultural ecosystem functions very well for agriculture but malfunctions for other ecological values. The main reason for this difference is that on the agricultural side, a wide range of efficient prices and institutions dictate production and technologies. In contrast, the ecological side is marked by sustained disinterest by the agricultural industry and a complete lack of price signals for all ecological functions except those that directly impact the costs of farming operations, such as groundwater levels. Agricultural market prices are, as noted earlier, the dominant driver of the levels, locations, and quantities of crop production. Many of California’s specialty crops supply a sufficient proportion of national and international demands to significantly influence the global market. For almonds, table wine, artichokes, and many other crops, California is by far the dominant U.S. producer; thus crop cooperatives, wholesalers, and marketing boards have strong incentives to promote their crops nationally and internationally. The most prominent recent example is the Blue Diamond almond cooperative, which has steadily expanded national and world consumption of almonds without sacrificing real prices. From the 1930s through the 1970s, many California specialty crops were marketed predominantly through cooperatives. However, for the last thirty years the difficulties of running a cooperative have in many cases outweighed the advantages. In industries such as rice, citrus, and canned fruits, the role of cooperatives has been usurped by larger, efficient, wholesale marketing firms. The demand-driven, commodity-specific signals from these wholesalers and cooperatives cut across several different producing areas in California. The other main driver of the California agricultural production function, water, has a very different functional structure. It is composed of more than a thousand water and flood

Water use

Water supply Projects Colorado Federal Local

Required delta outflow Managed wetlands Irrigated agriculture Instream flow Urban

Wild & scenic rivers

125

100

75

Water (km3)

50

25

Local imports Groundwater

Change in storage - Water (km3)

[Combined surface & groundwater storage]

Reuse

State

Instream environmental

Recycled1

2005 127%

+4.7

2004 94%

-17.1

2003 93%

-6.2

2002 81%

-11.7

2001 72%

-17.6

2000 97%

-7.0

1999 92%

-15.4

1998 175%

+6.8

0

0

% of Average precipitation

25

50

75

Note: One million acre-foot ~1.233 km3

Stippling in bars indicates depleted (irrecoverable) water use (water consumed through evapotranspiration, flowing to salt sinks like saline aquifers, or otherwise not available as a source of supply)

100

125

Water (km3) Recycled Detail of bar graph: for water years 1998-2005, recycled municipal water varied from 0.25 to 0.6 km3 of the water supply.

FIGURE 3 8.8 The balance of water use and supply in California, colloquially known as the butterfly diagram. Source: California Department of Water Resources.

control districts constituted and empowered under strict local benefit conditions. Each districts acts as the retailer of water for, typically, approximately 50,000 hectares of land linked to a particular river catchment or water supply project delivery branch. As such, water districts are fervent about local values and perceive themselves to be arbiters of the local public good, with powers over water supply that cut across many different crops produced in a district. California water districts are concerned almost exclusively with supplying surface water, while groundwater pumping is almost never regulated, recorded, or even measured in a meaningful way. As discussed earlier, another highly important input to California agriculture is labor. Farm labor takes two distinct forms: permanent, more skilled labor; and temporary, seasonal, piecework labor performed by migrant workers. Permanently employed labor is usually concentrated among managers, irrigators, tractor drivers, and other skilled field hands. Seasonal migrant labor is almost all contracted through a system of independent labor contractors who organize employees and transport gangs of laborers to particular fields for precisely defined tasks such as weeding, harvesting, or pruning. Since many of the migrant laborers are undocumented, the system of loosely regulated labor contractors has the potential to malfunction. Recently, the United Farm Workers union has increased facilities available for laborers in the field and provided a safer working environment from hazards such as pesticide contamination.

Management training for California agriculture and agribusiness is concentrated at three universities: Fresno State, California State Polytechnic University, and University of California–​Davis. A large, progressive research sector exists on agricultural technology, agronomy, livestock husbandry, and marketing. Early in the previous century, agricultural research was concentrated at the land-grant universities UC Davis and UC Berkeley. More recently, the state universities have taken a strong role in applied research, as have the private sector interests and the agricultural biotechnology industry.

The Structure of California’s Organic Agricultural Production Statistics on the recent growth in size of California organic agricultural production are based on the registration of growers and processors of organic products under the California Organic Products Act (COPA). The Agricultural Issues Center at UC Davis reports the very long list of crops produced under organic conditions in six major commodity groups: field crops; fruit and nut crops; livestock, dairy, poultry, and apiary products; nursery, greenhouse, and floriculture; pasture and rangeland; and vegetable crops (Klonsky and Healey 2012). From 2009 to 2012 the number of organic growers increased 16% to 2,693, crop area increased 36% to 238,455 hectares, and gross sales increased 55% to $1.5 billion. This Agriculture  877

80

Percentage

Growers Sales

60 40 20

+ 00

00 10 0–

50

10

0 50

0

0– 25

0–

25

00 10

50

–1

0 –5 10

10 5–

5

0

0–

expansion has taken place across all the commodity groups with the exception of floriculture and greenhouse products. Sales of fluid organic milk showed the fastest growth over the four years surveyed. Contrary to popular perception, California’s highly successful organic agricultural industry is dominated by large growers (Figure 38.9). In 2012, 9% of the growers, each producing more than $1 million of product, accounted for 78% of the total production value in the state. In contrast, 50% of the growers with the smallest production value accounted for about 2% of the value produced. While this may challenge social preconceptions, it does show that California’s organic industry has matured to a highly successful sector where specialized knowledge and economies of scale of operation have pushed towards large farming units as seen with conventional production systems. In many instances the same farm management units simultaneously produce both organic and conventional crops on different lands. This shift in the structure of organic production reflects strong and growing consumer preference for organic foods based on marketable commodities and characteristics rather than method of production. Given the increasing complexity of organic production, with the exception of some niche products, this concentration of organic production by the larger growers will likely be maintained into the future.

Grower revenue ($1000)

FIGURE 3 8.9 Structure of California organic agriculture. Source: Based on Klonsky and Healy 2012.

Ecological Malfunctions of the System Groundwater Overdraft The common-pool nature of most California groundwater basins coupled with the concept of correlative rights to pump groundwater, which in many cases practically means unlimited rights, has led to the classic resource exploitation problem known as the tragedy of the commons (Hardin 1968). The overcrowding of groundwater resources continues, albeit at the reduced rate; the California Department of Water Resources estimates that the average annual overdraft in California is 1.8 billion m3, or approximately 5% of the developed water supply. In the 1920s and 1930s, overdrafting on the west side of the San Joaquin Valley was so extensive that it caused the surface of the land to subside by 11 meters in fifty years (Figure 38.10). Subsidence can lead to substantial costs of upgrading and repairing infrastructure such as dikes, canals, roads, bridges, and buried gas lines. Groundwater overdrafting can also degrade of aquifer water by driving inflow of neighboring saline waters. More recently, aquifers have been recognized as a possible solution to the water storage problem in the southern part of the state. Recharging these aquifers is termed “groundwater banking”; in the southern San Joaquin Valley, three active groundwater banks have been in operation for more than ten years and proved extremely useful in the 2009 drought. While active groundwater banking does in some sense offset the costs of overdrafting, it by no means compensates for the lost stock of groundwater and the additional pumping cost that this generates. In several cases, California groundwater basins have been adjudicated and carefully managed and controlled, but these are invariably where the aquifer is threatened by an outside appropriation or salinity intrusion. Until wider adoption of monitoring and measurement of groundwater and clear, quantitative specification of overlying rights occur, the problem of overdrafting will continue. 878  Managed Systems

FIGURE 3 8.10 Demonstrating the effect of land subsidence. Photo: Ireland, Poland and Riley 1984.

Irrigation Water Drainage The twentieth-century expansion of irrigation resulted in several associated agricultural drainage problems. These were especially severe on the west side of the San Joaquin Valley, which had no natural drainage in the past due to a subterranean, impermeable marine layer. The region experienced accumulating concentrations of selenium and other elements in agricultural drainwater along with the normal salinity load. The toxic nature of this agricultural drainwater became apparent in 1981 when it emerged that waterfowl liv-

Irrigation water Atmospheric deposition 29 12

Atmospheric losses 38.1% Runoff 17.8 Leaching to GW 195

Synthetic fertilizer 204

CROPLAND NITROGEN INPUTS

CROPLAND NITROGEN OUTPUTS

Land-applied biosolids 4.77 Land applied liquids, WWTP-FP 3.43 Manure sold off-other CAFOs 0.862 Manure applied on-dairy 48.6

Manure sold off-dairy 78.2

Harvest 130

FIGURE 3 8.11 Cropland nitrogen sources and fates. Source: Harter et al. 2012.

GW = groundwater, CAFO = Confined Animal Feeding Operation, W WTP = Waste Water Treatment Plant. FP= Food Processor.

ing in Kesterson Lake were suffering from toxic exposure to selenium (see Chapter 31, “Wetlands”). The central problem was that Kesterson was essentially a sump into which drainage water was dumped to evaporate and thus concentrate any contaminants in it. Earlier plans had called for a drain to take drainage water north to San Francisco Bay. Completion of this plan was prevented by opposition to having agricultural drainage water from the San Joaquin Valley drain out through the Delta and the San Francisco Bay. Kesterson was decommissioned as a drainage sump, and a satisfactory alternative has not yet been found. As a result, most of the area’s highly saline drainwater is conveyed to evaporation ponds, where salt and other undesirable trace elements are concentrated and await disposal or use.

Groundwater Contamination by Nitrogen Nitrates are the predominant contaminant of groundwater in rural areas; more than 90% of these nitrates leach from irrigated crops and confined animal feeding operations (CAFOs). Harter et al. (2012) estimate that less than 40% of total nitrate input to fields is removed by harvested crops (Figure 38.11). Much of this section is drawn from a recent study by Harter et al. (2012) of the sources and causes of nitrate pollution in groundwater. The study focuses on two areas in California known to have significant nitrate pollution of groundwater: the Tulare Basin in the southern San Joaquin Valley and the Salinas Valley. As a result of leaching of nitrates into groundwater, 250,000 people in these regions are at risk from nitrates in drinking water, and 1.3 million people are affected by nitrate pollution in less threatening ways. The level of nitrate pollution in groundwater is expected to increase because of the long period of time required for nitrates to percolate down to active groundwater areas. Direct remediation of nitrate-polluted aquifers is economically prohibitive given currently available methods. However, nitrate levels can be reduced slowly over time using a simple “pump and fertilize” approach that involves persuading farmers to

use the nitrates present in groundwater as part of their fertilization requirements. Deep percolation from irrigation systems would reduce concentrations of nitrates or at least not increase them. Economic studies have shown that the modest reductions possible through the pump-and-fertilize approach can be achieved at a cost that can be accommodated by the agricultural industry. However, radical reductions in the amount of nitrogen applied would result in radical increases in the cost of production and an economic downshift in the agricultural industry, a critical source of livelihoods for many of the people who suffer from nitrate pollution (Harter et al. 2012). The optimal policy combination thus seems to be a subsidized program of monitoring and treatment for moderatesized, local water utilities and bottled water as an alternative drinking supply for the smallest water systems that exceed the state’s contamination standards. The cost of the programs should be borne by nitrate users; nitrate-use charges would both raise funds for remediation and monitoring and provide an incentive for farmers to reduce their nitrate use. Currently, farmers are exempt from state taxes on nitrate purchases. The first step would be to eliminate this exemption and add a small nitrate-use charge similar to fees already imposed on other commodities such as automobile tires, batteries, and some electronic goods—​a ll of which have higher-than-average disposal and social costs that generate social externalities. Harter et al. (2012) identify many practices that would reduce both the rate of nitrate accumulations in groundwater and applied quantities of nitrate on the surface. The impact of nitrates on groundwater can be reduced by more efficient irrigation practices, which would reduce the transfer rate of nitrates to groundwater and, because the water would reside longer in the root zone, allow a greater proportion of the dissolved nitrates to be taken up by plants. Thus, improved fertilizer use and irrigation efficiency combined with a deliberate pump-and-fertilize policy could stabilize the level of nitrogen currently contaminating groundwater. However, alternative water supply systems must be developed and financed for small rural communities. Agriculture  879

Waste Disposal from Livestock Confined Animal ­Feeding Operations (CAFOs)

2002 (4.9%)

The California livestock industry overwhelmingly functions through CAFOs; 1.7 million dairy cows are concentrated, mostly in the San Joaquin Valley, in confined conditions (see Table 38.2). In addition are several large beef feedlots and significant chicken operations. This animal production produces a substantial load of waste high in nitrogen, phosphates, and other nutrients (see Figure 38.11). As described earlier, increasing nitrates in drinking water present a health hazard and a significant health risk for young children (Harter et al. 2012).

1960 (27.6%) 1900 (100%) Current rice field Sacramento Stockton San Francisco

Agricultural Interactions with Natural Ecosystems California’s wide-ranging topography and Mediterranean climate give it a widely varied natural ecology. The introduction of European settlers and their resource-based economic activities has affected all but the high Sierra regions. Here, we examine the effect that the development of California agriculture has had on the natural ecosystems in the Central Valley and to a lesser extent in the coastal regions and temperate oak woodlands and savannas. The original Spanish landgrant ranches that produced cattle for tallow and hides fell on an east-west transect that ran from the coast to the Sierras, since the full range of California’s cross-section was needed to ensure some productive grazing at all times of the year. When first considered for agriculture, California’s Central Valley was both unproductive and unhealthy, with periodic flooding and droughts and endemic diseases such as malaria and dengue. The Central Valley needed effective control of floods, drainage, and irrigation to be feasible for crop agriculture. Initial development of irrigated agriculture took place along the Sacramento River with the development of orchards and crops. The extreme floods of 1862 were exacerbated by the residual sediment from hydraulic mining, and in 1868 water reclamation districts were authorized. Irrigation districts for additional water development were authorized in 1887. Both of these local institutions were used to finance and construct levees systems to control floods and drain farming areas. The extent of native wetlands was significant as recently as 1900 (Figure 38.12). Between 1900 and 2000, 95% of California’s natural wetlands that existed in 1900 were drained and converted to use in irrigated agriculture. This conversion of wetlands to irrigated agriculture has significantly affected many species. Most affected are the migratory birds using the Central Valley as part of the Pacific flyway. There have been some slight compensations, first of which was the inadvertent creation of the Salton Sea in the far south of the state by a misguided attempt to irrigation development in 1906, and second of which is the introduction of rice production, mostly concentrated in the Sacramento Valley, which produced significant habitat and food sources in the forms of rice tailings and flooded stubble fields. Nowhere has the change wrought by agricultural development been greater than in the Sacramento–​San Joaquin River Delta. Not only was the area diked, drained, and leveled for agriculture, but the elevation of the land was changed so drastically over the last hundred years that in some parts of the Delta fields are being farmed 4–​5 meters below sea level. In the southern San Joaquin Valley, the natural lake in Tulare 880  Managed Systems

Wetlands remaining (% of 1900)

Redding

Fresno

0

60

120

Bakersfield

180 km

FIGURE 3 8.12 A historical perspective on California’s wetlands. The red shaded areas show those wetlands that still remain one hundred years later. 1900 wetlands include the yellow, orange and red areas, and 1960 wetlands include the orange and red areas. Sacramento Valley rice fields provide some seasonal wetlands function for migrating birds and terrestrial and riparian species such as the giant garter snake. Sources: California State University, Chico (2003): Dahl and Allord 1997; by Hanak et al. 2011.

Lake Basin that was originally 130 kilometers long and 56 kilometers wide was completely controlled and drained during the early twentieth century. Folklore has it that the rusting remains of the engine of the paddle steamer that used to ply the lake remains landlocked in the corner of some irrigated field. Another dramatic ecological change caused by agricultural development was the damming of most of the upstream parts of the rivers leading into the Central Valley from the surrounding mountain ranges. These dams prevented the migration of salmon and other fishes to upstream spawning grounds. Overall, the loss of fish spawning and rearing habitat over the past hundred years due to damming and irrigation development has been substantial (Figure 38.13). In addition, exotic species have been introduced for fishing, through ship ballast water and aquarium releases, and from ponds and nurseries, making the San Francisco Bay-Delta one of the most invaded estuaries in the world (Cohen and Carlton 1998; see Chapter 19, “Estuaries: Life on the Edge”). Some exotics thrive in this highly altered ecosystem, competing with natives for food, preying upon them, and degrading their habitat. The combined result of changes in water quality, habitat, access upstream, and competition from invasive species, along with other stressors, includes widespread

Main stem habitat

(including migratory pathways)

Redding

Current Former Barrier Note: The map includes only habitat in larger rivers and major tributaries: the actual number of miles of stream cutoff is much higher than shown.

Sacramento

Fresno

Los Angeles 0

80

160

240

km San Diego

FIGURE 3 8.13 Loss of salmon spawning grounds. Source: Moyle

et al. 2005.

impacts to California’s native fishes. Hanak et al. (2011) estimate that 73% of these native species are in decline. In contrast, populations of exotic black bass (Micropterus sp.) in the Delta have increased rapidly, and the Sacramento–​San Joaquin Delta is now regarded as one of the premier bass fishing locations in the country (Bass Master Magazine 2013). The irony of an estuarine system being defined as among the top “lakes” in the country, based on the prevalence of an invasive species, has not escaped those concerned with sustaining what is left of a California estuarine system.

Synergies and Trade-offs After his seminal book Cadillac Desert, Mark Reisner (1986) thought that it was possible for irrigated agriculture to yield ecosystem benefits as well as costs. Accordingly, he turned his attention to the potential for joint production of agricultural and environmental services. Reisner focused on rice, a surprise since he had once described California rice as a tropical crop being grown in an arid environment. He focused on the potential for producing duck habitat as well as rice through additional flooding in the fall season. An additional motivation for the farmers to shift to fall flooding was a need to incorporate rice straw into the heavy, rice-growing soils. Beginning in 1992, rice farmers had to reduce use of burning to remove rice straw in order to mitigate significant air pollution effects from smoke. The phaseout of rice straw burning reduced the total area burned by 75% over ten years. Currently, only 12% of the total rice area is burned.

This reduction in burning required that an alternative straw disposal mechanism be developed. Fall flooding of rice fields significantly aided decomposition of straw and its incorporation into the soil before planting the following season. In this case, a shift in agricultural production technology simultaneously reduced air pollution and enhanced waterfowl habitat. The shift was not without cost to farmers, but the production of an environmental service while reducing pollution conferred ecosystem services with economic efficiency. A second example of joint ecosystem production with agriculture is emerging in the alternative management of floodplains. The Yolo Bypass in northern California has acted effectively as a flood control mechanism, in part protecting Sacramento, for the past seventy years. As part of its original design, the Bypass floodplain area is under active cropping by farmers who grow products ranging from processing tomatoes and rice to dryland and irrigated pasture. In addition, 6,800 hectares of the Bypass floodplain have been used for a wildlife reserve over the past twenty years, partially funded by revenues from the crop area. Most of the floodplain is farmed on crop leases reduced in price to compensate for the flood easement that occurs when a fixed weir, in the northern part of the Bypass, is overtopped by a rising river. The Bypass has flooded naturally in about half of the last twenty-six years, with some of the floods extending into the April–​May planting season. Over the past three years an additional ecosystem service, namely the provision of a short-term rearing ground for young salmonids from the winter salmon run, is emerging as potential and very valuable use for the Bypass floodplains. Ongoing experiments, in which the growth of tagged young salmon is compared in the artificial floodplains versus the river mainstem, show a growth advantage from floodplains covering crop stover from the previous year (Jeffres et al. 2008). Waterborne daphnia, food for the salmonids, thrive on the decomposing rice stubble in the floodplain. In short, some of the excess energy from floodplain cropping activity appears to be transferred over a short period to subsidize young salmon, allowing agricultural operations on the floodplain to provide an additional ecosystem. In addition to the fishery benefits, flooding at different depths and times supports waders, divers, and shorebirds that use the area at different times of year. In this form the Yolo Bypass and possibly other floodplains in California represent a truly multi-output system producing food, flood protection, salmonid habitat, and bird habitat in an interdependent system.

The Future of California’s Agroecosystem We judge the future of California agriculture as a managed ecosystem from two perspectives: the stability and sustainability of the current system and whether institutional and physical characteristics of the system provide enough feedback for it to adjust to future shocks such as climate change and market shifts. The past hundred years have shown that California agriculture is both highly productive and remarkably adept at adjusting to shifts in its resource base and the market demand on which it depends. Despite its rapid and successful development, several resources essential to California agriculture remain unsustainable. Groundwater stocks continue to suffer from overdraft, and many freshwater fish species in California are listed under the Endangered Species Act or declining—​a situation that recently has severely constrained producers’ ability to divert water from the Delta. Agriculture  881

While the reasons for declines in these fish populations are complex and varied, it is generally accepted that the amount of water exported from the Delta is a significant contributor, and recent legal decisions have tied exports of water for southern agricultural and urban uses to recovery of some of the Delta’s significant fish species. In addition, the salinity of groundwater in the Central Valley continues to rise. In particular, the large, productive Tulare Basin region has no natural outlet for drainage and presents the substantial challenge of identifying methods of extracting and storing the excess salt, which is imported into the basin via irrigation water and mobilized by agricultural drainage. The final, primary imbalance is in nitrates and other pollutants in groundwater. As discussed earlier in this chapter, contaminated groundwater is a significant but not insurmountable problem for the future. Johnston and McCalla (2004) identified six drivers of agricultural development in the past. Note that of those six, four—​public investment in water development, ready access to capital for agricultural development, a strong infrastructure of producer-based cooperatives and marketing institutions, and a mobile, plentiful, and low-cost labor supply—​ are unlikely to continue to drive agricultural development in the future. Two new entrants—​environmental and safety regulations and urban and environmental competition for resources—​w ill have a dampening effect on development of irrigated agriculture. This change in the drivers of development of irrigated agriculture is likely to result in several broad trends over the next twenty-five years. First, the growing scarcity of water will force producers to downsize irrigated agriculture in some parts of the state in terms of land area irrigated and amount of water used. However, there will be countervailing growth in demand for California agricultural products, which have the unique advantage of being upmarket products for which demand increases as personal incomes rise. Standard projections show that agriculture in the state can grow in economic value, profitability, and market influence in the foreseeable future. This increase in the proportion of high-value fruit, nut, and vegetable crops demanded by the market will influence how farmers can respond to California’s seasonal water scarcities and will place a premium on quality farm labor and increased mechanization. Another trend foreseen in the food market is an emphasis on the safety and quality of the commodities produced, even requiring traceability of products. This approach to food safety favors the kind of large-scale, organized production found in California over dispersed, small-scale peasant production from competing nations such as China. Another trend that favors California agricultural production and the resulting agroecosystem is the shift in public support away from commodity price-support programs and towards agricultural programs that generate public environmental goods. This trend, which has been predominant in Europe for the past fifteen years, is showing signs of traction in the United States in recent federal farm bills. To reconcile the current price-based structure, which made the California agriculture system what it is today, with increasing demand for environmental goods, transparency in production, and food safety, several shifts in institutional priorities will have to occur. Water is the driving and defining resource for both environmental goods and irrigated agriculture, so the institutions that have arisen over the past 150 years to control water allocation and development will have to adapt. Those institutions worked very well in the past but will be too rigid to 882  Managed Systems

accommodate future increases in scarcity and requirements for equitable reallocation among sectors and locations. The inherent tension between agricultural and environmental water use can be modified only if the environmental sector has some degree of control over its water resources, allowing it to actively respond to varying interannual scarcity, using trade-offs and triage to ensure stable environmental systems. Past interactions between California agriculture and demands for ecosystem services have often been adversarial, and conflicts over resource allocation between agricultural and environmental uses are often perceived as requiring substitution of one use for another. An alternative approach is to look for opportunities for joint production of ecosystem services and agricultural crops. It is likely that this approach, which can be incorporated into the structure of California agriculture, will be needed to accommodate the increasing demand for ecosystem goods in the future.

Summary The managed ecosystem that is California agriculture is highly variable in both the types and productivity levels of crops grown. This reflects the diversity of the soil and water resources and the microclimates that support California agricultural production. The development of the system has been driven predominantly by three factors: changing market demand for crops, water availability and its development, and technological improvements in both crops and growing methods. This massive managed ecosystem has not evolved without severe impacts on California’s natural ecosystem. Irrigated agricultural development on this scale has had a major impact on historical wetland areas and riparian corridors. The first major change in the ecosystem was the drainage of about 85% of California’s natural wetlands over an eighty-year period. Some partial compensation for this receding habitat occurred through the development of rice stubble as a food source for the Pacific flyway birds in the north and the mistakenly created Salton Sea as a source of food in the south. However, many of the other ecosystem services from wetlands were lost. Construction of dams for surface water storage on most of the state’s major rivers led to severe reductions in aquatic ecosystem functioning, such as the loss of natural salmon spawning grounds. The development of fish hatcheries has not compensated for this loss of spawning grounds. In addition, the development of highly sophisticated irrigated agriculture has led to a number of unintentional, byproduct pollutants in the ecosystem. Among these are nitrogen, selenium, and salts. Pesticide residues and metals are also a serious concern. The overdevelopment of groundwater pumping in some regions has led to land subsidence and loss of some vernal pools. On the positive side, California agriculture is immensely valuable in terms of the products produced, jobs and welfare generated, and the development of the California economy. In recent years the agricultural industry has become more conscious of both ecological impacts and values and has reduced its ecological impacts in some areas by reducing excessive pesticide use and agricultural burning. In addition, some crops are now grown in ways that enhance ecosystem benefits from their production. Notable examples include rice production in the Sacramento Valley and field crop production on floodplains, which have both been shown to directly contribute to the food webs of wild fishes

and birds. We anticipate that this trend towards reconciling agricultural production and ecosystem benefits will continue at an increasing rate in the future.

Recommended Reading Hanak, E., J. Lund, A. Dinar, B. Gray, R. Howitt, J. Mount, P. Moyle, and B. Thompson. 2011. Managing California’s water: From conflict to reconciliation. Public Policy Institute of California, San Francisco, California. Olmstead, A. L., and P. W. Rhode. 2004. The evolution of California agriculture, 1850–​2 000. Pages 1–​2 8 in J. B. Siebert, editor. California Agriculture: Dimensions and Issues. Giannini Foundation of Agricultural Economics, Berkeley, California. Reisner, M. 1993. Cadillac desert: The American West and its disappearing water. Revised and updated. Penguin Books, New York, New York.

Glossary Common-pool  Term describes a resource or good that—​ whether owned and/or regulated by government, communal group, private entity or no one—​is costly and difficult to exclude potential users from. Correlative rights  A system under California law in which groundwater is held in common by the overlying landowners, who are allowed to use any water that can be put to “beneficial use.” Farm gate value  The price of a product at which it is sold by a farm. Index of farmland values  A measure based on an initial index value of 100 for farmland in 1912–​1914. For example, an index of 160 means that farm values have appreciated 60% over 1912–​1914 values. Mother Lode  The California Mother Lode was a narrow region rich in gold deposits and extending about 190 kilometers north to south along the Sierra Nevada, through El Dorado, Amador, Calaveras, and Tuolumne Counties, along the edge of a geologic terrane called the Smartville Block.

References 100 Best Bass Lakes. . Accessed July 14, 2015. Adams, F. 1946. The historical background of California agriculture. Pages 1–​50 in C. B. Hutchison, editor. California Agriculture. University of California Press, Berkeley, California. Arax, M., and R. Wartzman. 2005. King of California: J. G. Boswell and the making of a secret American empire. Public Affairs, New York, New York. Benedict, M. R. 1946. The economic and social structure of California agriculture. Pages 395–​435 in C. B. Hutchison, editor. California Agriculture. University of California Press, Berkeley, California. Bradley, K. J. 1997. Cultivating the terrain: Public image and politics of California farming from the Depression to the post war years. UMI Dissertation Services, Ann Arbor, Michigan. California Department of Water Resources. 2009. California water plan. Bulletin 160-04. Sacramento, California.

California Department of Food & Agriculture. 2013. California Agricultural Statistics Review, 2012-2013. Sacramento, California. Cohen, A. N., and J. T. Carlton. 1998. Accelerating invasion rate in a highly invaded estuary. Science 279:555– ​57. Diamond, J. 1999. Guns, germs, and steel: The fates of human societies. W.W. Norton and Co., New York, New York. Durrenberger, R. 1999. California: The last frontier. Van Nostram and Reinhold Company, New York, New York. Hanak, E., J. Lund, A. Dinar, B. Gray, R. Howitt, J. Mount, P. Moyle, and B. Thompson. 2011. Managing California’s water: From conflict to reconciliation. Public Policy Institute of California, San Francisco, California. Hardin, G. 1968. The Tragedy of the Commons. Science 13:1243–1248 Harter, T., J. Lund, J. Darby, G. E. Fogg, R. E. Howitt, K. K. Jessoe, G. S. Pettygrove, J. F. Quinn, and J. H. Viers. 2012. Addressing nitrate in California’s drinking water. Center for Watershed Science, University of California, Davis, California. Hart, G. H., et al. 1946. Wealth pyramiding in the production of livestock. Pages 51–​112 in C. B. Hutchison, editor. California Agriculture. University of California Press, Berkeley, California. Ireland, R. L., J. F. Poland, and F. S. Riley. 1984. Land subsidence in the San Joaquin Valley, California, as of 1980. U.S. Geological Survey Professional Paper 437-I. Jeffres, C. A., J. J. Opperman, and P. B. Moyle. 2008. Ephemeral floodplain habitats provide best growth conditions for juvenile Chinook salmon in a California river. Environmental Biology of Fishes 83:449–​458. Jelinek, L. J. 1982. Harvest empire: A history of California agriculture. Second edition. Boyd and Fraser Publishing Company, San Francisco, California. Johnston, W. E., and A. F. McCalla. 2004. Whither California agriculture: Up, down, or out? Some thoughts about the future. Giannini Foundation Special Report 4. Kelley, R. 1989. Battling the inland sea. University of California Press, Berkeley, California. Klonsky, K., and B. D. Healy. 2012. Statistical review of California’s organic agriculture 2009—​2 012. Agricultural Issues Center, University of California, Davis, California. McClurg, S. 2000. Water and the shaping of California. Water Education Foundation, Sacramento, California. Moyle, P. B., J. A. Israel, and S. E. Purdy. 2008. Salmon, Steelhead and Trout in California. Watershed Science Center, University of California, Davis. Olmstead, A. L., and P. W. Rhode. 1997. An overview of the history of California agriculture. California agriculture: Issues and challenges. Pages 1–​27 in J. B. Siebert, editor. Giannini Foundation of Agricultural Economics, Berkeley, California. Reisner, M. 1993. Cadillac desert: The American West and its disappearing water. Revised and updated. Penguin Books, New York, New York. Rhode, P. W. 1995. Learning, capital accumulation, and the transformation of California agriculture. Journal of Economic History 55:773–​8 00. Scheuring, A. F. 2010. Valley empires: Hugh Glenn and Henry Miller in the shaping of California. Gold Oaks Press, Rumsey, California. Smith, B. D. 1998. The emergence of agriculture. American Scientific Library, New York, New York. Starrs, P., and P. Goin. 2010. Field guide to California agriculture. University of California Press, Berkeley, California. Stoll, S. 1998. The fruits of natural advantage: Making the industrial countryside in California. University of California Press, , Berkeley, California. U.S. Census of Agriculture. 1969–​2 007. Page 866. . Accessed 7/18/2015. U.S. Department of Commerce Bureau of Census. 1967. Economics and Statistics Administration < http://www.esa.doc.gov/>. Accessed July 18,2015.

Agriculture  883

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THIRT Y-NINE

Urban Ecosystems DIANE E . PATAK I , G . DARREL JENERE T TE , STEPHAN IE PINCE TL , TAR A L . E . TR AM M ELL , and L A’SHAYE ERVIN

Introduction The twentieth-century history of California is a history of urbanization. Yet until quite recently textbooks and field guides about California’s ecology omitted most mention of urban ecosystems except in the context of destruction of native habitat. Urban areas are ecosystems by virtually every definition. In addition to dynamic and complex humanenvironment relationships, cities host diverse plant, faunal, and microbial species, sometimes with greater species richness than in natural ecosystems (Luck 2007, Kowarik 2011). While some argue that many of these species are non-native and therefore of little value, urban residents and land managers invest enormous resources in creating and maintaining “novel” ecosystems—​species assemblages and environmental conditions that may not occur in non-urban ecosystems (Kowarik 2011, Hobbs et al. 2013). A body of literature is emerging on the many benefits that urban residents enjoy from proximity to novel, diverse habitats in cities such as private yards, parks, and other types of greenspace (Dunnett and Qasim 2000, Jackson 2003, Tzoulas et al. 2007, Musacchio 2009). An understanding of the ecology of these spaces is vital for optimizing their design and management and for

understanding their broader functioning in the context of local and regional sustainability. In this chapter we address three issues: (1) a review of current knowledge about the ecology of novel urban ecosystems in California; (2) a discussion of the known consequences, both intended and unintended, of creating and maintaining these novel ecosystems; and (3) some pathways toward integrating an understanding of the ecology of cities in planning a sustainable future for California.

Urban Ecosystems of California Urban ecosystems differ markedly from natural ecosystems in several critical ways. Like agricultural systems, they are strongly influenced by human choices and management. Unlike most of the other ecosystems described in this book, the built environment plays a central role in urban ecosystems. Built structures such as roads, buildings, and water and energy infrastructure dominate cities, and their functions and impacts both within cities and in the larger regional and 885

global context can be considered from an ecological perspective. In addition to the prominent role of humans in both the built environment and the biotic components of land cover in cities, the very definition of urban ecosystems is uniquely fluid compared to the ways natural ecosystems are defined. Urban ecosystems can be found within virtually every biome in California, from the coastal regions to the mountains and deserts. Spatial delineations of the location and extent of urban ecosystems in California vary, but common classifications include the extent of urban area itself as defined by political boundaries, U.S. census definitions of urban areas, and land cover classifications (Figure 39.1). More broadly, the larger area that encompasses both the urban area and its resource base is based in the ecological footprint concept, which was developed to determine the area of land required to sustain a given urban area. Given that California cities are sustained by resources imported from many regions nationally and internationally, the determination of specific ecological footprints can be complex (Wackernagel and Rees 1996, Luck et al. 2001). According to the U.S. Census, as of 2010 California was the most urban state in the nation with almost 95% of its population living in urban areas. More than half of this population lives in the state’s three largest metropolitan areas, all of which are located on the coast: the Los Angeles–​L ong Beach–​Santa Ana Metropolitan Statistical Area (MSA), the San Francisco–​Oakland–​Fremont MSA, and the San Diego–​ Carlsbad–​S an Marcos MSA. These urban areas are located in coastal plains and valleys and tend to be surrounded by mountain ranges and public lands. For the most part, the major coastal cities of California have almost entirely displaced the low-elevation shrublands and now border montane forested lands. This has several important consequences: as undeveloped land has become scarce on the coast, population densities have increased, housing costs have risen, and population growth has shifted to inland areas in the Central Valley and the Mojave Desert. In addition, the urban expansion of the twentieth and early twenty-first centuries has placed larger numbers of people in close proximity to the urban-wildland interface near U.S Forest Service and other public lands. As a result, forested ecosystems near coastal cities have experienced increased fire frequencies, pollution, and ex-urban development. Within and surrounding California’s major urban areas, other notable environmental conditions include urban heat islands, highly altered and/or contaminated soils, unique species composition, contamination of coastal and aquatic habitats, and rerouting of hydrologic flows, sometimes over hundreds of kilometers (Taha 1997; Dwight et al. 2002; Steiner et al. 2006; Lau et al. 2009; Fenn et al. 2010; Pataki, Boone et al. 2011). We describe these features in greater detail below.

The Biotic Environment of California’s Cities Cities are commonly assumed to reduce biodiversity due to extirpations and extinctions resulting from habitat destruction (Thompson and Jones 1999) and introduction of exotic species (Hobbs and Mooney 1998). Both processes contribute to biotic homogenization across regions—​t hat is, convergence in the species compositions of different urban areas (McKinney 2006). However, several synthetic studies have Photo on previous page: The San Francisco Bay Area at night. Photo: Shutterstock. 886  Managed Systems

reported that for plant and bird taxa, species richness (the number of species) tends to increase with human population density (McKinney 2002, Gaston 2005, Pautasso and McKinney 2007, Luck 2007, Kowarik 2011). While exotic invasive species contribute to this pattern, other mechanisms also increase species richness in densely populated areas, such as cultivation of garden and domesticated species across regions (Kendal et al. 2012) as well as the tendency for cities to be located in native biodiversity hotspots (Kühn et al. 2004). There is additional evidence that these processes result in biotic homogenization (Chace and Walsh 2006, McKinney 2006). In contrast to floral and bird species, other taxa such as butterflies decline in species richness with increasing urbanization density (Lizée et al. 2011). To date, these general patterns are consistent with studies of the urban biodiversity of California, where urbanization appears to have resulted in an increase in local species richness for some taxa at a cost of increasing prevalence of exotic species. Local studies of insect and bird species have shown this pattern (Frankie et al. 2005, Pawelek et al. 2009, Blair 2001, Connor et al. 2002). The pattern is particularly prevalent for plant species because unlike most faunal species, plant diversity is heavily affected by the continual introduction of new horticultural cultivars and exotic garden species imported to California from all over the world (Clarke et al. 2013; Pincetl, Prabhu et al. 2013). Some intentionally introduced species have gone on to become highly invasive in more natural ecosystems, resulting in a homogenization of the flora of urban localities in the state (Schwartz et al. 2006). Nevertheless, at the local scale, floral alpha diversity of urban ecosystems can far exceed the native habitat they replaced. Clarke et al. (2013) found 140 tree species by direct inventory and 214 (±10) by rarefaction in the city of Los Angeles. Pincetl, Prabhu et al. (2013) found more than 500 tree species available for purchase in plant nurseries in Los Angeles County. This is in comparison to the approximately 14 tree species native to the lower elevations of southern California, where urbanization has been concentrated (Schoenherr 1995, Rundel and Gustafson 2005). The Los Angeles urban forest contains several exotic species that have become naturalized or invasive, such as Schinus terebinthifolius, which disrupts natural riparian ecosystems (Clarke et al. 2013). Exotic disease-causing organisms are also prevalent in California cities; a current concern is Huanglongbing (HLB), a citrus virus spread by the Asian Citrus Psyllid that has already decimated citrus production in Florida (Stokstad 2012). In Los Angeles, tree diversity is highest in older neighborhoods, and plant cover increases with neighborhood income (Clarke et al. 2013), a phenomenon known as the “luxury effect” (Hope et al. 2003 and 2006, Kinzig et al. 2005). This effect has been reported throughout the southwestern U.S. (Jenerette et al. 2013) and could be associated with a hierarchy of needs rather than differences in residents’ desires (Avolio et al. in review, Wu 2013). In other words, residents of varying incomes could have similar desires for biodiversity, but only high-income residents are able to fully express these desires through investments in urban landscaping and its maintenance. It is therefore appropriate to consider adding these amenities to urban public spaces instead of relying largely on private spaces and remote public lands to provide access to biodiversity. In general, the resources required to maintain irrigated, non-native vegetation in this region appear to influence social and neighborhood inequities in tree and vegetation amenities.

U.S. Census

GRUMP

NLCD

N 0

75 150

300 km

FIGURE 3 9.1 The urbanized areas of California as defined by the U.S. Census. Because there is no single, standard definition of what constitutes “urban” land cover, a number of different classifications have been developed based on population density, political boundaries, remote sensing data, or a combination of approaches. The extent of urban land in the three estimates depicted here range from 5% of the area of the state (U.S. Census) to 7% (U.S. National Land Cover Database, NLCD) to 14% (Global Rural-Urban Mapping Project, GRUMP). Sources: U.S. Census, ; GRUMP, ; NLCD, .

Ecologists and social scientists are jointly studying the factors that influence decisions about planting and managing urban vegetation, soils, and wildlife. In Los Angeles a complex set of institutional factors determine the nature and fate of tree planting programs. As in most cities in the United States, Los Angeles has many agencies but few resources allocated to the planting and maintenance of city trees, which now tend to be associated with public-private partnerships and diverse sources of funding and fundraising (Pincetl 2010b). At the household level, Alvolio et al. (2015) reported that both socioeconomic and environmental factors influence residents’ preferences for urban trees in the Los Angeles metropolitan area. Affluent residents are more likely to prefer trees, and older residents are more likely to be concerned with their maintenance costs. Residents of urban areas in the Mojave Desert portion of southern California are more likely to value shade but less likely to consider public trees important than are residents of the naturally forested, montane region. Pataki et al. (2013) used surveys of residents in the Los Angeles area to develop a classification of the factors that influence the selection of garden tree species, such as their size, water use, phenology, and presence of showy flowers. They found that these factors were related to easily measurable leaf functional traits as well as each species’ biogeographic region of origin (the continent, climate, or habitat from which the species was imported to Los Angeles). Hence, it is increasingly possible to understand the relationships between institutions, household decisions, biodiversity, and ecosystem function in the highly urbanized regions of California. This can facilitate development of urban landscaping options and designs that meet criteria for both environmental performance and social acceptance, particularly if ecologists are integrated into the design process (Felson and Pickett 2005, Felson et al. 2013).

Fewer data exist for California on how patterns of biodiversity, species composition, and other alterations to urban environments influence ecosystem processes such as biogeochemistry, nutrient cycling, and mass/energy exchange. In general, urban landscapes in California tend to be irrigated, resulting in urban forest transpiration rates that can be similar to mesic natural forests depending on tree planting density (Pataki, McCarthy et al. 2011) and lawn evapotranspiration rates close to potential evapotranspiration (Litvak et al. 2014). These enhanced water fluxes alter climate on local to regional scales (Kueppers et al. 2007, Lobell et al. 2009). Although newly urbanized land cover tends to be associated with construction fill that is low in soil organic matter (Pouyat et al. 2006, Raciti et al. 2012), intensive resource inputs from irrigation and fertilization accelerate the soil carbon and nitrogen cycles (Kaye et al. 2005; Jenerette, Wu et al. 2006; Raciti et al. 2008). In southern California lawn soils that are not frequently tilled rapidly accumulate carbon for several decades (Townsend-Small and Czimczik 2010). However, the benefits of lawn soil carbon sequestration for mitigating global warming are offset by greenhouse gas emissions from lawn maintenance and fertilization (Townsend-Small and Czimczik 2010). Furthermore, the impacts of urban vegetation and soils on greenhouse gases and other types of pollution must be considered in the context of overall anthropogenic emissions. The direct carbon sequestration of urban vegetation is a negligible fraction of CO2 emissions from fossil fuel combustion, but urban plants and soils can substantially impact local urban air temperatures through evaporative cooling and shading, providing an ecosystem service. NO2 and N2O emissions and nitrate and pesticide leaching from fertilized soils are costs, however, of fertilized and chemically treated landscapes (Pataki, Carreiro et al. 2011). Urban Ecosystems   887

TA B L E 39.1 Imports, local sources, and outputs of energy, water, and food in Los Angeles County, 2000 Energy reported in terajoules (TJ) of production from imported fuels, local solar radiation, and waste heat of combustion

Imports

Local sources

Outputs

Energy and GHG emissionsA

2,362,300 TJ

9,500 TJ

2,371,800 TJ waste heat 124 MMT CO2Eq

WaterB

1,600 MMT

3,100 MMT precipitation 775 MMT groundwater

930 MMT wastewater 4,500 MMT runoff + ET + recharge

MaterialsC

8.3 MMT food

0.1 MMT food

8.3 MMT solid waste

Source: Ngo and Pataki 2008. A. Greenhouse gas (GHG) emissions are reported in million metric tonnes (MMT) of CO2 equivalents (CO2Eq), which is the effective warming of the atmosphere that would result from the equivalent amount of carbon dioxide (CO2). B. Water is reported as MMT imported from areas outside of Los Angeles County, as well as local precipitation and groundwater pumping. Some outputs of water are accounted for in wastewater discharge from sewage treatment plants; the remainder must be found either in surface runoff, evapotranspiration (ET) from vegetation and surfaces, or recharge of local groundwater. C. Only estimates of material imports and local production of food are available, while solid waste includes food plus other materials.

Urban Metabolism, Ecological Footprints, and Life Cycle Assessment For at least the last decade, urban ecologists have distinguished between the study of ecology in cities—​t hat is, biological or biophysical processes that happen to occur within cities—​a nd the ecology of cities, in which urban areas are studied more broadly as ecosystems that involve complex interactions between biophysical and social processes (Pickett et al. 1997, Grimm et al. 2000). The shift to studying the ecology of cities has caused a need for new tools and approaches and for modifications of approaches from other disciplines in order to understand ecological processes within cities as well as the interaction of cities with regional landscapes, climate, and resource availability. The challenge going forward in the twenty-first century is to understand cities as ecosystems in which human actions, the biophysical environment, the biotic environment, and the built environment all interact (Collins et al. 2000, Pickett et al. 2001, Alberti et al. 2003). Cities are complex systems in which many different social and natural processes operate simultaneously and influence each other at varying spatial and temporal scales. The sum total of these processes yields the total flows and transformations of energy and materials into and out of cities. Ecologists and urban planners alike are very interested in minimizing the environmental impacts of cities and maximizing the positive outcomes for human well-being. This requires understanding of the total resource use of cities, the impacts of the urban environment, and social and health outcomes—​ in other words, the ecology of cities. Cities are highly heterotrophic ecosystems in that they depend on enormous inputs of energy and materials to sustain urban functions. To put this in ecological terms, most natural ecosystems capture more energy in primary production (photosynthesis) than they consume in respiration. However, in cities primary production is orders of magnitude smaller than the energy consumed in transportation, electricity, and heat production. In addition, far more materials are imported into cities for building materials, food, and consumer goods than can be produced locally (Collins et al. 2000). For example, Los Angeles imports far more energy and materials than are available locally (Table 39.1). Several related methods have been developed to quantify the total 888  Managed Systems

amount of energy, materials, and land necessary to sustain cities. Urban metabolism describes the total inputs of energy and materials and the export of waste from a given urban area (Wolman 1965, Hanya and Ambe 1976, Newman 1999). The concept originally arose as an analogy in which the city was described as a living organism that consumes energy and produces waste. From an ecological perspective, this analogy is fairly limited because more strictly speaking, cities are complexes of organisms that interact with their environment and therefore are better described as “ecosystems” rather than “organisms” (Golubiewski 2012). Nevertheless, studies of urban metabolism introduced useful methods of quantifying total imports of energy, food, and/or other materials and exports of greenhouse gas emissions, solid waste, and other pollutants from various urban areas. These analyses tend to be quite data-limited, as energy and material consumption statistics are seldom easily available at the municipal scale (Pataki et al. 2006, Kennedy et al. 2007). However, Ngo and Pataki (2008) conducted an urban metabolism study of Los Angeles County from 1990 to 2000 and found that with the exception of food imports and wastewater outputs, inflows and outflows of energy and materials generally declined per capita over the study period, reflecting improvements in urban efficiency. Kennedy et al. (2009) used these results to compare energy use and greenhouse gas emissions from Los Angeles County to nine other metropolitan regions globally and found that per capita energy consumption was relatively low in Los Angeles, likely due to its mild climate and low energy demands, while per capita emissions from the transportation sector were quite high relative to Barcelona, Cape Town, London, Geneva, New York City, and Prague. Population density, one measure of urban sprawl , explained much of the variation in transportation emissions among cities (Figure 39.2). Hence, the perception of Los Angeles as a center of “car culture” (Selby 2012) and sprawl is borne out in urban metabolic analysis. These data provide insight into the mechanisms underlying patterns of emissions, which provides a means of prioritizing emissions reductions strategies. They also establish a baseline to evaluate future improvements in land use transportation planning and design. In studies of ecological footprints, estimates of urban energy and material consumption are used to calculate the total land area needed to support urban functions. The con-

GHG (t eCO2 cap.-2)

7 Denver

6 5 4

LA Toronto

3

Bangkok

NYC Geneva

2 Prague

1

Cape Town London

Barcelona

00 ,0 20

00 ,0 15

00 ,0 10

0 00 5,

0

0

Population density (persons km-2) FIGURE 3 9.2 The relationship between per capita (cap.) greenhouse gas emissions (GHG) from ground transportation and population density in Los Angeles (LA) in comparison to nine other cities worldwide. Source: Reprinted with permission from Kennedy et al. 2009. Copyright 2009 American Chemical Society.

cept was originally applied at the national scale to show inequities in resource use among countries (Wackernagel and Rees 1996, Wackernagel et al. 1999, Lenzen and Murray 2001) but has also been applied to cities (Luck et al. 2001; Jenerette, Marussich et al. 2006). At the urban scale, the classic ecological footprint concept has been criticized for its simplifying assumptions (Luck et al. 2001) and for implying in a general sense that cities are “bad” for the environment and natural resource consumption when, in fact, concentrating populations in small urban areas is probably more efficient than distributing the human population across the landscape (Kaye et al. 2006). Nevertheless, it is informative to establish a baseline for current urban resource and land consumption against which to evaluate future change. To do this, the assumptions underlying ecological footprint calculations must be considered very carefully. Studies of both urban metabolism and ecological footprints are highly sensitive to decisions about the size and scope of the urban region, which flows to consider, and the specific source regions for energy and waste. For example, Luck et al. (2001) showed that the estimated ecological footprint of Phoenix, Arizona, can vary by almost an order of magnitude depending on these embedded assumptions. Outside of Los Angeles, few if any studies of urban metabolism and ecological footprints have taken place in California to date. The most current approaches combine several methods: urban metabolism and ecological footprints as well as Life Cycle Assessment (LCA). LCA has been used extensively in analyses of industrial processes, services, and consumer products. Applied to an urban area, LCA involves inventories and impact assessments similar to urban metabolism but coupled with a “cradle-to-grave” analysis of urban activities and infrastructure including material supply chains, remote ecosystem and health impacts, and end-of-life waste disposal (Chester et al. 2012). For example, while urban metabolism studies have quantified direct greenhouse emissions from the urban transportation sector (Kennedy et al. 2009), Chester and Horvath (2009) showed that non-tailpipe emissions, such as those associated with vehicle and road construction, actually dominate total LCA passenger transportation emissions in the United States.

They used this methodology to compare scenarios of passenger car vehicle occupancy and air transportation with and without the construction of California’s proposed high-speed rail line. They found that high-speed rail was advantageous for greenhouse gas emissions under high-occupancy scenarios but potentially disadvantageous for sulphur dioxide emissions due to California’s current reliance on coal for electrical energy (Chester and Horvath 2010). This method shows great promise for evaluating sustainability options in California. A movement has begun, spurred by these studies, to establish a central repository of the data needed to conduct combined studies of urban metabolism and LCA in the state. Life cycle analysis can also account for the supply chain impacts of resource flows into cities. Chester et al. (2012) estimated the environmental impacts of materials embedded in existing buildings and roads in in Los Angeles County. This more complete approach leads to fuller accounting of the requirements for sustaining cities (Pincetl et al. 2012).

Integrating Biotic and Abiotic Components of Cities So far we have discussed the biotic features of cities—​their faunal and floral diversity and composition—​separately from other environmental dimensions such as resource appropriation and pollution emissions. This is because, to date, studies of biotic and built components of cities have not been well integrated. Conceptually, recognition exists that this integration is necessary. There has been a call to better incorporate the biotic environment—​parks, outdoor landscaping, street trees, and greenspace—​into calculations of urban metabolism, ecological footprints, and LCA (Chester et al. 2012). Conversely, a variety of urban planning tools available are currently available that consider scenarios and outcomes associated with the built environment, but these tools do yet explicitly consider the form and function of greenspace. Enormous interest exists in California in replacing “gray” or built infrastructure in cities with “green” or biologicallybased infrastructure to facilitate urban and regional sustainability. The many possible cobenefits of appropriate design and maintenance of urban green infrastructure have been described as ranging from pollution removal in air, soils, and water (Freer-Smith et al. 1997; Beckett et al. 2000; Davis et al. 2001; Nowak et al. 2006; Pataki, Carreiro et al. 2011) to food and water provision (Smit and Nasr 1992, Alaimo et al. 2008, Draper and Freedman 2010, Litt et al. 2011); habitat provision for native and rare species (Rudd et al. 2002, Matteson et al. 2008, Goddard et al. 2010); economic impacts, particularly in the real estate market (Tyrväinen 1997, Morancho 2003, Conway et al. 2010, Li and Saphores 2012); and more general benefits to human health and well-being. Common urban features in California designed to utilize biological processes to serve these functions include bioswales and bioretention basins for stormwater infiltration, urban farms and community gardens for food provision, greenbelts and corridors for wildlife habitat, green roofs, and a variety of other spaces that serve multiple functions. These features are hybrids of the built and biological environments, and developing science-based approaches for designing, implementing, and monitoring urban green infrastructure is an emerging research area (Pincetl 2010a; Pataki, Carreiro et al. 2011; Felson et al. 2013). Burian and Pomeroy (2010) and Pataki, Carreiro et al. (2011) have pointed out that much of Urban Ecosystems   889

the literature on urban green infrastructure originates from wetter environments and that appropriate designs for more arid environments in California cities remain to be tested. Design, construction, and maintenance of urban green infrastructure is associated with environmental costs as well as ecosystem services. Such costs have been termed ecosystem disservices and include monetary costs; negative environmental impacts such as energy, water, and fertilizer inputs to urban landscapes; and negative health impacts such as the release of allergens in pollen (Lyytimaki et al. 2008). Most studies have not yet evaluated the materials and embedded energy and emissions costs necessary to build and maintain these infrastructures. In general, far fewer studies of urban ecosystems address disservices than services, so the extent of these costs is usually uncertain. Most studies of urban ecosystem services in California have focused on a set of calculators about urban trees that was developed by the U.S. Forest Service (http://itree.org). To date, these tools have been used in several California cities to estimate various benefits of urban trees including carbon sequestration, air pollution removal, stormwater mitigation, and energy savings (McPherson 1996 and 2003, Maco and McPherson 2003, McPherson et al. 2005). However, these tools are very limited in their ability to evaluate urban ecosystem services as a whole, as they have been poorly constrained by available data and currently consider only the urban forest components of urban ecosystems. The next steps in these types of quantitative analyses are tools that integrate planning and assessment of greenspace and the built environment to more explicitly consider costs as well as benefits of alternative urban designs and can therefore contribute to scenario planning and adaptive governance . A variety of quantitative tools is available for land use and transportation in cities (Bartholomew 2007), but most do not currently incorporate greenspace planning. Integrating multiple aspects of ecosystem functioning in the built and nonbuilt environments is a promising future direction in urban ecological planning.

Impacts of Urban Activities on Surrounding Areas As urban metabolism studies have shown, urbanization can greatly impact adjacent or embedded undeveloped ecosystems through increased fire, pollution, species introductions, resource extraction, and altered climate and water availability. These effects occur through production within the city and transport to other regions (pollution, invasive species, altered climate), actions of urban residents within adjacent wildlands (fire), and direct appropriation of wildland resources (water) and other resources such as building materials, fossil fuels, and other materials. Interactions among these effects are pervasive, with urban-derived nitrogen pollution (see Chapter 7, “Atmospheric Chemistry”) leading to increased fire risk in adjacent deserts (Rao et al. 2010), fires leading to increased ozone (O3) precursor production (Bytnerowicz et al. 2010), and appropriations of water for urban uses leading to plant species and associated community changes (Elmore et al. 2003). At broader scales, urban areas are dominant contributors of greenhouse gas emissions through emissions of CO2 from buildings, industry, and transportation (Pataki et al. 2006) as well as emissions of N2O from irrigated and fertilized lawns (Townsend-Small and Czimczik 2010). 890  Managed Systems

Fire at the urban-wildland interface (UWI) is a perennial challenge in much of California. The state’s chaparral and forested regions are fire-prone, with a high risk of large fires near urban areas. The most recent five-year average (2011–​ 2006) had 5,084 fires that burned 8,439 hectares annually within California. The fire most destructive of infrastructure in California, the 1991 Tunnel Fire, was only 648 hectares but burned twenty-nine hundred structures (according to CalFire). Weather patterns of unusually high precipitation (leading to more fuel) or low precipitation (leading to more flammable fuel) both increase fire risks. Increasing fire risk near urbanized areas is linked with increasing ignition sources (Keeley et al. 1999), past practices of suppression that have increased standing fuel loads (Minnich 1983), and an increasingly complex matrix of low-density urban development in high-risk fire areas (Syphard et al. 2012). Pollution from urban areas is a major contributor to degradation of wildland ecosystems, with both extensive air and water pollution. Airborne urban pollutants that affect surrounding areas are dominated by nitrogen, O3, and particulates. The majority of NOx pollutants are produced by transportation (86%) and industrial activities (Fenn et al. 2010). On-road nitrogen emissions also likely include substantial NH3 beyond what is included in existing inventories (Battye et al. 2003, Bishop et al. 2010, Fenn et al. 2010). Correspondingly, emissions of nitrogenous pollutants differ greatly between weekday and weekend periods. For many urban sources, pollution emissions have declined since the enactment of vehicular emission controls despite increases in vehicle kilometers traveled (Fenn et al. 2003, Cox et al. 2009). At present, nitrogen emissions have led 29–​54% of nonurban vegetation cover to exceed biologically determined nitrogen deposition critical loads (Fenn et al. 2010). Tropospheric ozone formation is complex and related to NOx, temperature, water vapor, light, and volatile organic compounds (VOCs). These atmospheric drivers lead to diel and seasonal variations in O3 production. O3 concentrations in forests adjacent to urban areas, for example, in the San Bernardino Mountain forests adjacent to Los Angeles, can have severe effects (Bytnerowicz et al. 2008; see Chapter 7, “Atmospheric Chemistry”). High O3 concentrations near urban areas have led to declines of several native conifer trees through mortality and reduced recruitment (Bytnerowicz et al. 2007). O3 concentrations in the regions adjacent to Los Angeles have shown marked decreases from 1970 to the present with much larger decreases in peak concentrations compared to weekly means (Bytnerowicz et al. 2007). The Sierra Nevada are also affected by O3 emissions from urban areas, with the highest concentrations often driven by airflows from San Francisco (Bytnerowicz et al. 2013). Future projections suggest a potential for reduced tropospheric O3 but with large uncertainties and spatial variation (Steiner et al. 2006). Urban discharge of water pollutants to surrounding areas including surface water, groundwater, and coastal ecosystems can be substantial. Water pollutants include sewage, nutrients, pharmaceuticals, toxins, and growing amounts of nanoparticles. Sewage contamination of bacteria and viruses from urbanization into many coastal areas presents recurring human health issues (Dwight et al. 2002, Ahn et al. 2005). Urban inputs of nutrients including nitrogen, phosphorus, and carbon into waterways can contribute to disproportionately large total loads. For example, up to 17% of total organic carbon inputs to the Sacramento River might originate from urban lands (Sickman et al. 2007). Improvements in analytical

capabilities have allowed increasing detection of pharmaceuticals and personal care products in surface water and groundwater (Kolpin et al. 2002, Loraine and Pettigrove 2006). These materials likely have health consequences for humans (Santos et al. 2010) as well as wildlife (Nash et al. 2004). Metals and polycyclic aromatic hydrocarbon contamination of stormwater runoff are associated with transportation corridors (Lau et al. 2009, Kayhanian et al. 2012). Nanoparticles, a diverse class of materials defined solely by size, are increasingly of interest in urban discharge. Recent findings have shown direct movement of a widely used nanoparticle into urban waterways (Kaegi et al. 2008). Pollutant runoff patterns can interact with fires because airborne pollutants accumulate between fire intervals and become mobilized after fire events, causing large rates of surface water contamination (Stein et al. 2012). Urbanization is associated with several mechanisms for altering species communities in surrounding areas. Habitat fragmentation in wildlands adjacent to cities can reduce animal population sizes, decrease metapopulation connectivity, and increase invasive species, leading to increased risk of native species extirpations (Bolger et al. 1997, Suarez et al. 1998). Interactions between fragmentation and altered fire regimes can exacerbate these effects (Regan et al. 2010). Pollution by both nitrogen and O3 can alter plant communities, favoring species that can either use additional nitrogen inputs or tolerate higher O3 concentrations (Allen et al. 2007, Vourlitis and Pasquini 2009, Vallano et al. 2012). As discussed earlier, cities are also frequently sources of exotic species cultivated as landscape plants. Future studies of ecological footprints and LCA might be better able to incorporate the impacts of urban activities on wildland health (Chester et al. 2012). This information can then be used in regional planning and in “conservation development”—​a form of controlled development intended to preserve or restore natural ecosystems while allowing limited growth (Arendt 1999, Milder 2007).

The Future of California’s Urban Areas California has seen rapid changes in urbanization throughout its history. Following the Gold Rush, California urbanized far more rapidly than the rest of the United States, with particularly rapid population growth in the Bay Area (see Chapter 5, “Population and Land Use”). The early twentieth century saw a rapid increase in urbanization in and around Los Angeles, while the growth of outlying suburbs and smaller settlements dominated after World War II. In recent years there has been an effort to revive and repopulate California’s urban cores and central business districts. Today, California cities are moving in several directions with a growing emphasis on sustainable development. As we have discussed, interest has grown in urban greening activities, directed primarily towards increasing tree and greenspace cover. However, these activities are not necessarily compatible with efforts to reduce water demands and curtail outdoor irrigation. It will take careful planning to select species and designs for greening efforts that minimize water costs and other unintended consequences. Another emphasis for California’s cities is to increase reliance on renewable energy sources, with both large scale solar and wind power plants adjacent to urban centers as well as broad adoption of rooftop solar collectors. In 2013 the Los Angeles Department of Water and Power began its largest rooftop feed-in tariff program (100 MW) for rooftop solar energy (http://www.ladwp.com/fit).

These large-scale programs have potential monetary as well as environmental costs and require careful consideration (Hernandez et al. 2014). Interest is also increasing in mass transit options to increase connections within and between urban centers of California. Both San Francisco and Los Angeles metropolitan regions are expanding commuter rail services. Planned for current construction are high-speed rail lines connecting Los Angeles with San Francisco and Sacramento (California High Speed Rail Authority, http://www.hsr.ca.gov/). Land cover changes will continue to be a dominant component of urban ecosystems in the future. Within the Los Angeles region and the Bay Area, densities will likely increase because of the scarcity of available land. The age of these cities will also require retrofitting of existing infrastructure, with many opportunities for incorporating new designs. In the last few decades a shift has occurred away from singleuse zoning of urban areas in which residential, commercial, and industrial land covers are spatially separated, resulting in heavy dependence on automobile transit. Increasingly, alternative planning models such as transit-oriented development and new urbanism incorporate mixed land uses and increased reliance on walking and alternative transportation (Fulton 1996, Cervero et al. 2002, Lund 2003). However, urban development is still associated with suburban and periurban residential expansion. Urban land cover types are likely to continue ongoing expansion into historically agricultural and desert regions of California (see Chapter 5, “Population and Land Use”). Planning and decision making increasingly consider climate changes both exogenous and endogenous to California’s urban areas. Rising temperatures accompany both global warming trajectories and expansion of urban heat islands (Jenerette et al. 2011). These warming trends directly affect human health, electricity usage, and evaporation rates, with further indirect effects on air quality and fire frequency and intensity (Pincetl, Franco et al. 2013). Climate changes are also altering water availability for California’s urban areas (Tanaka et al. 2006, O’Hara and Georgakakos 2008). While urban areas are a relatively small component of total water demands, they are a rapidly growing water use sector (Christian-Smith et al. 2012). Reduced snowpack in the mountains and earlier snowmelt could have large effects on water availability. Reductions in available water could lead to reduced urban irrigation and some reductions of in-home, commercial, and industrial uses via increases in water rates and/or through other incentives to reduce urban water consumption. Notably, water availability is tightly connected to energy consumption; reductions in available water resources will likely increase energy demands to pump groundwater and reduce energy production from hydroelectric sources (Klein et al. 2005; Pincetl, Franco et al. 2013). Urban-driver effects on biotic communities will likely also grow. The availability of species imported from other regions and continents into California nurseries is increasing, and plant communities seem to follow “fads” in desirability that change at decadal scales (Pincetl, Prabhu et al. 2013). The once iconic palm trees of southern California are now frequently being replaced by other species (Figure 39.3). Eucalyptus species, once widely popular, are now less desirable (http://www.laparks.org/dos/forest/eucalyptus.htm). We expect different species to supplant current favorites in their desirability over time. As a result, new exotics and invasive plants will likely arrive in California cities and increase in prominence in less developed areas. Urban Ecosystems   891

pled human-environment interactions and the functioning of novel ecosystems, as well as effective consortia of ecologists, engineers, planners, designers, and decision makers in California cities.

Summary

FIGURE 3 9.3 Removal of a palm tree in Riverside, California. Photo: G. D. Jenerette.

In summary, California’s urban areas continue to rapidly change, with evolving decision making, implementation of new technologies, and sensitivity to the dynamic environment. Land cover changes will continue both within and outside of existing urban areas. While many coastal regions have been nearly completely developed, such as Los Angeles and San Francisco, more inland regions are sites of rapid urban expansion, such as Riverside and Fresno. Within all of these existing urban and urbanizing regions, future cities in California will not likely look like the cities of today, which themselves are quite different from the cities of fifty years ago. Climate change, increasing pollution loads, and declining resources now place pressure on cities to reduce their water, energy, and resource consumption. Significant changes in material and energy flow into and out of cities will likely require large changes in both infrastructure and governance. We have shown that there are many unanticipated consequences of urbanization for other aspects of California ecosystems, such as biodiversity, biogeochemistry, fire frequency and severity, and feedbacks to local climate. It has been difficult to predict these effects a priori, as some effects of urban development such as the prevalence of novel and diverse floral and floral communities have been surprising. To better anticipate the impacts of urban development choices on the complex landscapes of California, ecologists should play a role in informing urban planning and design. However, ecological science is only one of the many factors that determine land use decisions and outcomes. The sustainable future of California will require continual advancements and refinements in theories and concepts concerning cou892  Managed Systems

California is the most highly urbanized state in the U.S., with almost 95% of its population living in U.S. Census–​defined urban areas. Historically, most urban development took place on the coast, displacing native coastal ecosystems. Recently, urban development is taking place in higher-elevation areas and inland in the southern deserts and the Central Valley. Urban development has created hybrid ecosystems that consist of the built environment as well as novel assemblages of floral and faunal species. Floral diversity, as well the diversity of some animal taxa, is surprisingly high in California’s cities. However, some of these species are exotic invasives that have escaped from cities to impact California’s natural ecosystems. Numerous other unanticipated consequences of rapid urban development can occur, such as air and water pollution and changes in fire frequencies. Scientists are attempting to quantify these effects with studies of urban metabolism, ecological footprints, and life cycle assessment. These studies will help evaluate the effectiveness of new programs to reduce the negative environmental impacts of urbanization. California's cities are now implementing new programs in alternative energy generation, green infrastructure, and restrictions on outdoor water use. A major uncertainty in designing these new elements of urban landscapes is the relationship between the built and the biological environment. The extent to which biological characteristics such as biodiversity and species composition influence urban functioning and pollution is unknown. Many questions also persist as to how aspects of the urban environment influence human health and well-being. Overall, increasing walkability, reducing carbon-based energy use, and replacing some features of the built environment of cities with green infrastructure hopefully will contribute to mitigating environmental problems and improving quality of life. New strategies and policies in urban planning in California are attempting to integrate greenspace, open space, and ecosystem services into land use and transportation planning, which requires new consortia of urban planners, engineers, policy experts, ecologists, and other scientists. California is continuing to test new policies and strategies for urban sustainability, and the results of current programs will inform the future for both the state and beyond.

Acknowledgments This work was supported by U.S. National Science Foundation grants EAR 1204442, EAR 1204235, and EF 1065831.

Recommended Reading Adler, F. R., and C. J. Tanner. 2013. Urban ecosystems: Ecological principles for the built environment. Cambridge University Press, Cambridge, UK. Collins, J., A. Kinzig, N. B. Grimm, W. F. Fagan, D. Hope, J. Wu, and E. T. Borer. 2000. A new urban ecology: Modeling human communities as integral parts of ecosystems poses special problems for

the development and testing of ecological theory. American Scientist 88:416– ​425. Kennedy, C. A., J. Steinberger, B. Gasson, Y. Hansen, T. Hillman, and M. Havrenak. 2009. Greenhouse gas emissions from global cities. Environmental Science and Technology 43:7292–​7302. Pincetl, S., P. Bunje, and T. Holmes. 2012. An expanded urban metabolism method: Toward a systems approach for assessing urban energy processes and causes. Landscape and Urban Planning 107:193–​2 02. Schwartz, M. W., J. H. Thorne, and J. H. Viers. 2006. Biotic homogenization of the California flora in urban and urbanizing regions. Biological Conservation 127:282–​291. Selby, W. A. 2012. Rediscovering the Golden State: California geography. Third edition. Wiley, Hoboken, NJ.

Glossary Adaptive governance  A dynamic method of managing complex ecosystems in which decision-making and management change in response to experimentation and new information. Alpha diversity  The number of species at a specific location, in contrast to the diversity of multiple locations or a region as a whole. Bioretention  A landscape designed to infiltrate stormwater into soil. Bioswales  A type of bioretention feature, usually linearly designed to infiltrate stormwater into soil along streets, sidewalks, parking lots, or other paved areas. Biotic homogenization  A process in which localities become more similar in terms of biodiversity, ecosystem function, and other biological characteristics. Built environment  The human engineered and constructed environment, which includes roads, buildings, pavements, and other human-built structures.

(2) the adjacent counties with a high degree of social and economic integration measured by commutes to work in the urban core. New Urbanism  A revival of older, preautomobile neighborhood designs that encourage pedestrian and nonautomobile transit. Potential evapotranspiration  The amount of water that can be evaporated from a moist surface given atmospheric conditions. Rarefaction  A method of estimating numbers of species by the relationship between the area sampled and the numbers of species found in sampled plots. Scenario planning  A process of testing the form and impacts of alternative futures for urban growth in order to develop a shared community vision for local planning. Transit-oriented development  High-density urban development focused on public transportation hubs. Urban heat islands  Areas of elevated temperature and other alterations to local climate caused by urban materials and processes. These include heat retention by buildings and pavement, heat generation from combustion, and reductions in vegetation cover. Urban metabolism  Energy and materials inflows, transformations, and outflows from urban areas. Urban sprawl  Low-density urban development that may have a number of characteristics including low population density, reliance on automobile transportation, and single-use land use and building types as opposed to mixed commercial and residential development. Urban-wildland interface  The encroachment of urban development into largely natural areas. In California this interface often occurs in foothills and higher-elevation areas that developed more recently after valleys and lower-elevation agricultural lands were fully developed.

Ecological footprint  The area needed to supply resources to a given local population. Ecosystem disservices  The costs or negative impacts of creating or managing a given ecosystem. Ex-urban  Very low-density urban development, generally lower in population and housing density than suburban areas. Green infrastructure  Human-designed and -engineered features that rely on biological processes to provide environmental services. Heterotrophic  An organism or ecosystem that cannot synthesize its own sustenance. Cities are considered heterotrophic because their food, energy, and materials cannot be supplied entirely locally and must be imported from outside the city boundaries. Hierarchy of needs  A ranking of human needs proposed by Maslow (1943) in which humans fulfill basic needs in a certain order, beginning with basic physiological requirements and moving to other social and psychological needs only when basic needs have been fulfilled. This pattern is proposed to explain the relationship between urban biodiversity and affluence, in which high levels of biodiversity are only planted and maintained by residents who can afford to fulfill other basic needs. Life Cycle Assessment  A cradle-to-grave assessment of the total environmental impacts of a given process, product, or urban area. Mesic  A moist climate with adequate rainfall for forest growth. Metropolitan Statistical Area (MSA)  A U.S. Census–​ defined core urban area of fifty thousand or more residents that includes: (1) the counties containing the urban area, and

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PAR T S IX

POLICY AND STEWARDSHIP

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FORT Y

Land Use Regulation for Resource Conservation STEPHAN IE PINCE TL , TERRY WAT T, and M AR IA J. SANTOS

Introduction Over the past 150 years a great deal of California’s most scenic landscapes has been conserved because of California’s exceptional biodiversity. In this chapter we review the history of land conservation in California from a land use and governance perspective. We first introduce some key concepts in land use planning relevant to California, then we review and describe the historical process of conservation land acquisition and land use policies. Finally, we explain how and why land use regulations, fiscal constraints, and regulatory requirements today create a complex terrain for future land conservation. One important factor is that with urbanization, ecosystems that were once isolated are now close to the state’s cities (See Chapter 5, “Population and Land Use”). These lands are the new frontier for conservation efforts but also face unprecedented, unmet challenges such as those likely posed by climate change. New legislative and funding tools are emerging for protection, as we explain below, but a legacy of structural constraints remains in place. Surely, better land use is at the heart of conservation of the state’s magnificent heritage. In the end conservation rests upon political will and the support of the state’s residents. The state legislature will have to restructure the state’s taxation system

and the way in which private property rights are interpreted and will need to create new governmental regulations that ensure the state’s social and ecological resilience to climate change. This will require a change in land development patterns and fundamentally in the way California’s residents live on the land.

Key Concepts in Land Use Planning Regulations Federal, state, and local regulations all affect California’s ecosystems and land uses. Regulations are rules adopted to implement legislation passed by elected officials to guide public policies. These regulations can be prescriptive (for example, stipulating that houses cannot be built on flood plains) or procedural (for instance, requiring one to conduct an environmental impact statement before a decision can be made). Other legislative activities, such as establishing taxation rates, also are significant as they constrain or enable activities by governments and influence private behavior. Taxation rates 899

influence financial resource availability: fewer resources are available when tax rates are low, and more are available when tax revenues are higher. All of these are important factors in shaping natural resource conservation policies and outcomes in California.

Land Use Planning Land use planning stems from the need to plan for competing interests for land: land for housing, roads, commercial establishments, and schools as well as for the conservation of ecosystems, public open space, water, and other resources and values. Today, this competition takes place over privately held land. For greater conservation to occur, more land must be set aside from other uses; more often than not, this requires its purchase. This task is generally undertaken by a governmental entity, as private entities generally do not take on long-term conservation landownership and management as they typically lack the institutional infrastructure or funding to do so. The obligation to protect endangered fauna and flora imposed by the 1973 Endangered Species Act (ESA) has added another dimension to local land use planning that must be balanced with the protection of private property rights, requiring new strategies and approaches to land use planning and financing for conservation land acquisition. As for any land purchased by government, the Fifth Amendment of the U.S. Constitution ensures that private landowners are compensated fairly by the governmental entity in question. Government is constitutionally authorized to regulate land use, known as its police power. It can, for example, through land use zoning, prevent building in flood zones. Zoning is how land use is designated, and in California this power is in the hands of city and county elected officials. Land use designations include residential, industrial, commercial, institutional, open space (i.e., undeveloped land), and so forth. Land use designations change over time reflecting public opinion, pressures for urbanization, the need for additional tax revenues, and various other conditions because land uses generate value including tax revenues. The development potential of near-urban lands with conservation value makes them more expensive to purchase than distant rural lands as the former’s price reflects the potential for urban uses.

Management of Public Lands Congress recognized the value of public lands, declaring that they should remain in public ownership. However, federal lands in public ownership are managed by different laws depending on the land type and jurisdictional management. Most are managed by organic acts that allow for multiple uses and for the most part do not recognize conservation as an exclusive use. This means there can be hiking, hunting, logging, grazing, mining, or other uses allowed. Lands with designations such as Wilderness or National Park allow no extractive uses—​only recreation such as hiking and camping. The U.S. Forest Service (USFS) is subject to the National Forest Management Act (NFMA) and the Multiple Use Sustained Yield Act (MUSYA). Bureau of Land Management (BLM) lands Photo on previous page: The urban-wildland interface in Solano County, California. Photo: Larry Ford. 90 0   Policy and Stewardship

are commissioned in Federal Land Policy and Management Act (FLPMA) (Pub. L. 94-579)—​the least restrictive in terms of uses—​to allow a variety of uses on the agency’s land while simultaneously trying to preserve their natural resources. This concept is best summarized by the term multiple-use , defined in MUSYA as “management of the public lands and their various resource values so that they are utilized in the combination that will best meet the present and future needs of the American people.” As a result, much of these public lands have a myriad of uses including large-scale renewable energy facilities, transmission, roads, and recreation including parks and off-road vehicles.

Land Use Planning Regulation in California State planning law requires that local General Plans contain seven mandatory topical elements and meet other statutory criteria, which were added throughout the years. The elements include: noise, land use, circulation, conservation, safety, open space, and housing. Mandatory elements of a General Plan can be broken into two categories with some overlaps: the built environment (land use, housing, noise, and infrastructure) and the natural environment (open space and conservation). All elements of a General Plan are considered equal and must comprise an integrated, internally consistent statement of policies (Government Code Section 65300.5). The housing element, however, is the only element required to be updated, reviewed, and certified by the California State Office of Planning and Research (OPR). In some respects, the housing element has been elevated above other elements of the General Plan as California has experienced high levels of population growth through the early 2000s and suffered from insufficient affordable housing for its inhabitants. Cities may not deny or conditionally approve certain housing projects in a manner that renders the project infeasible without making efforts to curb discriminatory practices. (This requirement is often violated in practice). No similar certification exists for other land use actions otherwise consistent with the General Plan. The core natural environment components of a General Plan include the open space and conservation elements, which must together plan for the comprehensive and longterm conservation of open space and resource-providing lands. However, even though state law mandates ambitious and detailed planning efforts for open space, including an action program that the local government intends to pursue, the majority of elements developed by cities have increasingly shied away from mapping “proposed” parks and open space on private lands. Thus we see a trend away from a commitment to parks and open space. This is likely due to fear of lawsuits following Nollan vs. California Coastal Commission (483 U.S. 825). In essence, for a local government to require public access to public land from a landowner (for example, access to the beach or zoning of private land for open space), the Court requires the demonstration of a clear relationship between the regulation and the resulting access or use. In 1987 the Court determined there must be a logical relationship between the negative impact of the project and the need for a public easement across the owner’s property (a nexus). In the Nollan case, the Court rejected the requirement of an easement along the beach in southern California because it was not related to restoration of the view from the road—​a type of access—​and constituted a governmental tak-

ing of private property rights. This has had a strong effect on

local government’s willingness to plan for the zoning of private lands for open space, even when those lands are not currently adjacent to urbanized areas. The Court was concerned with the protection of private property interests under the Fifth Amendment of the U.S. Constitution. Nollan vs. California Coastal Commission, however, also had a significant chilling effect, as local governments have now shied away from even designating lands for future parks and open space, or preservation, for fear of landowner lawsuits. Local governments also expressed concerns about designating lands as open space, as designation could preclude the ability of these lands to be developed and thus generate future tax revenues. This marks a change in political culture away from government regulation and towards more “voluntary” approaches, such as one in which property owners themselves enter into agreements to preserve public access or to refrain from developing their lands in exchange for property tax relief or other tax relief. Such agreements can actually then be part of the property deed in perpetuity (Crew 1990). This brief overview of land use planning in California already hints at the paradox that the state faces: to be able to create revenue for land purchases, cities or regional governments must grow; in turn, this quest for growth limits the land available for conservation. In the next section we describe and illustrate the historical process of conservation land acquisition. We then describe the evolution of land use policies in the state and their current status and assess whether they will be able to meet the challenges ahead.

History of Conservation Land Acquisition Conservation of California’s remarkable natural resources was under way by the later decades of the nineteenth century. After the American Revolution, the U.S. federal government inherited all lands west of the Alleghenies with its new territories. It established a policy of selling and distributing these lands through many public land distribution acts (including the Homestead Act 1862). By the mid-nineteenth century, much land had shifted to private ownership, but mountains and desert lands often remained undesired by the public. A movement to protect some of these lands arose, and in 1864 Yosemite Valley and the Mariposa Grove of Giant Sequoia trees were protected by Abraham Lincoln, with a grant of 8,100 hectares of federal land to the state of California. Yosemite Valley protection was the first example of the president withdrawing lands of exceptional beauty from the possibility of sale to the public, and the same mechanism was used to reserve National Forests for conservation purposes. At this time, many of the nation’s forests had been purchased by private companies and logged. Fear of timber shortages over time, as well as a growing understanding of forests’ importance for watershed functioning, led to presidential acts to reserve remaining forests in public ownership, especially in the western states. Much of California’s early conserved land became the backbone of the lands of the National Park Service and the National Forest Service. California is arguably the state that led the implementation of the concept of resource reserves beginning in the nineteenth century (Barton 2000). Californians lobbied Congress and the president about the public interest in the public lands: that they should be conserved, not sold. The San Gabriel and San Bernardino Forest reserves, for example, were created

from the public domain in 1892 and 1893, respectively, by President Harrison for their watershed function. California is​ the birthplace of the Sierra Club (1892) and the Sempervirens Fund (1900), aimed at preserving redwood forests. Without question, interest in protecting the state’s scenic beauty has been important for nearly a century and a half. Local organization actions are exemplified by the early acquisition of remaining redwood forests in Santa Cruz County—​Big Basin State Park—​in 1902 by the Sempervirens Fund, whose ownership was then transferred to the State Parks Commission (see Chapter 26, “Coast Redwood Forests”). State interest in land conservation was reflected in 1927 by the then recently created State Parks Commission, which instructed Frederick Law Olmsted Jr. “to make a survey to determine what lands are suitable and desirable for the ultimate development of a comprehensive, well-balanced state park system, and to define the relation of such a system to other means of conserving and utilizing the scenic and recreational resources of the state” (Olmsted 1929). The resulting survey set the vision for the creation of the state parks system. California is not only one of the most biodiverse areas in the United States, it is one of the most biodiverse regions on Earth (Brooks et al. 2002; see Chapter 11, “Biodiversity”). Today, 25% to over 40% of California land is designated as Open Space depending on the definition used thereof (see Box Figure 1a). Open Space corresponds broadly to “lands protected through fee title ownership (absolute ownership) by a public agency or non-profit land conservation organization” (GreenInfo Network 2013). Open Space includes a wide array of property types, from city parks to National Parks, Bureau of Land Management (BLM) lands used for recreation and resource extraction, and Department of Defense lands. However, Open Space does not include private land in conservation easements. A conservation easement is an agreement by a property owner to maintain his or her land in private open space—​no development will occur—​in exchange for a tax rebate on that property. Conservation easements can have a time limit, after which they must be renewed or not. They are permanently tied to the title to the land and inhere at the sale of a property. Arriving at a definitive estimate of conserved land area is not easy as there are many types of land conservation in the state, including those just described as well as contracts under the Williamson Act (that can be exited after ten years) to keep agricultural lands in agriculture. Large tracts of land were set aside from the remaining public domain at the turn of the twentieth century. After that, additional park land had to be acquired (mostly by federal agencies) from the remaining, privately owned open space (see Box Figure 1a–​b). This is the process we describe below—​ the evolution of land use policies, their land development implications, and the revenue streams for conservation land acquisition. Geographically, early land conservation focused on mountain regions, mostly in the central and southern Sierras and in the mountains surrounding Los Angeles (see Box Figure 1a). Land set-asides and early acquisitions then expanded to the northern California mountain ranges until the 1920s (see Box Figures 1a and 4). In the 1930s, California State Parks took the lead on land acquisition for additional parks. In the 1940s, funded by the recently formed State Park system, large tracts in southern California were purchased; this pattern continued into the 1950s. These acquisitions were also greatly aided by regulations passed by the state legislature in the 1940s (continued on page 906) Land Use Regulation for Resource Conservation   901

BOX 40.1 RECONSTRUCTING CALIFORNIA’S LAND CONSERVATION HISTORY

Maria J. Santos Today, fully one quarter of all land in California is protected in some form or another, but our knowledge of how this came to be is fragmentary. In 1864, Abraham Lincoln granted approximately 8,100 hectares (20,000 acres) of land to the State of California—​land that would become Yosemite National Park and the Mariposa Big Tree Grove (Barton 2000). In the 1880s the state made initial steps to impose limits on the logging of its redwood forests. In the 1930s, U.S. Forest Service scientist Albert Wieslander painstakingly documented California forest resources (Wieslander 1935). Urban sprawl in the latter half of the twentieth century prompted communities to secure Open Space amenities close to residential areas (Pincetl 2004). All of these seemingly disparate actions have, in concert with hundreds of other actions over the past century and a half, shaped the current conservation network in the state. Since conservation decisions are intrinsically a human decision-making process, the likely future behavior of a human society is strongly tied with its past (Szabó and Hédl 2011). Unprecedented rates of change in land cover and climate over the past century challenge the ability of conservation networks to conserve biodiversity, among other goals. The timing of historic land acquisitions may have had an effect on whether conservation goals as currently defined have been met (Meir et al. 2004). The timing of conservation acquisitions may be a function of multiple factors, such as delays linked to requests for more information, the inherent dynamics of natural systems, and insufficient traction (funding, policies, etc.), among others (Czech 2004, Grantham et al. 2010, Knight et al. 2008). I thus took a step back to reconstruct the conservation land acquisition history of California and to assess the contribution of land additions at different time periods to the overall conservation network of the state. Conservation history describes the processes that created today’s parks, preserves, and refuges. It entails both a spatial and temporal depiction of conservation activities, implementation, and achievement. Once constructed, this history can point to past successes and shortcomings, aiding future conservation efforts. Many parks or Open Spaces have documented who inspired their acquisition, the steps towards their establishment, and the observed changes in them over time—​but this knowledge is not scalable to a region, state, or country. A sober appraisal of past efforts can galvanize future action, yet much work on the topic remains to be done. As a potentially emerging field, conservation history aims at linking historical ecology and environmental history. Historical ecology answers ecological questions using historical data, while environmental history looks into the effects of human presence on the environment they experience and modify. By their integration we can benefit from the quantitative perspec-

tive added by historical ecology (here narrowed to historical conservation biology) and the qualitative and legacy perspective from environmental history. Conservation decisions are more often than not a political decision, and thus the integration of both fields is only a necessary requirement to be able to assess the success of conservation. We know that conservation success is geographically and temporally concentrated in hotspots (Brooks et al. 2002) and hottimes (Radeloff et al. 2013). The process of reconstructing the state conservation history involved three steps: the spatial location of the Open Space properties, attribution of the time of acquisition and establishment, and data quality control and assessment. Open Space corresponds to “lands protected through fee title ownership by a public agency or nonprofit land conservation organization” (GreenInfo Network 2013). Below I describe the process and data sets used for such reconstruction and the resulting outputs. Over fifty thousand Open Space parcels exist in California. They are managed by over eight hundred agencies with varied archival structures. The development of geospatial data for all Open Spaces in the state has been a large effort lead by the GreenInfo Network, and it is now systematized in the California Protected Area Database (CPAD) (GreenInfo Network 2013). CPAD is a GIS inventory of all fee-protected Open Space properties in California assembled by collating and curating data from management agencies. Open Space management agencies retain fee title ownership —​t hat is, property rights over the land designated as Open Space. I selected this database rather than the World Protected Areas database (IUCN and UNEP-WCMC 2013) because the former has a more comprehensive representation of the Open Space properties in California. It includes properties that range from city parks to national parks, thus it is likely to better reflect the state’s conservation history. However, this database does not include attributes that describe the acquisition and establishment dates of every Open Space property in the state. To complement this information gap, my research team and I obtained from the federal and state agencies and nongovernmental organizations that manage Open Space properties data on acquisition date (when the fee title was purchased) and establishment date (when the property was open to the public). After one year of contacting agencies, we got an enormous amount of data. We retrieved information on acquisition date for a total of 35,807 properties (67%), corresponding to 110,300 square kilometers (98.85%) of the state’s Open Space (Figure 1a–​b). Merced, Orange, San Joaquin, Stanislaus, and Sutter Counties were harder to collect information from (dates were acquired for ca. 70% of their Open Space area). To complement this data gap, in collaboration with GreenInfo Network, we developed an online crowdsourcing platform for data collection, launched

Timeline of Open Space

A

B

Open Space

Not dated 1800–1809 1840–1849 1850–1859 1860–1869 1870–1879 1880–1889 1890–1899 1900–1909 1910–1919 1920–1929 1930–1939 1940–1949 1950–1959 1960–1969 1970–1979 1980–1989 1990–1999 2000–2009 2010–2019 Unknown

0

50 100

200

300

km 400

Agency City County State Federal NGO Special Districts Private Unknown

0

50 100

200

300

km 400

B O X 4 0 .1 F I G U R E 1   Timeline of Open Space acquisition and establishment in California over the past 150 years: (A) timeline of California current Open Space extent; and (B) Open Space land ownership/management responsibility, with governance types represented by different colors (city, special district, county, state, federal, nongovernmental organization [NGO], private, and unknown).

in mid-2013. It is geared to harvest the knowledge of the public on the dates and history of each Open Space property (http://www.mapsportal.org/mapcollab_ dates/). I used only data from the management agencies to construct the mapped history herein, amounting to 99% of the area of Open Space in the state. Sixteen thousand Open Space properties were acquired in California throughout the past 150 years as 53,337 parcels. These properties cover a quarter of the state’s area (112,156 square kilometers). The first hotspots of California conservation history were the mountains, followed by the southern coast and finally the deserts. The first hotspot of conservation also corresponds to the first hottime for conservation, which occurred at the turn of the twentieth century. However, throughout the twentieth century, hotspots and hottimes were not matched. During this time, hottimes in 1930s and 1970s corresponded to acquisition of land in many parts of the state but not in particular areas of the state (see Figure 1a–​b). The hotspots and hottimes of conservation are a result of the push and pull of pro-development and pro-conservation policies and action in the state’s history, as described elsewhere in this chapter (see Table 40.2). The growth of the conservation land networks in the state’s first most populous counties varied among locations. The early, most populous counties include Los Angeles, San Diego, San Francisco, Alameda, San

Mateo, Santa Clara, and Sacramento (Figure 2a–​g, Figure 3). While the counties vary in their area and proportion of Open Space (see Table 40.1), the earliest acquisitions occurred in the county of San Francisco with the reservation of the Presidio first by the Department of Defense and then with fee ownership transferred to Golden Gate Park (Figure 2c). Temporal Open Space acquisition patterns differ across locations, with (1) a stepwise pattern of acquisition in southern counties of Los Angeles and San Diego, (2) a steady acquisition rate for San Francisco, and (3) a very fast and more recent acquisition rate for other Bay Area counties and Sacramento (Figure 2a–​g, Figure 3). Despite these differing patterns, northern and southern California counties have similar proportions of Open Space land. While the southern coast of California was a hotspot of land conservation, its acquisition was phased, with an earlier history of less conservation followed by more recent acquisitions and the parallel development of systematic plans for biodiversity and their natural habitats. Today, counties like San Diego are leading the state in the development of Habitat Conservation Plans (HCPs) and Natural Community Conservation Plans (NCCPs). The outcomes of such experimentation with scientifically sound plans are commendable, but these actions are still of voluntary nature. They also unfold gradually; some of these HCPs and NCCPs took over ten years just in the planning phase. The hotspots and hottimes of land conservation in the (continued)

B O X 4 0 .1 F I G U R E 2  Timeline of conservation land acquisition in the most populous counties. Light blue areas indicate water. Sources: Santos et al. 2014a and 2014b.

4000

Area Properties

3000

8000

Los Angeles

6000

2000

4000

1000

2000 0

0 6000

5000

San Diego

4000

4000

3000 2000

2000

1000 0

0

25

300

San Francisco

20 10

100

Cumulative area (km2)

5 0

0

500

800

Alameda

400

600

300 200

400

100

200 0

0 500

800

400

San Mateo

300

600

Cumulative number of properties

200

15

400

200

200

100

0

0 800

800

Santa Clara

400

600

300

400

200

200

0

0

250

1200

Sacramento

150

800

200 100

400

50

0 Not dated

Unknown

2010–2019

2000–2009

1990–1999

1980–1989

1970–1979

1960–1969

1950–1959

1940–1949

1920–1929

1930–1939

1910–1919

1900–1909

1890–1899

1870–1879

1880–1889

1860–1869

1850–1859

1840–1849

1800–1809

0

B O X 4 0 .1 F I G U R E 3  Timeline of area and number of parcels conserved for the most populous counties as of 1910. The thick line represents the cumulative area in Open Space acquired. The fine line is the cumulative number of parcels acquired. Source: Santos et al. 2014a.

(continued)

(continued)

state can also be linked with the presence of governance structures that allow land purchase. Governance is the level (for example, federal, state, city, etc.) within which a given agency operates and has jurisdiction. Not all governance levels existed a century and a half ago (Figure 4). Federal agencies were one of the few governance levels that existed at the turn of the twentieth century and were the ones that could take

land from the public domain into private ownership. Other agencies appeared through the history of the state with a mandate to acquire land for conservation. For example, “special districts” emerged in the 1940s as a result of political victories over pro-development policies. Regulation in the 1970s also required Open Space elements to be included in land use General Plans (see Table 40.2).

Easements

Agencies

NGO’s County Parks Special Districts State Parks National Park Service

60000

100000

50000

800000

40000

600000

30000

400000

20000

200000

10000

Not dated

Unknown

2010-2019

2000-2009

1990-1999

1980-1989

1970-1979

1960-1969

1950-1959

1940-1949

1930-1939

1920-1929

1910-1919

1900-1909

1890-1899

1880-1889

1870-1879

1860-1869

1850-1859

1840-1849

0 1800-1809

0

Cumulative number of properties

120000

Cumulative area (km2)

Forest Service

1st National Park

B O X 4 0 .1 F I G U R E 4  Timeline of California conservation land acquisition and dates in which natural resource management agencies were created. The broad colored line represents the cumulative area in Open Space acquired. The fine gray line is the cumulative number of parcels acquired. Source: Santos et al. 2014a.

that allowed special districts (a type of government unit formed to provide a specific service, such as schools, sewers, street lighting, or parks) to also be created to acquire lands for conservation (see Box Figure 5a). From 1950 to 1980 more scattered and smaller properties filled in the conservation network as urbanization spread to adjacent private lands and larger tracts of remote lands were already publicly held. High federal involvement through funding for acquisitions occurred again in 1970–​1990 and was supplemented by nonprofit organizations’ land acquisitions (see Box Figure 5b). From the 1980s to today large tracts of desert lands have been acquired that had not been considered valuable until then (see Box Figure 1a–​b). While a great deal of land has been protected and kept or placed in the public domain, ecosystem losses continue as urbanization encroaches onto unprotected private lands. The most populous counties in 1910 showed very different processes and proportions of conservation land acquisi906   Policy and Stewardship

tion (Table 40.1). These counties’ population is today mostly confined to incorporated cities, with the exception of Sacramento. Even within heavily urbanizing areas, such as San Diego and Los Angeles, a high proportion of county land area has been dedicated to open space via a stepwise process (see Box Figures 2 and 3). San Francisco, the smallest county and the one with all its city population coincident with county boundaries, had earliest acquisitions and maintained a constant rate of acquisition throughout the twentieth century. San Francisco is also geographically bounded, and conservation options are limited. It is in this part of California—​ the Bay Area—​that the first ideas for regional planning were implemented. Alameda, San Mateo, and Santa Clara showed a peak of conservation land acquisition in 1950s. While California’s policies lead the nation in ecosystem conservation, they are inadequate to meet future challenges such as climate change and potential urban growth. This is because regulation in California is piecemeal, fragmented,

Prior to this time, General Plans could voluntarily include Open Space elements, but only in the 1970s did this requirement become mandatory. This created a way for the state to supervise development planning (which remains within the jurisdiction of the cities) and to be able to force the protection of natural resources at a state level. Finally, some of these governance levels appeared as a reaction to legislation such as Proposition 13, which affected the way funds were gathered for conservation land acquisition. Prior to Proposition 13, a part of development fees were set aside for land acquisition by special districts or other agencies. Proposition 13 altered the redistribution of development fees, greatly reducing the amount directed to conservation land acquisition. A supermajority vote was also included in the legislation, which requires a two-thirds’ vote in favor to change tax funds distribution. To respond to these decreases in development funds directed towards conservation land acquisition, an increase in land purchases by nongovern-

A. Timeline of Special Districts San Francisco Area

0

50

100 km

mental organizations took place (Figure 5a, b). Later legislation enabled conservation easements. These were innovations towards the maintenance of land as Open Space and the acquisition of its fee title at a time when conservation science was already advocating strategic planning. These forces led to complementary acquisitions of ecosystems that were yet to be included in the conservation lands network. For example, in the San Francisco Bay Area nonprofit organizations (NGOs), special districts, cities, and counties focused on grasslands, and federal agencies focused on coastal salt marsh, while all agencies acquired large extents of agricultural land. From this quick exploration of the state’s conservation history, it can be argued that a great deal of the state’s most spectacular and biodiverse areas have been conserved. Some of them were protected as early as the late nineteenth century, and early ideas from California about land conservation have been exported to varying extents to other parts of the state, nation, and world.

Not dated 1800–1809 1840–1849 1850–1859 1860–1869 1870–1879 1880–1889 1890–1899 1900–1909 1910–1919 1920–1929 1930–1939 1940–1949 1950–1959 1960–1969 1970–1979 1980–1989 1990–1999 2000–2009 2010–2019 Unknown

B. Timeline of NGOs San Diego Area

0

50

100 km

B O X 4 0 .1 F I G U R E 5  Timeline of special districts in the San Francisco (A) area and nongovernmental organizations in the San Diego (B) area. Source: Santos et al. 2014a.

and often too weak. There are, as we explain in this chapter, deep conflicting interests in funding allocation and priorities in the state. The following discussion reviews resource conservation issues in the state, the history of urban growth and land use regulation since the postwar period, and the rise and evolution of targeted conservation tools since the 1990s. The chapter addresses continued and new conservation challenges and innovative ideas and tools for conservation that emerge from the state’s history of land use planning and attempts to best preserve the state’s natural heritage.

Evolution of Land Use Planning and Policy in California Land use (and its planning) is at the heart of conservation of California’s magnificent heritage. As introduced earlier, land use planning stems from the need to plan for the compet-

ing interests for land. For greater conservation to occur into the future, more land must be set aside from other uses; more often than not, this requires its purchase. Next we review the historical process through which that came to be the case. We start with the Progressive Era (pre-1940s) when pro-development policies, such as the attribution of land use regulatory power to local entities, created the building blocks of California’s land use policy and established a legacy carried on until the twenty-first century. We then review the attempts to regulate urban growth when the boom of post–​World War II, progrowth policies began to be contested as impacts of local land use and planning “power” led to suburban expansion onto agricultural lands, creating inefficient and often segregated urban land use patterns. By the late 1970s a major change in revenue policies for local governments greatly affected their fiscal capacity, threatening the future of conservation land acquisition and adding a contradictory tension. While urban growth was Land Use Regulation for Resource Conservation   907

TA B LE 4 0 .1

Open space area in urban California counties

County

County area (km2)

Open space area (km2)

Area of county in open space (%)

San Diego

10,973.57

5,286.147

48.17

Los Angeles

10,242.41

3,375.847

32.96

San Mateo

1,430.55

435.12

30.42

Santa Clara

3,378.21

811.63

24.03

Alameda

2,126.87

469.39

22.07

Sacramento

2,580.21

225.94

8.76

277.02

22.12

7.98

San Francisco

Source: Santos et al. 2014a.

contested, a citizen-passed ballot initiative, Proposition 13, greatly decreased the property tax rate. This dramatically reduced city budgets and pushed land use zoning choices by cities and counties away from housing toward sales tax–​generating land uses. Today the property tax rate can only be changed by a supermajority vote, which means that at least two-thirds of the state population would have to approve the change. This has led to innovative solutions to raise funds for conservation land acquisition. A diversification of the portfolio of funding streams has occurred, including the participation of nongovernmental organizations (NGOs) and use of private philanthropic funds. Newer funding streams are also coming from impact mitigation banks and from carbon credits, but these are just emerging and little is known about how effective they will be. This history of land use policies and their cascading effects sets the stage for where California is today (Table 40.2).

Progressive Era Legacy: Pre-1940s A century-long series of efforts has occurred to create effective institutions to help guide metropolitan growth and development in California (Schrag 1998, Pincetl 1999b, Barbour 2002). Such institutions were proposed to operate at a regional or metropolitan scale, regulating where development could occur, infrastructure (like roads) would be built, and so forth. However, metropolitan-level approaches were pursued early and then consistently rejected in favor of local jurisdictional independence for cities and counties. Beginning in 1914, state legislation authorized charter cities to make and enforce all laws and regulations pertaining to municipal affairs, city-by-city and county-by-county (Silva and Barbour 1999), meaning that the state government could not tell cities what to do. Charter cities are cities that have developed a kind of constitution (a charter) that details its governmental arrangements, including the types of agencies and departments of the city, their roles, the role of the mayor, whether the mayor is elected or appointed, and other rules. Cities and counties have the authority to institute local control over land use, developing individual land use plans, zoning regulations, and so forth. Such local autonomy remains strong today. Local jurisdictional independence means the state 908   Policy and Stewardship

government has no authority over local land use decisions, including, in our case, preserving land for conservation. In order to understand the pressures facing the state’s ecosystems, it is important to review the patterns and governance structures of urban growth—​determined at the local level, as just explained—​in the state (see also Chapter 41, “Stewardship, Conservation, and Restoration in the Context of Environmental Change,” for additional pressures on the state’s ecosystems). Many of these patterns have not been significantly altered since the Progressive Era that set up the state government’s architecture in the period of 1912–​1918. This included local governmental autonomy in determining land uses and services (such as police and fire) and establishing elected representation in the form of city councils, county boards of supervisors, and planning and other representative bodies. General Plans for land use of cities and counties were required in 1937 by the state legislature—​though for charter cities they were voluntary—​and are documents that essentially establish the framework for directing the allowed uses of land as residential, industrial, commerce, or for institutions like schools and hospitals. General Plans guide where and what kind of development takes place over time. These documents are important legally for cities and counties and are the documents of reference when developments are proposed. They often can be amended to reflect changes in thinking by elected officials about land uses over time. For example, a development might be proposed that seems useful or appropriate, but the existing General Plan might show that land as Open Space. The local elected officials will then hold hearings and debate a change in the General Plan to accommodate that new proposal (see Fulton and Shigley 2012). These changes can be contentious and are the subject of democratic processes of decision making. With continued high urban population growth the state’s cities grew, and new cities were created and General Plans developed and revised. In the twentieth century, a time of seemingly abundant land and resources, inexpensive fossil fuels and building materials, land development was extensive rather than compact. These patterns were facilitated by urban and state policies including uniform subdivision regulations in 1929 (Fogelson 1967, Dear 1996, Hise 1997, Pincetl 1999b). Such subdivision regulations set out suggested patterns for subdivisions, including numbers of houses per area, sizes of roads, and other infrastructure patterns that were based on a perception of abundance. General Plans struggled to keep up with the pace of development, often overwhelmed by the tenacity of local developers and the proliferation of new municipalities, each with its own planning authority (Pincetl 1999b). California’s divestment of responsibility to localities during the Progressive Era continues to shape land use. Local governments base their ability to zone land and conduct business on “police power” as set forth in the California Constitution. This is an elastic power allowing cities and counties to broadly tailor policies and regulations to suit their community needs (California Constitution Article XI Section 7). Because each city and county controls its own land uses and development processes, governance is fragmented among many cities, none of which must coordinate with the others even when they are adjacent and when infrastructure deployment (like roads or sewers) would be more efficient through collaboration. While they often do collaborate, it is not required. Beyond local ­control over land use, land in the U.S. is parcelized, as parcels

TA B L E 4 0 . 2 Timeline of California legislation and regulation affecting open space, land use, and conservation

Year

Regulation

Purpose

Accomplishments

Unintended consequences

1914

Charter City Regulation

Codifies city control over municipal affairs

Regulate over where development could occur, infrastructure would be built, etc.

Lack of regional coordination for land use planning, mixed messages

1929

Uniform subdivision regulations

Codifies how subdivisions are to be defined

Extensive land development

Fragmentation of the land, encroachment on agriculture and natural ecosystems

1935

Central Valley Project

Water diversion project

Distribute water to agricultural areas in the Central Valley

Water rights, water dependencies

1937

General Plan

Comprehensive plan needs to be authorized for cities

Comprehensive plants that are guiding documents for cities and unincorporated land in counties that set out a general vision for the land in the jurisdiction of the city or county, and how it should be utilized

Charter cities receive autonomy from the requirement to adopt comprehensive plans

1947

State Highway Act

Determines regulations, plan and funds for the development of state highways

Growth of the road network

Exurban development towards agricultural and natural areas

1956

Federal Highway Act

Determines regulations, plan and funds for the development of federal highways

1956

Department of Water Resources

To manage state agencies over water use

Manages many water rights and water quality agencies

1957

California Water Plan

Determines who, where, and how much water can be used

New set of water projects including damming the Feather River and the State Water Project

1958

Bay Area Greenbelt Alliance

Regional planning and control growth through densification

Conservation of agriculture and open space

1959

State Planning Office

Centralized planning for land use

Reviews planning decisions; attempts to regional coordination of plans

1962

Federal Highway Act

Determines regulations, plan and funds for the development of federal highways

1962

California Tomorrow

NGO responsible for assessing federal and state actions over natural resources

1963

Local Area Formation Commission Act

County-level entity that reviews proposals to incorporate new cities, annex new territory to existing cities and create new special districts

Exurban development towards agricultural and natural areas

Overallocated existing available water resources; contributed to urban growth and the expansion of agriculture in the Central Valley

Advisory power

Exurban development towards agricultural and natural areas Proposals to reform state government and creation of regional government; impacts of sprawl on agricultural and natural lands No incentive to dissolve poorly functioning governments or take regional issues into consideration

(continued)

TA B L E 4 0 . 2 (continued)

Year

Regulation

Purpose

Accomplishments

Unintended consequences

1965

Bay Area Conservation and Development Commission

To overview development proposals towards a regional development plan

Stop Bay Infill projects; control development around the San Francisco Bay Area

1965

Williamson Act

Conservation of agricultural land

Allows farmers and ranchers to qualify for lower property taxes by keeping land as agriculture for ten years

Did not provide permanent protection because a termination fee can be used

1965

Quimby Act

Open space is required

Requires developers to set aside land as open space

Not a proactive planning tool or financial instrument thus cannot meet the open space needs; only helps new areas

1970

Update in General Plans

Requires open space elements

Open space elements are integrated in the general plans

OPR has little regulatory authority to impose them

1970

CEQA

California Environmental Quality Act

Requires state and local agencies to follow a protocol of analysis and public disclosure of environmental impacts of proposed projects and identify mitigation measures

1971

Update in General Plans

Become required for cities and counties

Require the inclusion of separate chapters—or elements—on transportation, open space, zoning, noise, housing, etc.; general plans are reviewed by the California State Office of Planning and Research (OPR)

Localities are still in charge of writing the general plans; there are no sanctions if the elements are poorly written, inconsistent, or out of date; the state has no substantive requirements for the plan’s elements

1972

California Wild and Scenic Rivers Act

Protection of rivers of extreme scenic beauty

No commitment from the state to provide water flows for maintenance of ecological processes

1972

Coastal Protection Act

Protection of California’s coast

1972

Coastal Commission

Protection of California’s coast

1972

Golden Gate National Recreation Area

Protection of San Francisco’s park and surroundings

Prior was land from the Department of Defense

1973

Federal Endangered Species Act

Protection of Endangered Species and their habitat

Strong regulatory power; defines take

1978

Proposition 13

Property tax reduction initiative

Cuts budgets for local governments dramatically

1978

Santa Monica National Recreation Area

Protection of the coast from development

Growth management efforts were hampered by the need for revenues; no funding for open space acquisition available

TA B L E 4 0 . 2 (continued)

Year

Regulation

Purpose

Accomplishments

Unintended consequences

1978

AB857 Urban Strategy

Infill versus further growth

Aimed to curb sprawl

No enforcement

1979

Proposition 4

Ceiling on state and local government expenditures

Further reduces local funds

Further constrains habitat acquisition

1982

Federal HCPs added to ESA

Habitat Conservation Plans

1988

Proposition 70

Specific park and recreation area projects

Sixty new specified projects were approved; sets the new era of voters approving specific and identified activities; voter trust, role of NGOs

No long-term funding

1991

NCCP

Natural Communities Conservation Planning

Voluntary participation necessary

Works best for large scale developers and large land areas

1996

Proposition 62 and 268

Control on local government spending

Extend supermajority rule (two-thirds) to all types of assessments, fees, or taxes used by local government

2000

AB 1427

Cortese Knox Hertzberg Local Government Reauthorization Act

Annexations must account for the protection of prime farmland

Procedural, no sanctions or consequences

2001

AB 857

Farmland Protection Act

Promote infill, protect environmental and agricultural resources, efficient development patterns

Same as above

2006

AB32

Global Warming Solutions Act

Set the new 2020 greenhouse gas emissions reduction goal

2007

SB 97

CEQA and Greenhouse Gas Emissions

Required OPR to develop, and the Natural Resources Agency to adopt, amendments to the CEQA Guidelines addressing the analysis and mitigation of greenhouse gas emissions

2008

SB 375

Sustainable Communities and Climate Protection Act

Unfunded mandate of the federal government

are what can be bought and sold. While this may seem like an obvious division of land, as we are accustomed to land regulated and on the market in parcels, this scale rarely is commensurate with ecosystems or habitats for ecosystem management and conservation purposes. Progressive Era land use and government frameworks set the stage for the rest of the century. At the same time, this new era brought increased public awareness and demands to curb development and preserve the natural resources of the state, including ecosystems and rivers as well as agricultural lands.

Attempts to Regulate Urban Growth: 1940s–​1980 Federal investments in World War II military industries greatly contributed to urban growth from the 1940s through

Procedural

Does not change land use law and local control over local land use

the 1950s as jobs were plentiful, attracting people to the state (McWilliams 1979). War and postwar industries, including the aerospace and automobile industries, made Los Angeles a manufacturing and industrial powerhouse. Mass building of Los Angeles suburbs like Winsor Hills, Westside Village, Toluca Lake, and Westchester was accompanied by new types of retail outlets such as supermarkets and malls, creating land use patterns still familiar today that are land-intensive and automobile-dependent. This pattern was replicated across the state (Hise 1997). Between 1950 and 1960 one hundred new California cities were incorporated in the state, with most of the growth located in suburban communities (Barbour 2002). Highways were also being built, launched and funded with the passage of the 1947 State Highway Act and the 1956 and 1962 Federal Highway Acts. These roads played critical roles in supporting suburban growth away from city centers Land Use Regulation for Resource Conservation   911

(Gregor 1957, Nash 1972). Federal subsidies aimed at supporting economic growth and modernizing the nation greatly assisted suburban expansion, an expansion that transformed agricultural lands, open spaces, and ecosystems. Post–​World War II urban growth also stressed the state’s existing water infrastructure and led to changes in it, affecting watersheds and groundwater resources. While we often think of conservation as the setting aside of lands for ecosystem preservation, water diversion and extraction also has significant impacts on ecosystem health. Water diversion and extraction from far-flung areas of the state has enabled the state’s urban growth to occur and agriculture to thrive (Reisner 1986, Hundley 1992). To meet new demand, water management entities proliferated (Pincetl 1999b). By the mid-1950s California had at least 165 irrigation districts, 69 county water districts, 55 reclamation districts, 39 water districts, 35 county water-works districts, 19 municipal water districts, 1,460 mutual water corporations, 456 commercial water companies, and 207 municipal water operations. In response, in 1956 a single state agency, the Department of Water Resources (DWR), was created by a special session of the legislature to bring together all of the state-level water agencies, and an appointed state water board was created. DWR was not given authority, however, over the myriad water management entities that had been created previously or over any subsequent water entities. In addition, groundwater pumping was excluded from regulatory purview and remains unregulated (pumping rate reporting has recently been required). A few urban groundwater basins are “adjudicated” or managed for safe yield, but this is recent and exceptional (see Chapter 38, “Agriculture”). In 1957, DWR issued a California Water Plan to call for a new set of water projects, including the damming of the Feather River and a new canal to bring water to the San Joaquin Valley conveyed through the San Francisco Bay-Delta. This ushered in the State Water Project (SWP), narrowly passed by the voters. The SWP has strongly contributed to urban growth in southern California and to expansion of agriculture in the Central Valley onto lands that had not previously been cultivated and are considered marginal. This resulted in severe impacts on the health of the Bay Delta as well as on the lands and groundwater of the west side of the San Joaquin Valley (Taylor 1975, Pincetl 1999a). By 1962, California had more than four thousand public and private water-related corporations, of which more than three thousand were concerned with some phase of water distribution while the rest were dedicated to flood control, drainage, and reclamation. This fragmentation of water responsibilities was not accompanied by any accounting of groundwater pumping, water sales, or monitoring of discharges. The State Water Rights Board existed at this time, but the dominant legal regime of “reasonable and beneficial use” limited its authority to incorporate public trust values in decisions. This situation since has proven difficult to remedy and is an important issue for the preservation of headwater ecosystems (ecosystems where water resources originate) and for determining how much water exists for ecosystem functions in the state. Unfortunately, current and future estimates of water volumes show that the State Water Project and the 1935 Central Valley Project (constructed earlier to provide farmers water in the Central Valley) lack adequate water to back up promised water contract deliveries (Table 40.3). Both the State Water Project and the Central Valley Project have allocated water contract rights that are a form of prop912   Policy and Stewardship

TA B LE 4 0 . 3 River flow and water right allocations in three major California river basins

Annual flows km3(MAF)

Water rights km3(MAF)D

Ratio

SacramentoA

26.6 (21.6)

148.6 (120.5)

5.58

San Joaquin

7.6 (6.2)

40.3 (32.7)

5.28

1.6 (1.283)

10.8 (8.725)

6.7

River basin

Trinity

C

B

Source: Adapted from California Water Impact Network: http://www .c-win.org/webfm_send/270, downloaded 3/4/12 MAF = million acre-feet. A. Includes the Sacramento, Trinity, Feather, Yuba, Bear, and American Rivers. B. Includes the San Joaquin, Merced, Tuolumne, and Stanislaus Rivers. C. Trinity River at Lewiston. D. Nonconsumptive hydropower rights are not included in this analysis.

erty right—​a usufructory right (right to use). The shortage of committed water—​even in average rainfall years—​means that the rights in contract could be breached, leading to potential lawsuits and perhaps compensation requirements. The long-term impacts of these projects are already being felt. In headwaters with significant water diversions for agricultural and urban use, rivers and streams no longer fully support fisheries and local fauna and flora. Lack of groundwater pumping regulations means heavy pumping, especially in drought years, is leading to subsidence and depletion, especially in the Central Valley. All in all, between the state and federal water projects and the limited authority of the Water Rights Board, not only are the state’s water resources severely overallocated, but groundwater resources are also poorly protected. Still, in recognition of the importance of the state’s natural resources, the 1972 California Wild and Scenic Rivers Act was passed to preserve the few, nondiverted, undammed remaining rivers possessing extraordinary scenic, recreation, fishery, or wildlife values. The act required the state to limit activities that might threaten flow, such as dam building or other diversions, on these wild rivers. Stress on water resources has polarized debate between agricultural users and those advocating for ecosystem water as well as between urban and agricultural water uses. Fragmented water management in all parts of the state, absence of water meters at residences in cities in the Central Valley, lack of central reporting of water use in urban areas, and weak reporting from agriculture all make it nearly impossible to accurately account for water use by different sectors. Only recently are farmers required to report groundwater pumping amounts. With growing concerns about climate impacts on the state’s water resources and recurring, serious drought episodes, such lack of data hampers better water management and exacerbates conflict among users. There have been efforts to better coordinate urban growth and water use in the state. Acknowledging these efforts and understanding why they were not implemented can inform future such efforts. The postwar growth boom spurred concern in the early 1960s from state policy makers and Governor Pat Brown about the effects of added fragmentation at the local level caused by increasing numbers of special districts formed to service urban growth (see Box Figure 5a) and the creation of new cities. The state legislature created a state Planning Office in 1959, located in the governor’s office, and

in 1960, Governor Brown appointed the Coordinating Council on Urban Policy to develop proposals to improve planning and coordination of growth. This emerged from the perceived need to prevent competition among local governments for new development and to mitigate adverse regional effects resulting from lack of development coordination. Adverse impacts included poor regional road integration, potentially redundant water agencies, duplication of services such as shopping malls authorized by individual cities and resulting competition between malls that undercut their financial viability, and competition for land uses that generate high tax revenues. The Coordinating Council’s report urged the creation of one multipurpose district in each of the state’s metropolitan areas to deal with regional issues such as air pollution, water supply, sewage, parks, and more. Today, regions have limited districts, including air quality management districts and flood control districts, though they are not elected and lack some of the initial far-reaching vision of the coordinating Council’s efforts. The council also suggested that incorporation, annexation, and the creation of special districts—​such as the myriad water districts mentioned earlier—​be coordinated and regulated at the state level to ensure more equitable growth throughout the state and better planning. Such an approach would have centralized decision making around creating new cities (incorporation), adding new land to a city (annexation), and the creation of new single-purpose agencies to provide a service (for example, a water district, a school district, a sanitation district). These proposals were opposed by cities and counties concerned about maintaining local authority over all aspects of urban growth and management. In 1963 the state legislature instead created watered-down, county-level local agency formation commissions (LAFCOs) that consist of five appointed commissioners empowered to approve or disapprove any petition for incorporation, special-district formation or dissolution, or annexation in each county. This was seen as a way to prevent some of the worst proposals—​for example, for new cities or districts that would not be able to generate sufficient taxes or fees to support necessary services, or for new cities too far away from existing infrastructure to be economically viable (Martin and Wagner 1978, Lewis 2000) One exception to the local-government rejection of coordinated regional planning was the creation in 1965 of the Bay Area Conservation and Development Commission (BCDC), a singe-purpose regional, land use planning entity. In this specific case the state legislature took initiative and designated a new entity to regulate and supervise land use and development around the San Francisco Bay, going against local government opposition (Dowall 1984). A strong citizen movement concerned about development impacts on the health of San Francisco Bay, including infill of the Bay to facilitate land creation for further development, had excellent relations with state legislative leaders and were able to convince the legislature to act. With continued urbanization pressure, additional grassroots movements to curb growth and to protect natural resources started to mobilize across the state in the 1960s and into the 1970s. The Sierra Club defeated an attempt by the state Highway Commission to improve Highway 101 by cutting through portions of the Prairie Creek Redwoods and Jedediah Smith Redwoods State Parks. Projects proposed by the Highway Commission often targeted state park lands for

transportation projects, as the land belonged to the state. The alternative was purchase of privately held lands. The many specific struggles are too numerous to enumerate here, but it is notable how conservation and land use issues became the subject of innovative proposals. One of the most significant, prescient organizations of the time leading these efforts and developing proposals to change California’s land use governance was California Tomorrow. In 1962 this nonprofit, educational institution began a decade-long effort of developing innovative proposals to reform state government and create regional government to manage growth. California Tomorrow published a monthly magazine; in 1967 it wrote: “Regional governments have not been created, in part because the white majority doesn’t want to face the full range of economic and social problems of the regional cities where, in fact, they live.” In essence, California Tomorrow pointed out the many challenges of suburbanization for good planning and also called out the role of white flight from inner-city neighborhoods that fueled the incorporation of suburban cities. California Tomorrow pointed out how sprawled urbanization patterns impacted agricultural and natural lands of the state and had negative effects on the fiscal health of older cities. It suggested infill development (using the unbuilt land in already urbanized areas) as an alternative, as well as curbs on incorporation despite the creation of the LAFCOs and their intended mission. Today these ideas are echoed in policies to develop more compact, transitfriendly cities to reduce greenhouse gas emissions. The Bay Area saw the rise of the Greenbelt Alliance in 1958 (then called People for Open Space), also organized around regional planning but with a focus on the conservation of both agricultural and open spaces and with the aim to control growth and encourage densification. This was a sophisticated approach that we can recognize today: to place further urbanization where it already exists and to maintain other lands for their ecosystem services, whether for food production or conservation. Other nongovernmental organizations that emerged during this period included the Planning and Conservation League (PCL) established in 1965 and the Trust for Public Land (TPL) created in 1972. TPL had been created in response to a development planned for the Marin Headlands north of San Francisco (TPL 2000). TPL pioneered the development of private land trusts that retain land in permanent open space and set a model for the rest of the country (see Box Figure 5b). In addition, the federal Golden Gate National Recreation Area (GGNRA) was created in 1972 partly as a result of the land conservation advocacy of the aforementioned nonprofit organizations but also reflecting the vigorous grassroots efforts to conserve important open space lands endangered by encroaching urbanization. In southern California citizen efforts led to the creation of the federal Santa Monica Mountains National Recreation Area (1978) in an attempt to preserve the unique ecosystem, one of the only east-west transverse mountains range in the U.S. The state legislature created additional opportunities for local governments to create open spaces, employing incentive-based approaches rather than regulatory requirements. The 1965 state legislature passed the Quimby Act, for example, which allowed local governments to require developers to set aside a portion of new subdivisions as parks or open space or to pay fees for park land acquisition and maintenance. The Quimby Act preserves open space as a condition of development and thus is not a proactive planning tool and financing instrument for open-space protection. In 1970 the legislature Land Use Regulation for Resource Conservation   913

also required open-space elements in city and county General Plans, but the numbers, sizes, and locations of these open spaces are at the discretion of cities and counties. In 1965 the state legislature also passed the California Land Conservation Act (the Williamson Act). This was once again an incentive program, and it reflected the legislature’s concern with urban growth impacts on the state’s agricultural lands. The Williamson Act allowed farmers and ranchers to qualify for lower property tax rates if they entered into contracts to keep their lands in agriculture for a minimum of ten years, and it provided counties financial compensation from the state for lost revenue. At the same time, the Williamson Act did not provide permanent agricultural land protection, as the contract could be abrogated by the property owner through payment of a termination fee (Sokolow 1990). Thus neither of these well-intentioned acts nor the open-space General Plan requirement met the need to protect open space and ecosystems through strategic planning and strong mechanisms for land conservation. Rather, they merely enabled the possibility of conservation, whether of agricultural or park lands. Today, counties are not compensated by the state for lost Williamson Act revenues. The 1969 Santa Barbara oil spill had a jarring effect on the state, prompting a number of legislative initiatives including the passage of the California Environmental Quality Act (CEQA) in 1970. Modeled on the federal National Environmental Policy Act, it requires state and local agencies within California to follow a protocol of analysis and public disclosure of environmental impacts of proposed projects and to adopt feasible measures for mitigating those impacts. It is a mandatory part of every California state and local agency process. Its passage was prompted by increased land development on the urban fringe and by the belief that greater transparency of impacts through their documentation would enable citizens to participate and demand better land use planning to protect critical resources (Pincetl 1999b). By the mid- to late 1960s the lack of integrated regional transportation infrastructure across metropolitan areas in the U.S. prompted Congress to incentivize the creation of metropolitan planning agencies (MPOs) or councils of government (COGs). Congress did so by offering transportation infrastructure grants to the regions that created these organizations—​another example of an incentive-based approach. California COGs were created in Los Angeles (the Southern California Council of Governments, SCAG), the San Francisco Bay Area (the Association of Bay Area Governments, ABAG), Sacramento, San Diego, and eventually the Central Valley. These voluntary organizations consist of local cities and counties, coordinate and collaborate across jurisdictions to plan and implement regional transportation projects, and can therefore qualify for federal funds to build them. In the early 1970s another regional governance organization also came about due to congressional law making—​a ir quality management districts. In California these were created by the state to implement the Clean Air Act at the air basin scale. Other regional governments, such as flood control districts, tend to be organized around single issues that might transcend jurisdictional boundaries but do not serve multiple purposes, certainly short of Pat Brown Coordinating Council proposals. Some have regulatory authority, such as air quality management districts; others are based on voluntary participation. While their actions affect ecosystems and environmental quality and must comply with the California Environmental Quality Act, none can facilitate land conser914   Policy and Stewardship

vation and thus have only indirect ability to protect ecosystem health. In 1971 the state legislature also required all local governments to develop, adopt, and follow General Plans—​which previously existed but were advisory for city planning. The General Plan was to be a city and county’s premier policy document and is now often referred to as the local constitution or blueprint for development. In 1990 the California Supreme Court held that the General Plan was the “constitution for all future development” (Lesher Communications, Inc. v. City of Walnut Creek, 52 Cal. 3d 531, 540 [1990]). After 1971 the government code was expanded to require that all land use approvals be consistent with the adopted General Plan and that all its elements be consistent with one another and the plan overall. At the time, this was seen as a great improvement in local planning processes. All General Plans and their elements were to be reviewed for consistency by the Office of Planning and Research (OPR) and revised every ten years. Unfortunately, the OPR has little regulatory authority to reject General Plans or demand revisions. It is often understaffed and unable to review plans in a timely manner, and no real sanctions exist for poor compliance. In another major, citizen-led change inspired by concerns over urban growth, state voters created the California Coastal Commission in 1972 by ballot initiative. The commission, an entity appointed by the governor, was given regulatory authority to protect California’s coast from overdevelopment. This was one of the strongest measures to emerge from a period described by Press (2002) as the state’s first slow-growth era, and a clear reaction to the increased pace of urbanization that was occurring. The 1970s were a period of active citizen initiatives to preserve land and environmental quality, both at the state level, as with the Coastal Commission, and at the local level with over one hundred ballot initiatives aimed to restrict or slow growth throughout the state. It was widely perceived that—​for different reasons in different parts of the state—​t he pace of growth was having negative effects. Reasons included concern about preserving agricultural land, impacts of sprawl on the environment, increased length of commuting, and desire to maintain property values. Citizens created ballot initiatives to attempt to change land-use designations to curb further sprawl. Some succeeded in passing while others did not, but overall they did little to curb the pace of urban growth fueled by population and economic growth. With the establishment of the suburban land use pattern, most new urbanization was landintensive and car-dependent, intruding further and further into the countryside. Unease about land development made increasingly profitable by the inflation of property values and leading to higher and higher property taxes led to another outcome: the passage in 1978 by state voters of Proposition 13, the property tax reduction initiative. Proposition 13 lowered the taxation rate for property and required a two-thirds majority vote for any new taxes. Local government revenues plummeted from about $10 billion just prior to the passage of the measure to approximately half of that shortly thereafter. In 1979, Proposition 4 furthermore placed a ceiling on both state and local government expenditures (Press 2002). The impacts of Proposition 13 on local government budgets and land use planning were immense. Growth management efforts were hampered by the need for revenues to purchase lands for conservation. Due to the cut in revenues from property taxes, cities and counties went the opposite direction, adopting land use poli-

cies most likely to replenish their depleted budgets: zoning for high sales tax businesses and annexing farmland and open space to add to developable land resources so as to increase their tax base (Pincetl 2004 and 2006, Wolch et al. 2004). In summary, the 1940s through the 1970s were complex and sometimes contradictory years. The era started with prodevelopment policies aided by the Federal and State Highway Acts that increased accessibility and further promoted development. But this huge growth and its impacts on the state’s scenic resources and agricultural lands, as well as on quality of life, led to ballot initiatives to guide and curb growth and to protect the environment. These efforts in turn encountered the entrenched interest of cities and counties in maintaining control over land use and the need to generate revenue to fund city and county services, revenues derived from developed land. Growth pressures ultimately led to the creation of some regional governance entities, including councils of government, air quality management districts, and others. At the same time, desire for the protection of habitat and species led to federal and state Endangered Species Acts, adding to the complexity of land use regulation. With the advent of property tax reductions, local governments found themselves increasingly squeezed among regulatory requirements, frameworks and voter sentiments. The next phase of California’s regulatory development was in part an attempt to overcome some of these contradictions in policy direction.

The 1980s and 1990s This abrupt shift in attitude and relationship towards government and governmental resources by the public—​t hat of reducing taxes, and radically reducing revenues while also wanting environmental protections—​led to new conservation strategies and approaches as the effects of Proposition 13’s tax reductions and of continued growth played out on the landscape. Ballot measures for open space funding using bonds became popular, and these measures had habitat and wildlife protection goals as their focus, earmarking substantial funds for acquisition of valuable ecosystems. Changes in intergovernmental relations and in the relations between government and civil society emerged as well (Press 2002). With the decline in local governments’ ability to fund land preservation with tax dollars, new mechanisms emerged. One such new strategy was developed successfully by the Planning and Conservation League (PCL) under the leadership of Gerald Meral (from 1983 to 2003). The PCL developed a set of inventive campaign funding tactics and ballot initiatives to finance the acquisition and development of park and recreation areas. Meral’s first successful measure (Proposition 70) in 1988 included sixty individually defined projects, each with a specified amount, in every corner of the state (Pincetl 2003). Never before had specific projects been prescribed in a ballot initiative, and each one was chosen with an eye to maximizing local voter appeal and was based on negotiations for support with local environmental groups and others ranging from homeowners to local governments (Schrag 1998, in Pincetl 2003). This marked a departure from having the state legislature determine where park or open space bond funds would be spent. It was part of the shift toward appealing to voters based on pre-specified projects, and it reassured voters that funding would go to projects they preferred. What was also novel in this approach was that the unallocated, but prescribed, funds

could be used by NGOs for certain types of projects. While a smaller percentage of the bond funding, these funds allowed the precedent of local groups participating in creating parks of their choosing rather than having parks established by the state legislature or state parks department. Unfortunately, this funding did not come with maintenance monies, thus creating longer-term financing problems for local nonprofits or governmental agencies responsible for managing the new parks. Nongovernmental organizations and quasi-governmental agencies thus emerged from this period as new stewards of public spaces, including ecosystems. State conservancies funded by grants and donations were created, such as the Coachella Valley Mountains Conservancy in 1991. Statechartered, with an appointed board approved by the governor and with nominal state funding, these conservancies are now found across the state working to preserve critical ecosystems. More nonprofit organizations including local land trusts were created and stepped in to raise funds for land conservation. Empowered by the new conservation regime established by the PCL, they also applied for state bond funds to purchase land (see Box Figure 5), often turning them over to public agencies to manage. In parallel, voters passed Propositions 62 and 218 in 1996, extending the supermajority approval requirement of Proposition 13 to virtually all types of assessments, fees, or taxes used by local government. This further constricted local government’s flexibility to raise funds for necessary services outside of voter-approved bond propositions, let alone to directly purchase additional lands for conservation purposes. Open space preservation efforts have therefore since been further squeezed by lack of state funds, creating new and different relationships between state and local governments and nonprofit organizations. For example, private conservation easements emerged in this period (Merenlender et al. 2004). Conservation easements are entirely private agreements that have little or no state or local public governmental supervision or regulation. These, and other new land conservation approaches like state-created but independently funded land conservancies, are still relatively novel; their consequences for long-term land preservation, management, and the funding of both remain to be examined over time. In the big picture the historical contribution of federal agencies to California open space has been quite disproportionate when compared to other levels of government (see Box Figure 1b). This is an artifact of California having a lot of spectacular public lands that had not been claimed for private ownership in the nineteenth and early twentieth centuries. It also reflects the dilemma of the fundamental mechanism for conserving land once the state grew—​t he necessity for land to be purchased when state and local land use regulators are unable or unwilling to conserve important ecosystem values through regulation or are constrained by the need for revenues. The 1980s and 1990s started off with an abrupt reduction in allocated governmental resources for conservation land acquisition and for local government operations in general. Rather than stalling conservation, it led to the development of innovative funding streams for land acquisition and the emergence of conservation easements. Ballot measures and bonds earmarked substantial funds for land acquisitions, and for the first time individual projects were identified in ballot initiatives. This era also saw the blossoming of nongovernmental organization land purchases. These innovations set the stage for the twenty-first-century conservation. Land Use Regulation for Resource Conservation   915

Twenty-first Century: Continued Conservation Challenges Given the historical antagonism of localities to state land use regulation, increasing fiscal pressure on local governments, and the simultaneous persistence of support for land conservation, the state has developed reactive and voluntary approaches to meet conservation demands. These create tensions and mixed messages between local power and regional planning efforts, and mismatches between policy goals and instruments for their implementation.

Tensions between Local Authority and Regional Planning Land use designations are an important part of municipal powers (the powers of city and county governments), as land use determines the character of cities and counties, their revenues, and their attractiveness for investors and businesses. Local tax revenues depend upon the way land is zoned or allocated among uses—​residential, commercial, and industrial. Zoning land is one of the main responsibilities of local governments and the way they control what goes on in their cities. Elected officials can zone for medical uses, auto body repair shops, single-family dwellings or apartments, parks and open space, and so forth. Since retail sales generate a great deal of tax revenue, zoning for this use has become very popular, and cities and counties do not want to give up any authority over zoning of their lands. Cities prefer to avoid zoning land uses that do not generate tax revenues—​like parks—​or that require high levels of expensive public services. As Nicolaides and Wiese (2013) point out, zoning can perpetuate inequality among cities and suburbs. Cities, in contrast to suburbs, are often older and have less flexibility to change their land uses to attract land uses that generate higher revenues. They have historic infrastructure like existing buildings and roads that are complex and expensive to change. Suburbs, which are newer, have planned their land uses to generate revenue. In essence, “the places themselves help to create wealth and poverty. . . . Adding to these [fiscal] advantages, each is carefully controlled by land-use restrictions that freeze in place their landscapes of high-end homes and freeze out almost everyone else” (Nicolaides and Wiese 2013). This is because localities have no incentive to be inclusionary (that is, to include people and land uses that generate little revenue), as that might reduce the already-fragile revenue stream. This is why a requirement now exists of state review through the state Office of Planning and Research of the housing element of General Plans, but its enforcement is very weak. There have been numerous attempts to create regional governments with regulatory authority and to encourage greater collaboration among local governments. A regional government would transcend individual city governments in a region. For example, Los Angeles County has eighty-eight cities. The county can only govern its own lands—​the lands that are not in the eighty-eight cities. Thus in Los Angeles County each city runs itself as though it was an island and its decisions had no impact on its neighbors. Funding is not shared, infrastructure projects are rarely coordinated, and each city tries to increase its own revenues through zoning decisions at the expense of its neighbor. Common goods like parks do not generate income; at the regional scale little incentive exists 916   Policy and Stewardship

for a city to create them, as park land is withdrawn from the taxable pool and the city is responsible for maintenance of the park even if it is a regional asset or benefit. Existing regional “special districts” like air quality management and flood control have narrow, single-issue mandates and do not serve to coordinate regional decisions outside their purviews. Notwithstanding prior attempts to create regional government entities with land use authority, today only the Tahoe Regional Planning Agency (with bistate jurisdiction) has land use authority. Even this organization has many difficulties, as the Nevada side is much more pro-development than the California side. Nevada’s position is driven primarily by the desire for more revenues, and Nevada has threatened to withdraw from the compact over stricter land use requirements proposed by California. Meanwhile the water quality of the Lake Tahoe—​the issue the agency was created to address—​continues to deteriorate due to land use impacts. This has led to something of an impasse. Conservation needs and ecosystems do not observe political boundaries; they are regional and even statewide. One could argue that many urban regional issues are similar, such as transportation, housing, water supply and reuse, and air quality. Ecosystem conservation often calls for landscape or regional conservation, keeping parcels together to provide connectivity and enough space for ecosystems to achieve full functioning and to preserve watersheds (see Chapter 41, “Stewardship, Conservation, and Restoration in the Context of Environmental Change”). This generates a fundamental mismatch and contradiction between how land use is planned and regulated and the scales at which conservation needs to occur. In addition, parcel-oriented planning is inimical to orderly urban growth, as it is not really possible to develop large and proactive plans for cities when the process that allocates land is piecemeal. Nothing in current planning law recognizes this fundamental incapacity to treat a region more holistically, including its ecosystems and topography, with the exception of the Natural Communities Conservation Planning (NCCP) process discussed below. One could argue that contemporary, parcel-based planning is anachronistic for the needs of more sustainable cities and landscape preservation and conservation.

Mixed Messages to Localities, and No Additional Resources With the decline in property taxes, cities and counties became increasingly reliant on impact fees (charging new development a special surcharge to mitigate its impact, for example, on habitat) and alternative property assessments to finance public infrastructure (Matute and Pincetl 2013). Impact fees charged at the time a development is authorized can be used to mitigate impacts from developments, and cities and counties often encourage large developments to reap significant impact fees. Often these fees are used to build schools and local parks, projects necessary to satisfy the additional population growth. Another strategy to increase local revenues has been to zone more land for retail sales, especially for highpriced items such as automobiles or high-volume sales from big-box developments since sales tax was not been affected by Proposition 13. These land uses have generated still greater consumption of land, especially on the urban fringe since bigbox developments need big areas. This has been described by a number of observers as the “fiscalization of land use” (Mat-

ute and Pincetl 2013, Barbour 2002, Pincetl 1999b, Schrag 1998, Fulton 1997). The fiscalization of land use has been a known barrier to sustainable land use outcomes for decades. This issue cannot be overstated; simply put, the tax structure resulting from Proposition 13 favors sprawl over infill since large-format retail and auto malls that generate high tax revenue better fill local government coffers. Paradoxically, while retail development is considered fiscally positive, it often generates a net fiscal drain due to impacts such as traffic, costly ongoing maintenance of an expanded road system, and induced growth in the form of residential development (Pincetl 1999a). Other policies reinforce current land use patterns and are deeply embedded in how cities do their business, such as school funding formulas that reward suburban schools and lender’s aversion to new housing formulas (for example, for infill or mixed-use development). These too are well-known, but policy remedies have not been adopted into law. Localities are now also being asked to consider development and transportation impacts on greenhouse gases by Senate Bill 375, to include greenhouse gas (GHG) impacts in CEQA documents, and to create climate actions plans. However, within an unchanged regulatory context no sanctions occur if these new requirements are not met, and no changes have been made to General Plan guidelines that require GHG reduction measures. Individual-city General Plans need not be consistent with those of adjoining cities in the region. Finally, no new sources of revenue mitigate the pressures towards further development (and especially retail development) for revenue. Thus localities have not been given any new tools that might encourage greater infill development, relieving pressure on undeveloped agricultural land and impacts on natural habitats nearby and in the hinterlands.

Mismatches between Policy Goals and Legislative Tools California has long recognized the costs of sprawl; it has identified the need for specific land use outcomes including orderly and efficient land use, infill, farmland protection, and conservation of resource lands. But the fundamental way in which land (and water) is regulated has not changed, only the urgency of state priorities. With the advent of climate change, priorities have expanded to include renewable energy and land use that reduces greenhouse gas emissions. All of these concerns are embodied in adopted state law and policy. Examples include the 1978 Urban Strategy turned into state law under Assembly Bill 857 (Wiggins codified at Section 65041.1 of the Government Code) in 2001; Assembly Bill 32 (California Global Warming Solutions Act), 2006; and Senate Bill 375 (Sustainable Communities and Climate Protection Act), 2008—​none of which actually alters land use planning. Among the broad goals in the adopted the 1978 Urban Strategy developed by Governor Brown’s Office of Planning and Research during his first administration (second term) were curbing urban sprawl and directing new urban growth to existing cities and suburbs, revitalizing central cities and neighborhoods, and protecting resource lands, in alignment with the proposals put forward by California Tomorrow and others. This was before understanding had emerged about climate change and was simply a strategy to deal with the state’s urban growth. The Urban Strategy laid out a framework for how planning should be done; it had no legal requirements

for implementation. In 2001, Assembly Bill 857 (the Farmland Protection Act) codified these planning objectives by establishing three priorities that encourage all state agencies to (1) promote infill development within existing communities, (2) protect the state’s most valuable environmental and agricultural resources, and (3) encourage efficient development patterns overall. It required the governor, in conjunction with the governor’s budget, to submit annually to the state legislature a proposed five-year infrastructure plan. The plan would have to show how all proposed infrastructure expenditures were consistent with these state planning priorities. There is little evidence that Assembly Bill 857 is being implemented, with the small exception of the current Strategic Growth Council’s use of the act’s planning principles as a prerequisite for grant giving. The current Strategic Growth Council was created in 2008 as part of the state legislature’s package of incentives to help localities plan for climate impacts. The council leadership embraced Assembly Bill 857’s languishing principles as good guidelines for their own grant funding (to which funds are provided by the state legislature). In 2006 the state legislature passed and Governor Arnold Schwarzenegger signed Assembly Bill 32, the Global Warming Solutions Act of 2006, which set into law a 2020 greenhouse gas emissions reduction goal to return overall emissions to 1990 levels. This law directed the California Air Resources Board to begin developing discrete, early actions to reduce greenhouse gases while also preparing a scoping plan to identify how best to reach the 2020 limit. Strategies in the scoping plan include recommendations for local policies that promote infill and more compact urban growth to reduce vehicle miles traveled (VTM) and greenhouse gas emissions. Senate Bill 97 (CEQA and Greenhouse Gas Emissions), passed in 2007, expressly added greenhouse gas emissions analysis as part of the CEQA process. Senate Bill 97 required the state Office of Planning and Research to develop, and the Natural Resources Agency to adopt, amendments to the CEQA guidelines addressing the analysis and mitigation of greenhouse gas emissions. These are targets and analyses requirements that do not have explicit required implementation steps and can be met however is deemed appropriate. Building on Assembly Bill 32, Senate Bill 375 (2008) then linked reduction of greenhouse gas emissions to land use and regional transportation planning. While this statute is possibly the closest California has come to requiring regional planning and encourages urban and suburban infill, clustered development, mixed land uses, transit-oriented development to reduce GHGs (California Transportation Commission 2010), Senate Bill 375 does not change land use law and local control over local land use. Compliance is not mandatory. All of the aforementioned statutes are aimed at the same overarching policy direction and land use strategies—​c reating more compact development, achieving greater regional integration of land use, infrastructure, and transportation planning—​but the fundamental structures of local control over land use were not touched, nor were the guidelines for General Plan. No additional sanctions or rules were created, nor revenue streams, with the exception of a “sleeper” provision that creates legal claims and remedies for violation of housing policies. Unfortunately, as in previous attempts to regulate land use in the state, the gap between these state policy guidance regulations and local land use planning and actions was not filled. For example, Cordova Hills, a 1,100-hectare housing development not adjacent to any curLand Use Regulation for Resource Conservation   917

rent development and inconsistent with the recently adopted Sacramento regional plan and with the goals of SB 375, was approved by the Sacramento County Board of Supervisors as a measure to encourage economic activity and growth in March 2013. Local governments opposed regulatory requirements for the consistency of local plans with SB 375, and for good reason. They did not want to be constrained in land use decisions they consider their prerogative. As a result, SB 375 contains incentives rather than penalties that fall short of what is necessary to align state and local policy and state desired outcomes.

Innovations toward a Change Current and recent policy innovations have developed to try to improve conservation planning in lieu of strong state and regional planning (see also Chapter 41, “Stewardship, Conservation, and Restoration in the Context of Environmental Change”). They are speculative—​that is, not yet fully tried out—​complex, and highly technical. They are often predicated on large-scale transportation infrastructure (to accommodate assumed future growth), tying funding for habitat preservation and compliance with the Endangered Species Act with the construction of freeways to service future and further urbanization. These innovations demonstrate the direction of habitat conservation going forward—​fi nding opportunities for funding in continued urban growth, and using existing programs and infrastructure plans rather than regulatory reform or the creation of new funding streams.

Habitat Conservation Plans and Natural Communities Conservation Plans California, due to its unique geography, has more endemic species than any other state in the continental U.S. (see Chapter 11, “Biodiversity”). With urbanization pressures throughout the state, developments have thus been held up by the federal Endangered Species Act (ESA), passed in 1973 with the aim to preserve threatened and endangered species and their habitat. Development conflicts began to abound. Environmental organizations used the ESA to challenge development with some success, and developers found themselves stymied and slowed down by ESA challenges to development projects. ESA challenges were based on a species-by-species threat reflecting the law, and it became clear that policies based on protecting individual species would scarcely achieve the goal of preserving species as they relied on habitats that needed to be preserved as well. Habitat conservation plans (HCPs) were added to the ESA by Congress in 1982. Congress viewed HCPs as a win-win for imperiled species. HCPs took habitat into consideration and created incidental take permits (ITPs), which allowed—​for the first time—​the taking of a limited amount of habitat via permit. This take permit would be issued in exchange for a commitment to protect and manage other habitat areas, maintaining the species’ chances of overall recovery (Kostyack 2001). Kostyack (2001) describes: “Facing the possibility of significant development restrictions due to the ESA’s prohibition against taking of listed species and the possibility of liability for issuing permits in violation of this prohibition, local governments have negotiated with federal agencies to ensure that their development plans are consistent with ESA stan918   Policy and Stewardship

dards,” thus allowing development to go forward. However, habitat conservation is fundamentally an issue of how much habitat is required and where it must be located to ensure a species’ long-term viability. As HCPs were created to protect one species at time, California, under Governor Pete Wilson, created Multiple Species Habitat Conservation Planning (MSHCP) and the 1991 Natural Communities Conservation Planning (NCCP) process in an effort to encompass entire ecosystems and their processes. The NCCP process was predicated on volunteer landowner participation. It was launched to construct preserves for coastal sage scrub habitat in southern California as a way to enable large-scale developers in Orange County to proceed with development but to also preserve the coastal sage scrub habitat and its fauna. It was intended to bring all stakeholders to the table in order to set aside coherent, regional habitat preserves (Fulton 1997, Pincetl 1999a). NCCPs were endorsed by secretary of the interior at the time, Bruce Babbitt, as a legitimate implementation of the ESA that would allow the protection of endangered species but not halt development. At that point, some feared that the ESA would be challenged by hostile forces in Congress. If the NCCP could be shown to work, it would obviate the calls for reform or deauthorization of the ESA. The NCCP process is similar to the HCP process in that it allows the taking of habitat and individual species in exchange for habitat set-asides. It differs from the federal HCP in that it is a state-led program that is voluntary, locally initiated, potentially applicable at a county or regional scale, and initiated before the landscape becomes degraded and to protect ecosystems at the landscape scale while accommodating compatible development. It is a landscape-oriented approach that can potentially resolve habitat fragmentation while also ensuring development. NCCPs have been successful in creating conservation set-asides where there have been willing landowners and funding to purchase lands as the process is voluntary. The NCCP process was an important attempt to address the scale mismatch between land use policies and natural resource protection discussed earlier. Because land planning is done parcel by parcel, and habitat often needs to be protected at a much larger scale that often transcends several local political jurisdictions, implementing the NCCP opened up new conservation options to draw in participants at multiple scales and different political jurisdictions. Nongovernmental organizations’ land acquisitions (see Box Figure 4) post-1980s also resulted in large tracts of land in southern California being added to the conservation network (see Box Figures 1a, 4, and 5 ). Landscape-level planning provides a wide range of benefits, particularly when in the form of HCPs and NCCPs:

. It permanently protects large, interconnected, and bio. . .

logically rich blocks of habitat—​a strategy for conserving species and habitats far more effective than protecting small and isolated reserves. It increases efficiency and provides more certainty to project proponents seeking state and federal permits from wildlife agencies. It provides economic incentives and fair compensations to private landowners for permanently protecting natural resources on their lands. It establishes partnerships among local government agencies and state and federal wildlife agencies for the conservation of natural resources.

. It establishes partnerships between state conservation agencies and state infrastructure agencies to align conservation goals and implementation and to plan for California’s water, transportation, and energy needs, for example, while avoiding its most sensitive resources and reducing the further fragmentation of habitat.

General Plans: The Blueprint Process Not to be confused with local General Plans, the California Department of Transportation (CALTRANS) developed a Regional Blueprint program that preceded Senate Bill 375. Regional Blueprints came about as a result of increasing traffic congestion and the long-term inability of CALTRANS to build sufficient road infrastructure to address the state’s needs. Not only are funds insufficient, but growing evidence also shows that building more road capacity simply induces more traffic. Regional blueprints are collaborative planning processes that engage residents in articulating a vision for the long term of their region and assessing trade-offs. Regional visioning efforts used advanced Geographic Information Systems (GIS) scenario mapping to demonstrate the comparative impacts of different regional development scenarios. They involved numerous stakeholders who were challenged to allocate projected growth across each transportation region. The concept was that if people—​stakeholders—​were involved and learned of the difficult trade-offs by trying to make projected growth fit in an existing region, good policy would emerge from the ground up. These early blueprints inspired the Sustainable Communities Strategies (SCS) planning requirements under SB 375 and have become a touchstone for local general planning. However, the blueprints and SCS requirements have no implementation and enforcement mechanisms, and they focus on the existing built environment, leaving aside existing open space lands and their future (CALTRANS 2013). Few regions have exhibited ground-up results from these processes. Sustainable Communities Strategies under SB 375 require the state’s regional metropolitan planning organizations (MPOs) to develop plans for how to reduce vehicle miles traveled and better match housing locations to jobs. SB 375 does not provide the MPOs with any greater authority to require their members to implement the SCS, however, nor are there any penalties if the SCS is not abided by. Further, neither Regional Blueprints nor SCSs under SB 375 are required to be consistent with local General Plans. Consistency would, at a minimum, oblige localities to address how their land use decisions might affect what is outlined in the SCSs and/or Blueprints. But only the state legislature has the authority to require consistency through legislation, and it has not chosen thus far to exercise this authority. General Plan guidelines were last updated by the California Office of Planning and Research (OPR) in 2003. OPR has recently initiated an update with a focus on the policy outcomes of the state—​as articulated in AB 857 and the 1978 Urban Strategy for California—​and specific “metrics or measures” General Plans should meet to do so. Under the Brown administration, enforcing existing rules is seen as a way to address climate impacts and encourage better planning. Creating metrics or measures that General Plans should meet could provide a pathway to effective change in land use decision making.

Regional Advance Mitigation Planning (RAMP) Another recent innovation is the Regional Advance Mitigation Planning (RAMP) in 2010. This is being implemented with an early example in Transnet, San Diego’s sales tax for local transportation, and Measure M2, Orange County’s reauthorized sales tax measure. These are parallel and nearly invisible efforts (probably because of their novelty) that might provide a new framework for conservation. RAMP is an effort to develop a more comprehensive approach to mitigate biological resource impacts caused by major infrastructure projects. It allows for natural resources to be protected or restored as compensatory mitigation before projects are constructed, and often years in advance (Thorne et al. 2014). However, the fear of takings looms, and it remains to be seen if this approach will obviate that concern.

Sustainable Communities Strategies, Conservation, and Open Space Plans: The Greenprint As discussed earlier, the Endangered Species Act has been a driver of habitat conservation in the state. The Natural Communities Conservation Plan process created a more comprehensive, regional approach to conserve habitats based on science. Yet the NCCP does little to address the fiscal pressures on localities, and in fact NCCP success is often predicated on growth because land preservation is possible only through development fees that can fund land acquisition. Localities, given the dearth of tax revenue financing, use growth opportunities to leverage new infrastructure and to levy fees that go to paying for habitat conservation or other programs. While some localities might want to plan proactively, including for conservation, this is difficult where other cities continue to plan for growth and in a time of tax revenue scarcity. An emerging concept to creating sustainable communities is inclusion of Regional Greenprints (different from the Blueprints just discussed) as an integral part of local General Plans and Sustainable Community Strategies. While Greenprints are still conceptual, they would offer a new way to improve conservation planning by providing a process to map a region’s important open space for a full range of ecosystem services, including habitat, farmland, recreation, water resources, and more. Greenprints go beyond conservation plans because they are inclusive of all open space values. Since a region’s natural systems provide the basis for ecosystem services, planning for and integrating green infrastructure with the planning of hard infrastructures like freeways would serve to identify natural areas that should be protected or managed appropriately. This concept is not new; Ian McHarg (1970) laid out a compelling vision for this approach to planning and development. Today, with GIS tools, advances in conservation biology and interest in sustainability and resilience, more and better tools for protecting ecosystems are available. Conservation planning could provide a baseline from which land use plans are developed and could be linked to CO2 sequestration—​a new and popular idea. Requiring conservation priorities to be respected in land use and transportation planning would provide a number of important benefits. These include: 1. Promoting strategies that would reduce risk and project delays for infrastructure agencies and development, as plans would be based on better information Land Use Regulation for Resource Conservation   919







about landscapes, water resources, and working lands, to help guide decisions on infrastructure investment. 2. Driving mitigation (e.g., regional advanced mitigation) and conservation dollars to protect priority conservation areas. 3. Directing development to areas more suitable for development and avoiding hazards such as infill development. 4. Preserving lands that will sequester CO2. 5. Reducing conflicts by incorporating conservation planning into infrastructure and land use scenarios, providing an element of certainty for development and allowing for more systematic economic growth.

CO2 sequestration credits or land preservation, currently another land use conservation strategy under consideration, is associated with a number of uncertainties and caveats. These include the uncertain science regarding CO2 sequestration by specific lands in specific places in the state, including the state’s forests; whether carbon offsets for forestry or land use conservation represents “additional” sequestration; and long-term management of these lands and forces such as fire, floods, and other potentially destructive events. Further, many CO2 sequestration estimates are based on data from lands outside California and not previously degraded. The idea of linking Greenprints to CO2 sequestration is new and may take time to learn how to implement. It could also be funded by California’s new carbon market, helping relieve the pressure on development fees to fund conservation.

Future Challenges Other ongoing activities will also challenge habitat preservation. Paradoxically, concern about greenhouse gas emissions is fueling Governor Jerry Brown’s drive to have a highspeed rail line link southern California to the Bay Area and Sacramento. This project is highly contentious. Its route, as currently designed, goes through some of the state’s prime agricultural land and could support even more suburbanization in the already urbanizing Central Valley. More suburbanization will likely bring more air pollution, demand for more water, and impacts on the Sierra Nevada mountain ecosystems and water areas of origin, undermining GHG reductions. Without strong land use controls, this outcome is likely. Governor Brown is also pursuing more energy resources, including oil and natural gas fracking. “Fracking” is the fracturing of rock by a pressurized liquid, often water-based and supplemented by chemicals that increase pressure and keep fractures open such that the oil or gas can be pumped. A great deal of the state’s natural gas and petroleum resources are found in the Central Valley and compete for water, adding another demand to the state’s scarce water resources in addition to agriculture, cities, and places of origin. Fracking will also produce highly problematic wastewater, which will have to be disposed of, and could pose threats to fauna and flora as well as contribute to already contaminated groundwater. Fracking activities will increase air pollution, which in the Central Valley affects ecosystem functioning in the Sierras (see Chapter 7, “Atmospheric Chemistry”). As with many water extractions in the state, little accounting is taking place of the quantities of water being used for fracking or of current quantities of water used by the active petroleum extraction industry in the southern Joaquin Val920   Policy and Stewardship

ley. Another important potential area of energy exploitation is the Monterey Formation in the California Coast Ranges and Peninsular Ranges. The formation was first considered by the U.S. Energy Information Administration to have yield potential of approximately 15.5 billion barrels of oil. Though this estimate has been downgraded considerably, it still illustrates the pressure resources are under for multiple use. While recent legislation (SB 4, 2013) has been passed to regulate fracking in the state, it is considered by some as not strong enough. Protected Forest Service and BLM federal lands are not exempt from other energy projects in California. Large-scale renewable electricity generation projects in the high desert of the Mojave have been given the green light by the Department of the Interior despite documented impacts on the desert tortoise (Goperhus agassizii), among other species (see Chapter 11, “Biodiversity”). Addressing climate change policy through renewable energy development is therefore in tension with land conservation in some areas.

Climate Change Climate change is an important emerging factor affecting conservation (see Chapters 14, “Climate Change Impacts”; 41, “Stewardship, Conservation, and Restoration in the Context of Environmental Change”; and other chapters throughout this volume). Its impacts on California, which will likely include reduced snowpack, increased temperatures, sea level rise, and other less understood changes, have significant implications for the state’s future. As elsewhere in the world, changes in temperature affect fauna and flora dramatically, and accompanying changes in the hydrologic regime will exacerbate those impacts. Greater competition for water is probable among existing users and will also increase for reserved water for ecosystems and places of origin. Ironically, transportation is the single highest contributor to GHGs in the state, a result of the state’s still predominantly automobile-oriented urbanization process. This means that cities remain one of the biggest contributors to GHG emissions, affecting habitat in this way as well as competing for land. Increased temperatures will also aggravate air pollution, affecting human and ecosystem health. Places like southern California and the Central Valley will see much worse air quality, with increased ozone levels impairing ecosystem health in adjacent national forests and parks. Climate change is the culminating phenomenon resulting from human reliance on fossil energy, an energy source that is deeply intertwined in all aspects of daily life and the economy.

Where Do We Go from Here? California has been one of the most urbanized states in the nation since its early days. Local control over local land use is strongly protected by localities and local elites who defend their prerogatives to maintain the character and type of housing, and thus the population that can live in that locality. However, to improve the future of conservation of the state’s biodiversity and unique habitats and to address other important regional and statewide issues, new approaches are necessary. As recommended throughout this book, it is necessary to take an ecosystem stewardship approach—​“a strategy to respond to and shape social-ecological systems under con-

ditions of uncertainty and change to sustain the supply and opportunities for use of ecosystem services to support human well-being” (Chapin et al. 2010). A renewed commitment to stewardship of the state’s public lands is needed. This will not be easy given the current regulatory framework—​local control over local land use, fiscal constraints, and protection of private property rights, among other regulatory structures. This chapter has outlined some newer initiatives that attempt to use infrastructure development to advance habitat conservation on a landscape level as well as potential funding from carbon credits. Admittedly these initiatives are tentative and complex; they rely on voluntary willingness to implement them. Ultimately, conservation of the state’s resources is predicated upon better state urban policy, directing growth inward and constraining further growth on greenfields. While the state has a few tools at its disposal, such as Assembly Bill 857, the GHG requirements in CEQA, and the nudging requirements in Senate Bill 375, real progress will require the state legislature to reform General Plan requirements and elements. It will require changing the fiscal rewards of cities and the provision of new funding streams for conservation. Fundamentally, conservation will be successful only by virtue of better urbanization patterns. This chapter has not even considered impacts on habitat of development in the wildlandurban interface, another artifact of local land use regulations. Many of the arising issues from that phenomenon—​such as fire, habitat fragmentation, flooding, and the spread of invasive species—​are covered elsewhere in this book in the chapters exploring the affected ecosystems.

Summary California has been fortunate to have had some of its magnificent ecosystems conserved starting in the late nineteenth century. At the same time, the state has experienced tremendous population growth and urbanization. This has led to conflicts between land conservation and development. These conflicts are local because land use planning—​determining zoning and where development will occur—​is the prerogative of cities and counties, and there is no regional coordination or goal setting for future growth. Urban growth is still seen as the key to the prosperity of localities, so little incentive exists to conserve land, to build denser more contained cities, and to collaborate across jurisdictions for infrastructure, land use, or revenue sharing. This is further complicated by the legacy of the Fifth Amendment of the Constitution, which stipulates that in order for land to be conserved, it must be purchased at fair price from its legal owners, in the context of increasing constraints on local government’s ability to raise taxes or other funds. This chapter describes how the state legislature, and voters through the ballot initiative process, have passed a number of laws and created new agencies to attempt to improve conservation planning and to protect conservation lands. These include the California Coastal Act and the creation of the Coastal Conservancy, the California Environmental Quality Act, HCPs and NCCPs, the Williamson and Quimby Acts, and public involvement via voting or lobbying to protect the state’s natural resources, as with the Bay Conservation and Development Commission. But with the reduction of funding since the early 1980s due to ballot initiatives, conserving California’s lands has been more difficult. Because government must purchase its land, and since land in California is

expensive, conserving more lands is costly. Despite this challenge, innovative funding streams have emerged—​especially with the greater participation of public parties in conservation land acquisition—​as well as alternative forms of land conservation such as conservation easements. Better land use is at the heart of conservation of California’s magnificent heritage. In the end, conservation rests upon political will and the support of the state’s residents. The state legislature will have to restructure the state’s taxation system and the way in which private property rights are interpreted, and create new governmental regulations to ensure the state’s social and ecological resilience to climate change. Successful conservation will require a change in land development patterns and fundamentally in the way people live on the land.

Acknowledgments for Box 40.1 “Reconstructing California’s Land Conservation History” (page 902) would have not been possible without the funding from the Spatial History Project (Wallenberg Foundation) and the Bill Lane Center for the American West, and the ViceProvost Undergraduate Education program at Stanford University. Several acknowledgments are also due to individuals and institutions for valuable contributions and data: Jon Christensen at the University of California at Los Angeles and Zephyr Frank at Stanford University for their support, mentorship, and insights to the development of the idea of the reconstruction of California conservation history. Jon Christensen for the access to the Herbarium records data sets, and Zephyr Frank and Richard White for the access to Spatial History’s data sets. James H. Thorne at the Information Center for the Environment at the University of California at Davis for the access to the Wieslander project, especially the historic and modern vegetation data. Maggie Kelly and David Ackerly at the University of California at Berkeley for their insights on Wieslander data and collections, and the conservation environment of the Bay Area. Open Space Council for providing the inspiration and the connections between academic research and the on-the-ground managers. GreenInfo Network for creating and making available the invaluable California Protected Areas Database, one of the building blocks of this work. Jake Coolidge, Erik Steiner, and Geoff McGhee for their insights on cartographic design and how to visualize this type of information. And last but not least, to all the students and collaborators who contributed to the collection of the data set herein presented: Alice Avery, Alexandra Peers, Emily Francis, Claudia Preciado, Taz George, Matthew Walter, Alex Powel, Ma’ayan Dembo, and Salma Zahedi.

Recommended Reading Curtin, D. J., and C. T. Talbert. 2004. Curtin’s California land use and planning law. 24th edition. Solano Press Books, Point Arena, California. Fairfax, S. K., L. Gwin, M. A. King, L. Raymond, and L. A. Watt. 2005. Buying nature: The limits of land acquisition as a conservation strategy. MIT Press, Cambridge, Massachusetts. Fulton, W. 1997. The reluctant metropolis: The politics of urban growth in Los Angeles. Solano Press Books, Point Arena, California. Jensen, D. B., M. S. Torn, and J. Harte. 1993. In our own hands: A Land Use Regulation for Resource Conservation   921

strategy for conserving California’s biological diversity. University of California Press, Oakland, California. Meyer, A. 2006. New guardians for the golden gate: How America got a great national park. University of California Press, Oakland, California. Schrag, P. 1999. Paradise lost: California’s experience, America’s future. University of California Press, Oakland, California. Sellers, C. C. 2012. Crabgrass crucible, suburban nature, and the rise of environmentalism in twentieth-century America. University of North Carolina Press, Chapel Hill, North Carolina. Walker, R. 2007. The country in the city: The greening of the San Francisco Bay Area. University of Washington Press, Seattle, Washington.

Glossary Annexation  The addition of new land to a city. California Environmental Quality Act (CEQA)  This act requires that state and local agencies follow a protocol for analysis and public disclosure of environmental impacts of proposed projects and adopt feasible measures for mitigating those impacts. California State Office of Planning and Research (OPR)  This was created by statute in 1970 and constitutes the state planning agency. Carbon credits  Refers to funds or credit in carbon markets, originated from offsetting carbon emissions from operations such as air travel, industries, and so on. Conservation easement  An agreement by a property owner to maintain his or her land in private open space—​ no development will occur—​in exchange for a tax rebate on that property. Conservation easements may have a time limit, after which they must be renewed or not. They are permanently tied to the title to the land and inhere at the sale of a property. Coordinating council  Refers to a council created to develop policy proposals to improve planning and coordination of growth. Councils of government  Metropolitan planning agencies that coordinate regional transportation infrastructure. Endangered Species Act (ESA)  Act passed in 1973 with the aim to preserve threatened and endangered species such as eagles, wolves, and grizzly bears and their habitat. The ESA can be (and has been) used to challenge development. The ESA aims to preserve species and the habitats they depend on. Federal Land Policy and Management Act (FLPMA)  Passed in 1976, the FLPMA is the driving force behind the Bureau of Land Management (BLM), and it establishes the multiple use of the bureau. According to BLM’s website (http://www. blm.gov/flpma/FLPMA.pdf), the act states that “the public lands be retained in Federal ownership [ . . .] disposal of a particular parcel will serve the national interest” and also that “the public lands be managed in a manner that will protect the quality of scientific, scenic, historical, ecological, environmental, air and atmospheric, water resource, and archaeological values; that, were appropriate, will preserve and protect certain public lands in their natural condition; that will provide food and habitat for fish and wildlife and domestic animals; and that will provide for outdoor recreation and human occupancy and use.” Fee title ownership  Refers to when an Open Space management agency owns the title of that property (which means it has its property rights). Habitat conservation plans (HCPs)  Plans to implement the mandates of the ESA for the protection of threatened and endangered species and their habitat. HCPs took habitat into consideration, and they created incidental take permits (ITPs), 922   Policy and Stewardship

which allowed—​for the first time—​t he taking of a limited amount of habitat via permit. Homestead Act  Passed in 1862, this act secured homesteads to actual settlers on the public domain. Hotspots  Spatial locations with high frequency of conservation areas. Hottimes  Time periods with high concentration of conservation activity, as, for example, acquisition of land for conservation. Incorporation  The addition of new cities with a threshold in the number of residents. Local agency formation commission  County-level agencies with the responsibility to approve or disapprove any petition for incorporation, annexation, and special district formation or dissolution. Mitigation banks  Funds from development fees, mitigation, and other sources are held in these “banks” for purchasing land in the future. Multiple use  Defined in the Multiple Use Sustained Yield Act as “management of the public lands and their various resource values so that they are utilized in the combination that will best meet the present and future needs of the American people.” Multiple Use Sustained Yield Act (MUSYA)  Passed in 1960, this act authorizes and directs “that the national forests be managed under principles of multiple use and to produce a sustained yield of products and services, and for other purposes,” according to the act’s website (http://www.fs.fed. us/emc/nfma/includes/musya60.pdf). National Forest Management Act (NFMA)  Passed in 1976, this act mandates that “the public interest is served by the Forest Service, Department of Agriculture, in cooperation with other agencies, assessing the Nation’s renewable resources, and developing and preparing a national renewable resource and program, which is periodically reviewed and updated,” according to the act’s website (http://www.fs.fed.us/ emc/nfma/includes/NFMA1976.pdf). Natural communities conservation plans (NCCP)  These are an effort to encompass entire ecosystems and their processes, in contrast to HCPs (which are aimed at protecting one species at a time). Open Space  Open Space (as opposed to open space, or undeveloped land) refers to “lands protected through fee title ownership by a public agency or non-profit land conservation organization” (GreenInfo Network 2013). Open Spaces include a wide array of types of properties from city parks to national parks, Bureau of Land Management lands used for recreation and resource extraction, or Department of Defense lands. However, Open Space does not include private land in conservation easements. Organic Act  A law that regulates how a certain agency acts on behalf of the public good. Police power  Constitutional right provided to regulate land use. It follows from the Tenth Amendment to the Constitution and gives state rights and powers not delegated to the United States. Special districts  A type of government unit to provide a specific service—​such as schools, sewers, street lighting, parks, or acquisition of lands—​for conservation Taking law  Habitat taking is a component of the Endangered Species Act. It means that in specific cases habitat of listed species can be taken (for other use) with the trade-off of mitigation land. Usufructory right  A right to use.

References Barbour, E. 2002. Metropolitan growth planning in California, 1900–​2 000. Public Policy Institute of California, San Francisco, California. Barton, G. A. 2000. Empire forestry and American environmentalism. Environment and History 6:187–​2 03. Brooks, T. M., R. A. Mittermeir, C. G. Mittermeier, G.A.B. da Fonseca, A. B. Rylands, W. R. Konstant, P. Flick, J. Pilgrim, S. Oldfield, G. Magin, and C. Haig-Taylor. 2002. Habitat loss and extinction in the hotspots of biodiversity. Conservation Biology 16:909–​923. California Department of Transportation (CALTRANS). 2013. Regional blueprints. . Accessed March 12, 2013. California Transportation Commission. 2010. Regional Transportation Guidelines. . Accessed March 12, 2013. Chapin, F. S.,S. R. Carpenter, G. P. Kofinas, C. Folke, N. Abel, W.C. Clark, P. Olsson, D. M. Stafford Smith, B. H. Walker, O. R. Young, F. Berkes, R. Biggs, J. M. Grove, R. L. Naylor, E. Pinkerton, W. Steffen, and F. J. Swanson. 2010. Ecosystem stewardship: Sustainability strategies for a rapidly changing planet. Trends in Ecology and Evolution 25:241–​249. Crew, M. H. 1990. Development agreements after Nollan vs. California Coastal Commission, 483 US 825 (1987). The Urban Lawyer 22:23– ​58. Czech, B. 2004. A transdisciplinary approach to conservation land acquisition. Conservation Biology 16:1488–​1497. Dear, M. 1996. In the city, time becomes visible: Intentionality and urbanism in Los Angeles, 1781–​1991. Pages 760105 in Allen J. Scott and Edward W. Soja, editors. The City: Los Angeles and Urban Theory at the End of the Twentieth Century. University of California Press, Berkeley, California. Dowall, D. E. 1984. The suburban squeeze: Land conversion and regulation in the San Francisco Bay Area. University of California Press, Berkeley, California. Fogelson, R. M. 1967. The fragmented metropolis Los Angeles, 1850–​ 1930. University of California Press, Berkeley, California. Fulton, W. 1997. The reluctant metropolis: The politics of urban growth in Los Angeles. Solano Press Books, Point Arena, California. Fulton, W., and P. Shigley. 2012. Guide to California planning. Fourth edition. Solano Press Books, Point Arena, California. Grantham, H. S., M. Bode, E. McDonald-Madden, E. T. Game, A. T. Knight, and H. P. Possingham. 2010. Effective conservation planning requires learning and adaptation. Frontiers in Ecology and the Environment 8:431–​437. GreenInfo Network. 2013. Bay Area protected areas database (BPAD). . Accessed November 20, 2013. Gregor, H. F. 1957. Urban pressures on California land. Land Economics 33:311–​325. Hise, G. 1997. Magnetic Los Angeles: Planning the twentieth century metropolis. Johns Hopkins University Press, Baltimore, Maryland. Hundley, N. 1992. The great thirst: Californians and water, 1770s–​ 1990s. University of California Press, Berkeley, California. IUCN - International Union for Nature Conservation and UNEPWCMC - United Nations Environmental Program - World Conservation Monitoring Center. 2013. The world database on protected areas (WDPA). UNEP-WCMC, Cambridge, UK. . Accessed November 20, 2013. Knight, A. T., R. M. Cowling, M. Rouget, A. Balmford, A. T. Lombard, and B. M. Campbell. 2008. Knowing but not doing: Selecting priority conservation areas and the research-implementation gap. Conservation Biology 22:610–​617. Kostyack, J. 2001. NWF v. Babbitt: Victory of smart growth and imperiled wildlife. Environmental Law Reporter 31:10712–​10718. Lewis, P. G. 2000. The durability of local government structure: Evidence from California. State and Local Government Review 32:34–​48. Martin, D. T., and R. E. Wagner. 1978. The institutional framework for municipal incorporation: An economic analysis of Local Agency Formation Commissions in California. Journal of Law and Economics 21:409–​425.

Matute, J., and S. Pincetl. 2013. Unraveling petroleum. Report for Next Ten. . Accessed October 10, 2014. McHarg, I. 1969. Design with nature. American Museum of Natural History, New York, New York. McWilliams, C. 1979 [1949]. California, the great exception. Peregrine Smith, Santa Barbara, California. Meir, E., S. Andelman, and H. P. Possingham. 2004. Does conservation planning matter in a dynamic and uncertain world? Ecology Letters 7:615– ​622. Merenlender, A. M., L. Huntsinger, G. Guthey, and S. K. Fairfax. 2004. Land trusts and conservation easements: Who is conserving what for whom? Conservation Biology 18:65–​76. Nash, G. D. 1972. Stages of California’s economic growth, 1870–​1970: An interpretation. California Historical Quarterly 51:315–​330. Nicolaides, B. M., and A. Wiese. 2013. Suburban Disequilibrium. New York Times. Week in Review. April 7. Page 5. Olmsted, F. L. 1929. Report of state parks survey of California. California State Park Commission. Sacramento, CA. Pincetl, S. 2006. Conservation planning in the West: Problems, new strategies, and entrenched obstacles. Geoforum 37(2):246–​255. ———. 2004. The preservation of nature at the urban fringe. Pages 225–254 in J. Wolch, M. Pastor, and P. Drier, editors. Up against the Sprawl: Public Policy and the Making of Southern California. University of Minnesota Press, Minneapolis, Minnesota. ———. 2003. Nonprofits and park provision in Los Angeles: An exploration of the rise of governance approaches to the provision of local services. Social Science Quarterly 84(4):989–​1001. ———. 1999a. The politics of influence: Democracy and the growth machine in Orange County, U.S. Pages 195–212 in A. Jonas and D. Wilson, editors. The Urban Growth Machine: Critical Perspectives, Two Decades Later. State University of New York Press, Buffalo, New York. ———. 1999b. Transforming California: The political history of land use in the state. Johns Hopkins University Press, Baltimore, Maryland. Press, D. 2002. Saving Open Space: The politics of local preservation in California. University of California Press, Berkeley, California. Radeloff, V. C., F. Beaudry, T. M. Brooks, V. Butsic, M. Dubinin, T. Kuermmerle, and A. M. Pidgeon. 2013. Hot moments for biodiversity conservation. Conservation Letters 6:58–​66. Reisner, M. 1986. Cadillac desert. Viking, New York, New York. Santos, M. J., T. Watts, and S. Pincetl. 2014a. The push and pull of Land Use Policy: Reconstructing 150 Years of Development and Conservation Land Acquisition. PlosONE. doi: 10.1371/journal. pone.0103489. Santos, M. J., J. H. Thorne, J. Christensen, and Z. Frank. 2014b. An historical land conservation analysis in the San Francisco Bay Area, USA: 1850 to 2010. Landscape and Urban Planning. 127:114–123. Schrag, P. 1998. Paradise lost: California’s experience, America’s future. University of California Press, Berkeley, California. Silva, F. J., and E. Barbour. 1999. The state-local fiscal relationship in California: A changing balance of power. Public Policy Institute of California, San Francisco, California. Sokolow A. D. 1990. The Williamson Act: Twenty-five years of land conservation. The Resources Agency, Department of Conservation, Sacramento, California. Szabó, P., and R. Hédl. 2011. Advancing the integration of history and ecology for conservation. Conservation Biology 25:680–​687. Taylor, P. S. 1975. California Water Project: Law and politics. Ecology Law Quarterly 5:1–​52. Thorne, J. H., P. R. Huber, E. O’Donoghue, and M. J. Santos. 2014. The use of regional advance mitigation planning (RAMP) to integrate transportation infrastructure impacts with sustainability: A perspective from the USA. Environmental Research Letters 9: 065001. Trust for Public Lands (TPL). 2000. Building green infrastructure. TPL, San Francisco, California. . Accessed October 10, 2014. Wieslander, A. E. 1935. A vegetation type map for California. Madroño 3:140–​144. Wolch, J., M. Pastor, and P. Drier, editors. 2004. Up against the sprawl: Public policy and the making of southern California. University of Minnesota Press, Minneapolis, Minnesota.

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FORT Y- ONE

Stewardship, Conservation, and Restoration in the Context of Environmental Change AD INA M . M EREN LENDER , DAVID D. AC KER LY, K ATHAR INE SUD ING , M . REBECCA SHAW, and ER I K A Z AVALE TA

Introduction California’s ecosystems are diverse and dynamic; they have always experienced change. However, unprecedented human activity and impacts have brought about increasingly dramatic changes that require our attention. California’s rapidly growing human system will include inevitable increases in land and water use. In the next fifteen years we expect California’s population to grow by more than eight million people, a number as large as the current size of New York City. Continued development at current, low densities—​w ith only 10% infill—​w ill consume an estimated 2.1 million hectares (5.1 million acres) of undeveloped land in that time (Landis and Reilly 2003). The existing water supply cannot accommodate this expected growth while also sustaining California’s diverse and important agricultural sector. Global climate change will exacerbate this shortfall, with expected 30% declines in snowfall and a further increase of 0.5°C to 1.0°C by 2025 and even more dramatic changes through the rest of the century (see Chapter 14, “Climate Change Impacts”). The rising likelihood of extreme rainfall events and intensifying drought, as we experienced in the winter of 2013–​2014, will profoundly affect California’s natural and human sys-

tems. We must anticipate and prepare for change to California’s ecosystems from both old forces like development and new, interactive forces like climate change. A range of anthropogenic changes—​including climate change and other changes related to human development such as ocean acidification, biodiversity loss, land use change, freshwater use, and higher nitrogen and phosphorus levels—​a re all on the rise around the world (Fenn et al. 2010, Hayhoe et al. 2004). With the widespread recognition of these dramatic changes and the documented accelerating rates of change, many argue that Earth has experienced a decisive break from what it experienced during most of the Holocene and we have entered a new epoch: the Anthropocene (Crutzen 2002). This new epoch could be dated as early as the beginning of the Industrial Revolution, or when the great acceleration in population began in the 1950s, or at the start of the current trajectory of anthropogenic climate change around 1970 (Zalasiewicz et al. 2010). While the question of whether this current time in history is distinct and lasting enough to be designated a new epoch on a geological time scale is still in debate, humans are indisput925

ably leaving their mark across the planet and changing the trajectory of Earth’s biological and physical systems (Vince 2011). This recent period of industrialization and globalization has led to increasing homogenization of natural communities at unprecedented rates because humans have facilitated massive biotic exchanges among regions (Olden et al. 2004). Biotic homogenization refers to simplification of a community’s composition and happens when native ecosystems are assimilated by widespread exotic or weedy species reducing diversity (McKinney and Lockwood 1999). In California this is best illustrated in the prairies and grasslands, where nonnative species make up over half of the California flora (Bartolome et al. 2007, Stromberg et al. 2007; see Chapter 13, “Biological Invasions”). Given the trajectory of projected change in the state, our greatest challenge is to ensure that California’s ecosystems persist and retain their capacity to sustain vital ecosystem services and biological diversity while building a sustainable and resilient society. Rates of change in California’s ecosystems will increase in this century unless dramatic action is taken to limit land conversion, realign water management, and dramatically slow greenhouse gas emissions into the atmosphere. Growth and conservation policies related to land use, such as increasing urban infill, could greatly enhance conservation outcomes in California. Land use patterns reflect land economics, which in turn reflect taxation structures; for example, Proposition 13 reduced local government’s ability to keep up with urban infrastructure, stimulated growth in suburban communities, and reduced options to finance infill developments (see Chapter 40, “Land Use Regulation for Resource Conversation”). However, even with dramatic improvements to land and water use patterns, we face the highest rates of climate change recorded in human history. This means that California’s ecosystems that we observe today, as detailed throughout this book, could be transient, albeit on the scale of thousands of years, and that the trajectories of the transitions they will undergo are dramatic and highly uncertain. For example, forecasted increases in summer temperatures suggest that desert plants could grow in Bakersfield and the Central Valley; and in the San Francisco Bay Area the vegetation is expected to be more similar to what we currently observe in southern California (Loarie et al. 2008). This reality requires that researchers, managers, decision makers, and the public make a concerted break from the notion of equilibria or stationary ecosystems. Recognizing increased dynamism in ecosystems does not imply abandoning biodiversity conservation. Rather, a more concerted conservation effort and improvements to ecosystem management that incorporate an increased understanding of global and local forces of change are more important than ever to increase social and ecological system resilience to rapid change. In this chapter we examine the challenges environmental change presents to conservation, stewardship, and restoration in California. Many of these have been raised in earlier chapters of this book. We attempt to emphasize and integrate the actions various authors have suggested to best address future protection and management of California’s ecosystems through the lens of inevitable change.

Photo on previous page: Environmental stewards of all ages at Pepper­wood Preserve, in Sonoma County. Photo: Greg Damron, . ­

926   Policy and Stewardship

Ecological Responses to Environmental Change Protection and management begin with understanding of how ecosystems respond to environmental changes. While land use and other changes have been occurring for centuries, rates of change and associated responses such as species introductions are accelerating. Other changes are essentially new and unprecedented, such as accelerating climate changes to which we are already committed. Ecological responses to these growing challenges can take millennia to be fully realized. Inertia within ecosystems resists change; systems can appear to be in equilibrium when in fact they are simply experiencing lags preceding more radical ecosystem reorganization. One of the most important lessons from the paleoecological record is that time lags of one thousand to five thousand years after episodes of major changes, such as in climate, are expected in the observed responses of ecosystems (see Chapter 8, “Ecosystems Past: Vegetation Prehistory”). For example, climate change induces disequilibria at both leading and trailing edges of species ranges—​w ith time lags at the leading edge due to delays in migration, succession, and evolutionary adaptation, and on the trailing-edge due to delayed local extinctions and disruptions of species interactions (Svenning and Sandel 2013). Species respond to change at varying rates depending on their life history characteristics, tolerances to environmental variation, and degrees of interaction with and dependence on other species in the community. For example, species’ lifespan and dispersal ability influence their response rates and the duration of lag effects. Species responses can also lag shifts in ecosystem processes such as soil formation and biochemical cycles (Jackson and Sax 2010). At the community level, ecological responses to multiple stressors can interact as well as lag changes in drivers. In some California ecosystems, climate change will likely favor exotic species that are more productive under the new climate conditions, with cascading ecosystem effects such as increased carbon storage rates and community effects such as displacement of native species (Ehrenfeld 2003). Lags in species and community effects are reflected by the fact that only twenty documented extinctions of California’s plant species have occurred in the past hundred or so years despite the conversion of about 43% of the land to agriculture and development throughout the twentieth century (see Chapter 11, “Biodiversity”). With an estimated 612 plant species now considered threatened across the state, far more extinctions are expected. Plant species extinctions are observed as they gradually unfold over human lifetimes through long-term monitoring at university research stations. For example, ecologists over a decade ago witnessed what could have been the last observation of beaked tracyina (Tracyina rostrata) at the Hopland Research and Extension Center (Heise and Merenlender 2002). Attribution of ecological changes to particular stressors, while more often straightforward in the past—​such as the demise through overexploitation of animals like the California grizzly bear (Ursus arctos californicus)—​has also become more challenging in some cases. For example, it is clear that climate change is driving extinctions worldwide. However, attributing biotic response to even the most well-documented aspect of climate change—​warming—​is difficult. The complexity inherent in natural ecosystems and the disconnection between the temporal and spatial scales that global change operates on and those at which ecosystems respond makes it

especially difficult to attribute ecological changes to globalscale drivers alone. Some notable successes documenting climate-driven transitions discussed previously are worth replicating, such as documented shifts in distributions of small mammals in the Sierra Nevada and intertidal invertebrates at the coast (see Chapter 14, “Climate Change Impacts”). The topographic complexity of California makes it harder to generalize how species ranges may shift, as widely anticipated shifts northward and to higher elevations might also be accompanied by some downward movements, such as towards moister locations, and westward to cooler coastal areas. Detection and attribution of change will almost certainly improve as we combine high-quality baseline data and realtime tracking of ecological responses to change. Still, Mediterranean climates pose considerable challenges to climate change response detection because of their inherently high interannual variability in precipitation patterns and the potential for biotic changes to be strongly driven by climatic change during any season. A wide range of climate change responses have been documented in California and attributed to contrasting effects of temperature and precipitation (Rapacciuolo et al. 2014). California’s diversity of ecosystems and concomitant diversity of responses and feedback loops also complicate generalized predictions about the outcomes of global change forces. In California’s deserts, for example, increases in atmospheric carbon dioxide (CO2) and nitrogen are of concern as they could facilitate the proliferation of non-native species and associated wildfires at the expense of native species more tolerant to nutrient-poor conditions (see Chapter 30, “Deserts”). This potential shift could be counterbalanced by soil moisture reductions that limit non-native grass invasion, but other interactions between elevated atmospheric CO2 and both water use and temperature-sensitivity of different species might further complicate these responses (Dole et al. 2003). Alterations to water and land management required for humans to adapt will make predicting ecosystem response still more challenging. In sum, California’s current ecosystems will continue to change at varying rates; while strong lags may dampen initial rates of extinction, introduction, and disruption, ecosystem shifts and novel community types will eventually result. The future of California’s ecosystems remains unknown, but change is certain. Tackling these changes despite some areas of increased uncertainty are essential to guide effective stewardship and conservation.

Stewardship, Conservation, and Restoration The time lags and uncertainties inherent in ecosystem responses to climate and other environmental changes require us to rethink how we approach conservation and restoration and necessitate nimble adaptive management strategies to manage inherently transient ecosystems. Despite these challenges, there is much we can do to conserve, restore, and manage California’s ecosystems to safeguard biodiversity and human welfare into the future (Caro et al. 2012). Conservation in an era of change will need to: (1) manage toward a dynamic future rather than a static past; (2) be open to the challenge of reconsidering long-held conservation goals and objectives; and (3) extend collaboration to nontraditional private and institutional partners. Equally important, it requires that we reexamine and amend the scaffold of environmental, regulatory, and institutional policies that have begun to con-

strain our ability to respond effectively and that we build in sufficient flexibility to meet the future needs of conservation. Defining and obtaining consensus on achievable goals, measuring and monitoring conservation success against those goals, implementing adaptive responses and partnering with diverse institutions have all proven difficult in the conservation community even without considering climate change (Salafsky et al. 2002, Ferraro and Pattanayak 2006). Successful conservation will require “managing according to clear goals by which decisions are made and modified as a function of what is known and learned about the system, including information about the effect of previous management actions” (Parma et al. 1998). Stating achievable goals and building adaptive plans to achieve those goals are essential. In the future there will likely need to be a shift from a focus on maintaining specific patterns of species and habitat diversity towards sustaining underlying ecological and evolutionary processes that promote continued ecological functioning in the face of change. Conservation will be defined less by the persistence of current patterns and conditions, and more by managing change in ways that sustain core conservation values and human needs.

An Ecosystem Stewardship Framework In this period of unprecedented change, intervention may be needed simply to remain in the same spot along a degradation/transformation trajectory. Much as Lewis Carroll wrote in Through the Looking Glass (1871), “Now, here, you see, it takes all the running you can do, to keep in the same place. If you want to get somewhere else, you must run at least twice as fast as that!” This context of directional environmental changes forces us to consider new management approaches not premised on maintaining a particular community or genetic composition, but rather capable of shifting to maintain diversity as well as core ecosystem functions and important services over time (Choi et al. 2008). Although this type of approach has been discussed for some time among researchers, in practice it remains relatively new to explicitly incorporate ecosystem dynamism into land management strategies and to target outcomes that might vary and potentially include novel ecosystems (discussed later in this chapter). Ecosystem stewardship has been defined as “a strategy to respond to and shape social-ecological systems under conditions of uncertainty and change to sustain the supply and opportunities for use of ecosystem services to support human well-being” (Chapin et al. 2010). Defined in this way, the framework relies less on maintaining historical conditions of a community and more on managing the dynamics, pathways, and rates of ecological change. If the current community status is viewed as desirable, efforts should be made to maintain the components of the current system by increasing stabilizing feedbacks that contribute to community persistence, and/or to maintain components of the system in shifting distribution over time. These goals should be defined and articulated as part of the ecosystem resilience concept. Resilience is conceptualized as the amount of change a system can undergo and still maintain similar function and structure with minimal intervention (Gunderson and Holling 2002). It is a critically and widely invoked concept, though still difficult to translate into practice in specific instances (Zavaleta and Chapin 2010). Stewardship, Conservation, and Restor ation   927

FIGURE 41.1 Many rangelands in California can be considered novel ecosystems due to the abundance of non-native grasses. These grasslands provide important ecosystem services, including forage. Conserving native species in these ecosystems is sometimes facilitated by livestock grazing, which can reduce competition from annual grasses that otherwise dominate. Photo: Jack Kelly Clark.

is one of the most widespread and intense sources of disturbance. Prescribed fire has been a preferred management strategy for thousands of years in California, and in some areas the use of fire by Native people has influenced the composition of ecosystems we see today (see Chapter 10, “Indigenous California”). Maintaining fire as a management tool is critical to the protection of many ecosystems, and when fire is not feasible, other types of disturbance such as selective harvesting and grazing could be required (Stromberg et al. 2007). In California’s serpentine grasslands, for example, selective grazing on nitrogen-rich exotic grasses has been used with success to combat effects of nitrogen deposition (Weiss 1999) (Figure 41.1). However, this same method in coastal sage scrub is less effective because sufficient grass forage occurs only in very wet years (see Chapter 22, “Coastal Sage Scrub”). In many valley grasslands, however, invasions occurred so long ago and are spatially so extensive that fire and mowing often no longer suffice for large-scale restoration (see Chapter 23, “Grasslands”). Though grazing or mowing have been successfully used to mitigate the effects of nitrogen deposition in some grasslands, they might not be effective or feasible in other systems. Here, new management techniques need to be developed or goals altered to reflect the transformation of the systems in question to new states. Adopting an ecosystem stewardship framework acknowledges the need to incorporate expected changes into interventions and to consider how to maintain system integrity over large potential ranges of variability rather within narrow bounds.

Conservation Stewardship on California’s Tahoe National Forest (TNF) illustrates this approach. Here, as with many forest systems in the western United States, drought concerns, amplified by rising temperatures and lowered snowpacks, interact with dense forest structure from fire suppression to create severe forest hazards, extreme flood events, and new opportunities for invasive species spread (see Chapter 27, “Montane Forests”). Managers on the TNF recognized several general principles as opportunities for stewardship in the face of rapidly changing climate: (1) managing for drought- and heat-tolerant species and ecotypes, (2) reducing impact of current anthropogenic stressors, (3) managing for diverse successional stages, (4) spreading risks by including buffers and redundancies in natural environments and plantations, and (5) increasing collaboration with interested stakeholders. In addition, for reforestation they considered more drought-tolerant germplasm and species mixes and prioritized sensitive-species management actions at the “leading edges” of species ranges (likely favorable future habitats) rather than at “trailing edges” (Joyce et al. 2008). North and colleagues place a similar emphasis on forest resilience in this volume (Chapter 27, “Montane Forests”), in which large trees are recognized as providing more secure carbon storage compared to high-density, small trees that are prone to pests, pathogens, and fire. The ecosystem stewardship framework also emphasizes the value of expecting and exploiting disturbances such as fires, floods, and high rates of nitrogen deposition rather than attempting to prevent such events. Managing for disturbance is an important aspect of maintaining healthy ecosystems (see, e.g., Chapter 3, “Fire as an Ecosystem Process”). Natural disturbance plays a well-documented, essential role in developing ecosystem structure and function (Attiwill 1994). Fire 928   Policy and Stewardship

Halting habitat loss for all of California’s ecosystems is crucial to maintain biological diversity. Diversity of genotypes, species, and functional groups influences a range of biogeochemical processes, trophic interactions, resistance to biological invasions, and dimensions of temporal ecosystem variability (Hillebrand and Matthiessen 2009, Hughes et al. 2008, Báez and Collins 2008, Isbell et al. 2009; see Chapter 15, “Introduction to Concepts of Biodiversity, Ecosystem Functioning, Ecosystem Services, and Natural Capital”). A review of over two hundred studies worldwide revealed that species losses result in decreases in productivity and influence decomposition and can be as disruptive to ecosystems as warming temperatures and other global change factors (Hooper et al. 2012). Habitat loss can thus, in addition to increasing risks of species extirpations and extinctions (see Chapter 11, “Biodiversity”), have biophysical impacts on ecosystems. For example, tree removal in oak woodland ecosystems can affect local climate, hydrology, and ecosystem processes. Trees in oak woodlands create microclimates distinct from grasslands, influencing soil development as well as nutrient cycling (see Chapter 25, “Oak Woodlands”). Avoiding land development patterns that continue to increase habitat loss and fragmentation is critical to maintain biodiversity, ecosystem processes, functions, and services. All development involves trade-offs, including green energy efforts such as planned solar development that threatens sensitive valley and desert habitat for a number of listed species (see Chapter 11, “Biodiversity”). International agreements provide a guide to setting largescale conservation goals for the state. While the United States has not signed the International Convention on Biodiversity

TA B L E 41.1

Protection Status

Percentage of surface area protected, selected regions

Region Antarctic

1

16.87

Brazil

18.70

22.45

East Asia

13.99

Eastern and Southern Africa

14.06

Europe

12.37

4

9.79

North America

16.21

North Eurasia

7.74

Pacific

1.89

South America

Military

8.17

Central America

North Africa and Middle East

3

0.50

Australia/New Zealand

Caribbean

2

Percentage

0

75 150

300 km

19.33

South Asia

6.53

South East Asia

9.55

FIGURE 41.2 Management status of protected lands in California, as categorized by the U.S. Geological Survey Gap Analysis Program (GAP), Protected Area Database v1.3 (2012). Map: Frank Davis.

Western and Central Africa

8.65

STATUS 1 L ANDS: Permanently protected and managed to maintain a

Total

11.58

Source: UNEP–World Conservation Monitoring Centre 2006.

natural state (e.g., national parks and wilderness areas). STATUS 2 L ANDS: Permanently protected and maintained primarily

in a natural state but may include some management actions such as wildfire suppression. STATUS 3 L ANDS: Permanent protection but subject to extractive uses

(ICB) (Convention on Biological Diversity 2011), California could play a leadership role in meeting these standards much the way it has in reducing carbon emissions following the adoption of the Kyoto Protocol. To meet the IBC standards, California would need to bring its rate of natural habitat loss to zero where possible and commit to reducing current carbon emissions by at least one half. California would need to ensure that 17% or more of every terrestrial ecosystem has full protection, along with 10% of the state’s marine and coastal areas. The international standards also require restoration of at least 15% of degraded areas within each ecosystem. Meeting these conservation targets is a critical step to protecting California’s spectacular biodiversity that took millions of years to evolve (see Chapter 11, “Biodiversity”). The California GAP analysis (1998) provides an estimate on the amount of protection afforded to each of California’s ecosystems. As part of this effort, levels of protection were ranked from 1 to 4, with 1 the highest, most permanent level of protection against habitat conversion and use and 4 the lowest level of protection. Approximately 15% of California currently falls under the most protected status (1), with an additional 3% in the second-highest protective category (Figure 41.2). Over half of the state lacks protection from habitat conversion; most of this is privately owned. The remaining 30% is managed for multiple uses by agencies such as the U.S. Forest Service and the Bureau of Land Management. While 15% might seem high compared to other areas of the world (Table 41.1), these protected areas in California are almost entirely in high-elevation Sierra Nevada or in Mojave and

such as logging, Off Highway Vehicle recreation and mining. STATUS 4 L ANDS: Includes private lands of California that lack formal

protection (such as legally recognized easements or other known mandates) and where development and conversion of natural habitats and development are generally allowed.

Sonoran Desert ecosystems. Other low-elevation areas in the Great Valley and along the coast, along with many rare habitat types, have very little protection to date. The nearshore marine ecosystems present a different story. Under the Marine Life Protection Act of 1999, a public-private partnership established marine protected areas using the best available scientific methods. As a result, 16.4% of the marine ecosystems are captured in protected areas with varying goals as compared to the 2% global average. For example, five National Marine Sanctuaries now lie along the West Coast, protecting over 31,263 square kilometers in the California Current system and especially benefiting kelp conservation (see Chapter 17, “Shallow Rocky Reefs and Kelp Forests”). With approximately 18% of state marine area under designated protected status, however, only 243 square kilometers (94 square miles) fall under the highest level of protection, which prevents all fishing or habitat loss. Current conservation actions required to fill the gaps among these areas remain piecemeal and reliant on private sector investment. To meet and hopefully exceed these conservation goals, California will need to continue to embrace the importance of biodiversity for human well-being and the moral, ethical, Stewardship, Conservation, and Restor ation   929

Essential Connectivity Areas Less cost Most cost Natural landscape blocks Potential riparian connections Interstate connections Ecoregions

0

80

160

240

N km

FIGURE 41.3 California’s Essential Habitat Connectivity Project. Proposed linkages across California based on expert opinion of habitat requirements for a select group of focal species. Data and map: California Department of Fish and Wildlife, Spencer et al. 2010.

and practical responsibility to protect species and their habitats, as well as to strengthen public connections to nature and increase general understanding of the role that species and natural systems serve in supporting society. All of these elements are important to increase public and political support for conservation.

Habitat Connectivity Remnant protected areas are often too small to allow persistence of viable species populations. Connecting protected areas into networks can increase persistence; in particular, the need to recover endangered species and rare habitat types has driven much of the demand for habitat connectivity. Plans to increase connectivity through corridors, ecological networks, and other landscape features to minimize continued fragmentation and associated species extinctions are widespread. The most common approach to connectivity is to maintain and restore habitat that will provide pathways among protected natural areas for wildlife movement. However, it is not always clear that connecting wildlands through linear habitat features across disturbed landscapes really does enhance species persistence within reserves. The goal of reconnecting landscapes comes from our theoretical and empirical understanding of how habitat fragmentation contributes to rates and patterns of species extinction. Fragmentation is defined as the transformation of a continuous habitat into habitat patches varying in size and configuration (Fahrig 2003). Habitat loss and consequential fragmentation is the largest current threat to the world’s biodiversity (e.g., Dirzo and Raven 2003). The extent to which species distributions are expected to shift due to climate change has increased the justification for additional protected areas. Many existing protected areas are currently too small to be resilient in the face of today’s changes, let alone those projected into the future (Heller and Zavaleta 2009). Climate change compounded by habitat loss and fragmentation could impede species’ range shifts to such an extent that population and species extinctions could result (Nuñez et al. 2013). The built environment poses barriers to species movement between protected and unprotected remnant natural areas, and even areas with high habitat connectivity could prevent species to adapt to a changing climate and shifting ecological zones for a variety of reasons. In particular, as conditions warm, alpine plants will have nowhere to go, and those occurring at the edges of their ranges or within small isolated patches might not survive. Species’ ability to move through landscapes of varying composition and through corridors of varying size and configuration range widely; what constitutes connectivity for one group of organisms might not for another (Chester and Hilty 2010). Many scientists and land managers have assumed that improving the connectivity of existing protected area networks will improve the chances of species movement and adaptation. In California, consultants have proposed linkages between many protected lands across the state using simplified rulebased models of species habitat preferences (California Essential Habitat Connectivity Project 2010) (Figure 41.3). Habitat connectivity to facilitate animal and plant movement is one of the most frequently promoted strategies addressing rapid climate change resulting from anthropogenic disturbance (Heller and Zavaleta 2009). As part of this strategy, habitat corridors have been adopted to make protected area networks more resilient to climate change (Vos et

al. 2008). Despite the fact that habitat connectivity is the go-to solution for increasing reserve resilience to climate change, most methods remain untested. The most common approach is to establish corridors that track the species expected range shifts due to climate change in order to permit the expected movement of the species to more suitable habitat in the future (Lawler et al. 2013). This approach depends on models that have high levels of uncertainty about species climate sensitivity as well as the climate change models themselves. A more problematic aspect of this species modeling approach is forecasting the distribution of species across novel climates when we lack information about whether these novel climates and the vegetation they support in both the near- and long-term will provide suitable habitat. A simpler alternative that avoids the inherent uncertainties in a species-based approach is to design linkages based on expected rates of climate change and the distributions of climates across space and time (Loarie et al. 2009, Ackerly et al. 2010). This approach can take advantage of high-resolution, downscaled climate models (e.g., Flint 2013) and avoids some of the pitfalls associated with species habitat modeling. It focuses on features of climate that could influence reserve network resilience, based on the following assumptions: (1) the advantages of connectivity are greatest for areas that will experience faster rates of change, (2) a reserve network that harbors greater climatic diversity will allow greater adaptation, (3) maintaining access to cooler climates is a high priority, and (4) corridors that track isotherm movement will facilitate climate migrants trying to remain in the same climate analog over time. In other words, it is useful to think about climate space and diversity across the entire landscape as well as how fast organisms will need to move to stay in the current climate (Loarie et al. 2009; Burrows et al. 2014). Unique physiographic features could be important to protect in order to maintain landscape connectivity for multiple species. These include alluvial valleys, river mouths, and other places where hydrologic flow paths provide conditions that promote increased diversity and are often restricted geographically (Klausmeyer et al. 2011). Rarely are these different goals for maintaining connectivity met by the same corridors across the landscape, and it is not possible to secure corridors among all habitat fragments. Hence it could be more important to enhance the permeability of the matrix landscape surrounding protected areas through measures such as diversifying agricultural landscapes, adding highway crossings, and removing fences to facilitate as much species movement as possible across moderately modified landscapes. In some cases, species might only require short-distance movement to persist. In these cases, enlarging core reserves could be more effective than establishing corridors between reserves. More research is needed to examine the ecological and economic trade-offs between augmenting and acquiring new reserves and conserving habitat corridors (Shaw et al. 2012). Conservation planning needs to be coupled with land use planning. Conservation strategies that include multiple environmental benefits, reduce time delays, and increase predictability for the development community are desirable and will be implemented more readily.

Ecological Restoration The chapters in this book describe many ecosystems that persist in California today. Yet we know that land use change Stewardship, Conservation, and Restor ation   931

over the years has taken its toll and that large areas have been converted to urban and agricultural land. Some ecosystem loss stands out as particularly severe, such as the loss of 90% of California’s wetlands (Dahl 1990; see Chapter 31, “Wetlands”). With less than 10% of this ecosystem remaining, wetland restoration is essential to recover some of what has been lost. Other ecosystems in the state remain widespread but have been greatly modified by agriculture and residential development, such as California’s oak woodlands (Chapter 25). In oak woodlands, field surveys show that 67% of trees on average are greater than 13 centimeters in diameter at breast height (cm dbh) and 37% of trees were found to be greater than 61 cm dbh (Gaman and Firman 2008). Inadequate regeneration could be affecting at least three oak species (blue oak Quercus douglasii, valley oak Q. lobata, and Engelmann oak Q. engelmannii) (Koenig and Ashley 2003, Zavaleta et al. 2007, McLaughlin and Zavaleta 2013). Oak restoration is often required to mitigate for continued habitat loss, degradation, and lack of regeneration (McCreary 2009). Today, much restoration is shifting away from goals of returning communities to particular reference assemblages and towards rehabilitating systems to a point along a development trajectory that allows for self-sustaining population, community, and ecosystem processes (Suding 2011, Choi et al. 2008, Hobbs and Cramer 2007). Restoration of these processes can also lead to higher or desired levels of particular functions or services, such as erosion control and drinking water quality. However, we lack experience in implementing ecosystem management goals other than the traditional goals that emphasize native species and permanence. In some cases, ecosystems have been so strongly impacted that it is not an option to use a historical analogue as a target for restoration. These systems are often termed “novel ecosystems.” Hobbs and others (Hobbs et al. 2006, Hobbs et al. 2009) have defined a novel ecosystem as one that lacks a historical analog and has crossed a threshold to where reversing these impacts is not feasible. In these cases, goals based on historical conditions can be impossible to attain or require more intensive and continual management than can be reasonably accomplished. Eventual failure in cases like this can both waste valuable public investment and leave ecosystems in poor condition. Instead, consideration of how these “novel” systems can have value and of how to manage these systems for other goals, such as providing ecosystem services or natural areas for recreation and enjoyment, is warranted (Daily et al. 2009; see Chapter 39, “Urban Ecosystems”). The widespread existence of and forecasted increase in novel California ecosystems impels us to adopt equally novel approaches to restoration and to make increased investments in managing ecosystems. For example, an approach to protect important ecosystem processes, functions, and services could focus on ensuring persistence of diverse functional groups (Suding et al. 2008). A functional group is a set of species that have similar traits and that thus are likely to be similar in their effects on ecosystem functioning. Functional diversity can be very important because it goes beyond diversity and community composition and focuses on how species influence ecosystem dynamics, stability, productivity, nutrient balance, and other functions (Tilman 2001). Some express concern that by recognizing the role of novel ecosystems we will abandon efforts to prevent continued modification of natural ecosystems. However, acknowledging that novel ecosystems are providing essential ecosys-

932   Policy and Stewardship

tem services and might not warrant investment to recover to their historical composition can free resources to protect more pristine ecosystems and restore native plant and animal communities where they are more vital for biodiversity conservation. An increased level of uncertainty is associated with trying to predict how animal populations will respond to future change; models are aimed at forecasting changes to California’s bird communities involving novel communities as a result of climate change (see Chapter 11, “Biodiversity”). Biodiversity goals such as native species persistence and diversity require more and innovative attention in the context of accelerating environmental change. Intensive and continued management (essentially gardening) can be justified in rare cases due to unique aspects of natural heritage or place, but more of biodiversity protection will need to emphasize movement and change, and limited resources may often be better used targeting less-affected areas with more certain recovery trajectories. Generally, we should not expect to be able to maintain existing communities in their current, static form. A long-term perspective, including paleohistory, can provide useful information on how species migrated under specific climate conditions in the past and provides a rich context from which to consider current possibilities. Understanding what stakeholders need and involving them in planning and implementation also encourages long-term public investment in site stewardship even as sites change. Often, both historical and stakeholder perspectives can be achieved with minimal trade-offs. For instance, urban creek restoration can generate high native plant diversity and abundant wildlife habitat while providing an open understory that allows children to explore (Lyytimaki and Sipila 2009). Even in more protected wilderness areas, where emphasis shifts towards historical goals, anticipation of and reconciliation with future climate and disturbance regimes are critical (Baron et al. 2009). Assisted migration is a controversial approach to conserving biodiversity under climate change (Vitt et al. 2010). It involves the purposeful introduction of genotypes or species not part of any (past) reference condition but adaptable to current and future likely conditions. Although thousands of conservation translocations have taken place in the past for other purposes (Godefroid et al. 2011.), assisted migration is a new approach to preventing species losses under climate change and pursuing increased ecosystem resilience. Some assisted migration efforts aim to prevent extinctions by deliberately expanding the ranges of endangered species, while others might pursue improved future vegetation cover and plant survivorship at a restored site by planting more drought-tolerant varieties. The latter contrasts starkly with the standard approach of relying on local genetic stock for planting efforts (Vander Mijnsbrugge et al. 2010). Hence, assisted migration is contentious because it can place different conservation objectives at odds with one another (McLachlan et al. 2007). While in some contexts it is time to rethink the focus on local genetic stock for plant restoration projects (as has begun to take place, for instance, in reforestation efforts in British Columbia, Canada), the risks of using alternative genetic stocks should be considered carefully before implementation. In particular, those that could dilute locally adapted gene pools could present greater biodiversity risks than introductions of more distinct southerly or lower-elevation taxa. Given the effort made to care for and maintain early restoration efforts, strong natural selection to maintain locally adapted

FIGURE 41.5 South Bay Salt Pond Restoration in San Francisco Bay. The lighter color areas surrounding the bay are urban. Because both the salt flats and the tidal marshes provide critical bird habitat, the restoration project is taking a piecewise approach, restoring tidal flow in some areas while maintaining drier salt flats in others. The diversified approach is allowing multiple restoration goals to be achieved. Photo: Cris Benton.

FIGURE 41.4 Don Edwards National Wildlife Refuge workshop participants viewing restored wetlands at the Environmental Education Center in New Chicago Marsh near Alviso, California. Photo: Cris Benton.

genotypes is less likely, thus newly introduced genotypes could degrade remnant native genotypes through hybridization or competition under temporarily subsidized conditions. Maintaining relatively local genetic solutions, such as by using genotypes and ecotypes found within a larger, given watershed area and in similar microclimates, is the safest option for restoring and maintaining local biodiversity. While the extent of invasion and land use impacts in California may pose serious challenges to ecological restoration, several large-scale solutions and innovative strategies have successfully overcome these impacts. As part of one of the largest wetland restoration programs in the western United States, 6,100 hectares of former commercial-production salt ponds are being reconnected to the South San Francisco Bay (Thebault et al. 2008) (Figures 41.4, 41.5). Restoration program managers are balancing salt pond retention with reconversion to tidal salt marsh because both currently provide critical habitat for endangered bird species. Given the uncertainties inherent in restoring a mixed complex of wetland types without historical analog, the project developed targets for a range of restoration goals, with predefined triggers for reevaluation if monitored components deviate from the desired range. This approach allows both specificity in restoration planning and the ability to adapt and address uncertainty. This is a good example of a large-scale project, that along with the estua-

rine habitat being protected under the marine protected area (MPA) and 26,000 hectares of tidal habitat recovery being planned for the Sacramento–​San Joaquin Delta, takes us in the direction of restoring and protecting large-scale ecosystems rather than the more typical small-scale or piecemeal approach to mitigation of endangered species habitat loss (see Chapter 19, “Estuaries: Life on the Edge”). The Channel Islands represent the largest experiment in California testing how a combination of active management, mostly of invasive species, and passive restoration, resulting from human depopulation and declining land use, is leading to the ecological recovery of unique habitats (see Chapter 34, “Managed Island Ecosystems”). While the recovering island ecosystems are not identical in composition to their historical counterparts, with certain elements entirely missing and with well-established new species, the dynamics of recovery and change on the islands are remarkable and critical to observe as they “revert back to nature.” Guided by a desire to recover ecological processes, managers have ended livestock grazing, reduced human population densities, and removed feral non-native animals from the islands, including some voracious herbivores that have greatly affected plant growth and soil structure. Removal of ranch animals was an essential first step to trigger restoration of these islands, which has progressed considerably without additional interventions (Beltran et al. 2014). Some efforts have also successfully removed invasive plants that represent a high risk to restricted native communities. Other invasive plant populations are waning due to declines in disturbance from intensive land use and overgrazing; still others, such as common, naturalized annual grasses found throughout California, have been left untreated and will likely remain on the islands indefinitely. Positive outcomes apparent today on the islands include intact riparian vegetation along stream corridors that had been reduced to exposed cobble and mud, and increased channel sinuosity, which together reduce flow velocities and increase sediment capture. Early colonizing shrub species are now facilitating establishment of native herbs over annual exotics along with increases in rare species that were reduced to low numbers by trampling by non-native animals. Clear

Stewardship, Conservation, and Restor ation   933

goals and working assumptions have been critical to the success of the Channel Islands restoration approach. Land managers recognized that some change is inevitable, and some human-caused changes are irreversible, alleviating expectations that the islands will return to a particular historical baseline. They also acknowledge that restoration outcomes depend on initial ecosystem and weather conditions (precipitation in particular) and that self-repair might be slow, especially for soil development and other processes that operate on longer time scales. Finally, the importance of local, regional, and global contexts for the recovery of wide-ranging species on the islands has been clearly articulated, clarifying the role of the Channel Islands in broader, mutually contingent efforts to conserve these taxa. Restoration is not only essential for maintaining many of California’s ecosystems but also for enhancing quality of life; it can reconnect people with nature and improve air and water quality. Restoration of degraded habitat and conservation of intact habitats are both essential, with the latter in part providing the source propagules for future biodiversity. Restoration requires source material, and its success is greatly influenced by the configuration of natural areas and their composition present in the surrounding landscape. Scientists increasingly appreciate the value of participatory research and collaborate with land stewards, managers, and citizen naturalists to implement realistic restoration and management goals for most of California’s wildlands.

Future Directions for Stewardship and Policy across California’s Ecosystems Interwoven in many of the chapters in this volume are important suggestions for land management, research directions, and revisions to policy. These focus largely on increasing ecosystem resilience to help California’s ecosystems adapt to global environmental changes. This focus is critical to minimize future regrets and provide desired outcomes and benefits no matter how climate and other changes proceed, to provide environmental co-benefits including the alleviation of potential threats to human communities associated with climate and other environmental changes, to dynamically restoring natural processes, and to address the areas of greatest risk in the state. With this in mind we review and augment the calls to action presented throughout this book for the benefit of California’s ecosystems and the people who inhabit them. We emphasize the importance of fire management, managing for natural hydrologic flow regimes, emerging challenges to open space stewardship, and ecological monitoring as essential elements of adaptive management that each apply to a diversity of ecosystem types in the state.

Fire Management Many of the authors in this book point to the challenges and importance of fire as a disturbance agent integral to California’s plant community dynamics and an essential management tool for replicating natural ecological processes. Native Californians used fire to drive and concentrate game, open paths for travel, alter habitat mosaics, protect against enemies, safeguard villages, and promote the growth of desirable plants and specific plant parts (see Chapter 10, “Indigenous California”) (DeNevers et al. 2013). These prescribed burns 934   Policy and Stewardship

certainly enhanced habitat diversity in some parts of California; however, the influence of prescribed burning on plant communities by indigenous people may be less widespread than is sometimes asserted. From the forests to the deserts, authors stress that the combined effects of warming, decreases in rainfall, increases in the extent and intensity of human land use, and atmospheric changes due to pollution could increase fuel loads and fire frequencies, which in turn could greatly alter plant community composition. In particular, fire suppression over the past century has produced dense, even-aged forests that, in combination with increased human ignitions and increased tree mortality from a variety of sources such as beetle kill, present a “perfect storm” for catastrophic fire (see Chapter 3, “Fire as an Ecosystem Process”). Most models predict that the frequency and severity of fire will rise, with 5–​8% increases in area burned (see Chapter 27, “Montane Forests”). The chaparral ecosystem, characterized by drought-tolerant, fire-adapted species, is expected to expand as conditions warm and fire intervals increase. California’s already extensive urban-wildland interface in chaparral ecosystems continues to grow; damage to structure and loss of life could get much worse without immediate attention to land use policies that encourage expansion of development into these wildlands (see Chapter 24, “Chaparral”). Montane forests face similar challenges. Changing fire regimes across the state could be the most important consequence for California’s coupled human and natural systems of changing climate and continued sprawl. Ecological restoration requires fire in many ecosystems, but many issues lamentably limit or prevent use of prescribed burns. These include potential threats from escaped fires to structures and lives, air quality restrictions, and a lack of resources to conduct controlled burns. Alternatives to fire such as mechanical tree or shrub removal can benefit tree stand management and reduce risk of future fires, but these alternatives do not restore important ecological processes such as nutrient cycling that accompany fire. Prescribed fire is described as the most efficient means to promote montane forest resilience, especially in the face of climatic stress (see Chapter 27, “Montane Forests”).

Freshwater Landscapes Water is one of California’s most important natural resources. A central need to improve conditions in several important California ecosystems is to address past and continued hydrological alterations and improve freshwater conservation for environmental processes, functions, and species. Water has long been a critical resource in California whose natural capture and delivery have been impacted by dams and other water management projects. Many threatened and endangered species in the state rely on freshwater ecosystems. Of the California species that went extinct in the past hundred years, 19% occupied freshwater systems (see Chapter 11, “Biodiversity”), and close to 80% of California’s 129 native freshwater fishes are at risk of becoming extinct (Moyle et al. 2011; see Chapter 33, “Rivers”). A remarkable 82% of California’s native fish species are considered highly vulnerable to climate change (Moyle et al. 2013). Conflicts surrounding water management are at the heart of many divisive political agendas, contrasting cultural and ecological values. It is hard to imagine how California’s ecosystems, especially those highly dependent on dynamic hydrologic processes, will fare

as the state’s population approaches fifty million and drought increases. For example, wetlands are highly sensitive to alterations throughout both the terrestrial and the aquatic components of watersheds. Landscape configuration and seasonal flow conditions affect wetlands; thus watershed alterations that create hydrologic rerouting and change flow dynamics can greatly affect wetland processes and lead to dysfunction and species extirpation. Wetland features technically protected under the Clean Water Act can be severely degraded by hydrological alterations caused by permitted, adjacent, or nearby development. The same can be said for riparian communities that rely on seasonal streamflow variation. For example, germination of Fremont cottonwood (Populus fremontii) in riparian areas depends on the spring and early summer inundation associated with floods from Sierran snowmelt (Mahoney and Rood 1998). These large, overstory trees provide vital habitat and shade for salmon and other freshwater species. The magnitude, frequency, and duration of flooding have been widely altered by reservoir management specifically to avoid flooding and to provide a more constant water supply. The problems of addressing wetland management independent from the management of surrounding, often highly modified, landscapes has crippled many ecosystems (see Chapter 31, “Wetlands”). Managing wetlands and fresh water at a larger scale requires balancing socioeconomic demands on the system, often expressed as political influence, with protecting the dynamic hydrologic processes required for ecosystem function and self-maintenance. California’s water policies are completely inadequate to structure effective allocation of this resource (see Chapter 40, “Land Use Regulation for Resource Conversation”). This fact has severely hampered attempts to protect streamflows for environmental benefits including endangered species recovery. In rare instances such as the Mono Lake case (see Chapter 32, “Lakes”), fresh water was returned to an ecosystem for public trust benefits. However, other fresh waters such as coastal streams around the San Francisco Bay have been the subjects of long legal struggles pitting urban, agricultural, and environmental interests against each other without clear resolution.

Protected Area Challenges Many of California’s parks and preserves face a growing number of management challenges. A renewed commitment and stewardship to these precious public lands is needed. The recent establishment of marine protected areas (MPA) along California is one of the most promising conservation actions recently taken in the region. These reserves are intended to provide refuge from fishing, creating core habitat areas for fish and other marine species to reproduce and grow, and providing a source for replenishment of currently overharvested areas. Performance of marine protected areas is rather mixed, mostly due to a lack of enforcement and monitoring. This underscores the importance of building capacity to better manage MPAs and to effectively manage and evaluate their strategies and actions (Pomeroy et al. 2005). Effective marine protected area management along with implementation of the catch shares or individual transferable quotas in fisheries management can go a long way towards restoring both marine biodiversity, habitats, and productive fisheries (Costello et al. 2008).

While large areas of California’s public lands are designated as parks, preserves, and wilderness areas, activities still occur within these lands that can degrade ecosystems. Demand continues to grow for multiple-use considerations in California’s open spaces, and particular tensions surround off-road vehicle use. The impacts on natural ecosystems of rapidly growing recreation can be extensive yet remain understudied and underappreciated. In the United States the popularity of outdoor recreation activities—​such as hiking, backpacking, and birdwatching—​has doubled in the past twenty years (Cordell et al. 2005). In California a majority of residents (65%) prefer to recreate in undeveloped and nature-oriented parks, a trend that has increased over the past fifteen years. Public forested lands in California receive an average annual total of fifty million to sixty million visitor-days. Yosemite National Park received four million visitors in 2010; while not all activities equally affect ecological systems, these sheer numbers present concerns for montane forest health. On the other hand, access for recreation is a key component of plans to generate public support and revenue for land conservation. Public parks and open space preserves are the primary places where most people access nature, and contact with nature has a range of human health benefits (Frumkin 2001). Ecologists have identified recreation as an ecosystem service supporting human populations (Costanza et al. 1997), estimating the monetary value of recreation opportunities that are provided when land is protected from development (Chan et al. 2006). The Lake Tahoe Basin has a tourist population of 4.5 million yr-1, and with revenues of $4.7 billion annually, tourism is at the heart of the region’s economy (see Chapter 32, “Lakes”). Outdoor recreation enthusiasts are vocal advocates for land conservation, and public access is an important platform for generating tax and bond revenue for protected area acquisition. For example, California voters passed Proposition 84 in 2006, which as part of over $5 billion in bond commitments included nearly $2.5 billion for conservation ranging from park facilities to forest and coastal protection. In the same year voters passed Measure F in Sonoma County, extending a quarter-cent sales tax in the county entirely dedicated to open space protection. Still, recreation is not always compatible with other conservation objectives. Recreational activities are the second largest cause of endangerment to species occurring on U.S. federal lands (Losos et al. 1995). California is the state with the greatest number of listed species threatened by recreation. This is in part because the threat of recreation is most frequently associated with urbanization, another important cause of endangerment (Czech et al. 2000). Growing evidence indicates that nonmotorized activities have negative impacts on a wide range of wildlife species (Knight and Gutzwiller 1995). For example, recreational activity correlates with decreases in species abundances and activity levels (Garber and Burger 1995), causes wildlife to flee (Papouchis et al. 2001) or avoid otherwise suitable habitat (Taylor and Knight 2003), and alters species composition and behavior (Ikuta and Blumstein 2003). Reed and Merenlender (2008) found that the key factors associated with recreational effects on carnivores appear to be the presence and number of human visitors to protected areas in California oak woodlands. In this study abundances of native coyotes (Canis latrans) and bobcats (Lynx rufus) were over four times greater in sites not open for public access than in adjacent park lands. Expanding human communities demand additional land and public access for recreation, underscoring the need for refined understanding Stewardship, Conservation, and Restor ation   935

of the consequences of recreation for ecosystems and biodiversity that can guide co-development of management strategies with park managers to minimize these impacts while providing opportunities for people to experience nature in a diversity of ways.

Ecological Monitoring for Adaptive Management Active ecosystem stewardship requires increased attention to ecological monitoring in California. Monitoring is integral to adaptive management and is stressed by authors throughout this volume as a vital and undersupported activity. Only through monitoring biodiversity is early detection of invasive species possible and equally important is impact monitoring to assess the effectiveness of eradication efforts, which generally require ongoing efforts to achieve long-term control (see Chapter 13, “Biological Invasions”). For instance, multimillion-dollar restoration efforts in the San Francisco Bay-Delta and elsewhere to recover estuaries and wetlands in general still lack effective monitoring and evaluation programs to measure the environmental and social benefits of these efforts and to determine whether management or policy changes are required to improve success (see Chapter 19, “Estuaries: Life on the Edge”). Effective management requires clear identification of monitoring objectives, stressors that contribute to ecosystem degradation, and indicators that can be used to monitor ecosystem responses to change (see Chapter 30, “Deserts”). Monitoring is especially critical given the uncertainties associated with increased rates of change. In some systems, such as oak woodlands, the dominance of private ownership and growing prevalence of small individual land parcels pose special challenges (see Chapter 25, “Oak Woodlands”). These include lack of landowner experience, training, financial incentives, and general willingness to participate and mean that improving stewardship in these privately owned landscapes can only be done with outreach and education. Remote monitoring technologies are also increasingly important tools to assess the condition, composition, and changes that will occur on woodlands and other privately owned landscapes. While monitoring protocols are well developed for some systems, such as wetlands and watersheds, less information exists in other ecosystems to build on to inform monitoring programs. For example, reliable methods for restoration and monitoring under different environmental contexts are largely undeveloped for desert systems (see Chapter 30, “Deserts”). Likewise, effective management of newly established marine protected areas requires monitoring, evaluation, communication, and adaptation. Evaluation involves quantitative tracking of ecosystems before and after management actions are taken and assessing whether these actions are producing the desired outcomes. Evaluation is a routine part of the management process yet is underdeveloped for most managed ecosystems. Even in places that are rebounding and might require less new action, such as the Channel Islands, monitoring remains crucial for detecting changes over time. Beyond resources and will, monitoring and assessing management effectiveness rely on sound, scientifically based design. Much can be learned from well-documented experiments designed to test hypotheses and provide information for natural resource management. Research forests and field stations exist throughout California that could be used

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to compare management alternatives as well as to monitor global change within and between the ecosystems that span the state. Intensive climate monitoring efforts are under way on sites such as Blue Oak Ranch Reserve, a University of California reserve in the Mount Hamilton Range, Santa Clara County, California (Blue Oak Ranch 2008). Here, a wireless environmental monitoring system has been deployed to collect fine-resolution data on spatial and temporal variability in the landscape. This type of passive monitoring system provides an excellent opportunity to understand how environmental changes impact biophysical processes and functions and, ultimately, organismal and community responses. This type of intensive, fine-scale monitoring provides an important complement to the more coarse-grained monitoring feasible over larger spatial extents.

Policy Shifting and expanding stewardship goals will create an enormous challenge for business-as-usual conservation policy making. Existing policies (e.g., the National Environmental Protection Act [NEPA], the California Environmental Quality Act [CEQA], the federal Endangered Species Act [ESA], and the California ESA [CESA]) and their current implementation offer few options for future flexibility and adaptation. There is a critical need to analyze existing policy tools and define future policy needs from four perspectives: (1) the extent to which existing state and federal policies hinder the reevaluation of conservation goals in the context of directional change, (2) the degree to which existing state and federal laws across multiple sectors hinder or facilitate implementation of conservation goals, (3) whether existing state and federal programs that provide financial resources are aligned to help conservation adapt to directional change, and (4) the extent to which tools for state agencies, local governments, and nonprofit organizations are available or under development to facilitate conservation implementation in the context of regional change. The California Natural Resources Agency has made statewide adaptation planning for climate change a priority for addressing complex and large-scale challenges to conserving biodiversity and habitats (California Climate Adaptation Strategy for Biodiversity and Habitat 2009). Initial planning efforts focus on helping species persist in a changing environment. Towards this goal, the state agencies responsible for stewarding the state’s biological diversity, the Department of Fish and Wildlife and California State Parks, have committed to evaluate internal policies related to regulatory responsibilities and to communicate openly with other agencies and the public. In their adaptation plan, these agencies outline nearand long-term climate adaptation strategies that will require additional collaborative efforts with multiple state agencies as well as sustainable funding and long-term state support. The California Global Warming Solutions Act of 2006, AB 32, SB 375, and the Sustainable Communities and Climate Protection Act of 2008 set statewide and regional targets for GHG emissions reductions and demonstrate a willingness to address climate change in California. While deep structural change is needed to address the environmental improvements needed in California, demands for mitigation and conservation protection that are emerging to support greenhouse gas emission reductions could provide a way forward through

climate change policy (see Chapter 40, “Land Use Regulation for Resource Conservation”). Climate mitigation funds and efforts should be directed or incentivized to protect the conservation priorities outlined in existing regional and state conservation plans. Climate mitigation needs to avoid restoration projects that fail to deliver functioning ecosystems and conservation of small, unconnected mitigation sites and projects disconnected from existing conservation priorities and unlikely to contribute lasting conservation value. The state needs to expand the approach used by regional transportation plans that fully assess cumulative and site impacts of development and that aggregate proposed greenhouse gas offsets and direct them toward meaningful conservation investments. Equally important, counties need resources to improve urban services and incentives to direct development toward infill and other suitable areas in order to avoid habitat conversion and fragmentation. We focus on aligning local land use planning and state conservation with the changing needs of the state’s biological resources, but the state will require support from academic and other science institutions to rethink policy and implementation in the context of the changes we are now experiencing.

Future Research, Education, and Outreach for Addressing Environmental Change Fundamental to future actions is the need for sound science and adaptive, effective management, along with knowledge sharing and comprehensive and sustainable data collection and monitoring systems. We bring the following areas of investigation into focus because of their importance in improving our understanding of how to best steward California’s ecosystems given future environmental change and the increased uncertainty that this presents for biodiversity conservation and ecosystem services. Most of these priorities span more than one ecosystem, and all reflect the scholarship contained in this volume. The following specific areas are loosely divided into biophysical research, interdisciplinary and applied research, and public participatory efforts.

Biophysical Research Understanding interactions among climate conditions, ecosystem processes, and vegetation dynamics is the foundation for the stewardship approach we propose as a framework for ecosystem conservation in California. The impacts of climate change will be mediated largely through extreme physical events, including drought, wildfire, heat waves, loss of chilling hours, and relaxation of freezing events. Continued efforts to enhance and validate downscaling methods, coupled with enhanced understanding of biological responses at all levels to changing abiotic and biotic conditions, will be important for bringing the large-scale projections of GCMs to the local scale of organismal and population responses (e.g., Flint et al. 2013). Large-scale tree mortality due to heat or drought has already been observed on all continents (Allen et al. 2010), including increased mortality of large trees in California and the American West (van Mantgem et al. 2009). In California’s Mediterranean-type climate, changes in pre-

cipitation and water balance will in many cases be more important than increased temperatures, leading to heterogeneous responses across complex gradients and microclimates (Crimmins et al. 2011, Rapacciuolo et al. in review). Specific research needs that emphasize climate change as an emerging driver include:

. Food webs and species interactions, including the

. . .

. . .

.

. . .

top-down roles of meso- and top predators such as mountain lions, bobcats, and coyotes in community dynamics; the bottom-up impacts of invasive plants on animal community composition; and mechanisms controlling community invasibility. Disturbance regimes and successional dynamics in nonstationary environments, including identification of leading and trailing edges of existing ecosystems and emergence of novel ecosystems. Interactions between climate, including extreme drought events, and disease. Ecosystem inventory to identify remnant native grasslands; species assemblages associated with hot, dry microclimates within each ecosystem type to pinpoint likely beneficiaries of future change; species assemblages in cool climate areas across each ecosystem type that are not likely to change; and potential refugia in need of protection. Paleoecology and historical ecology to provide historical context on recovery and dynamics of vegetation change across ecosystems and historical responses to drought. Improved understanding of current and likely future fog patterns and how coastal microclimates and fogreliant communities respond to changes in fog timing and extent. Ecophysiology and climate change biology of California’s dominant woody plant species, including climate and microclimate controls on seedling establishment, early growth, and mortality; and patterns and scales of genetic climate adaptation. Ecohydrology, including interactions among major plant growth forms, associated spatio-temporal patterns of water availability, and spatial and temporal variation in soil moisture available across complex geologies and landforms. Belowground processes, including the influence of water deficit on microbial processes; and root function and climate sensitivity. Microevolutionary dynamics and rapid evolution in relation to dispersal rates and gene flow in spatially heterogeneous environments. Carbon, water, and energy exchange of California’s ecosystems, including rates of carbon sequestration.

Interdisciplinary and Applied Research While many of our identified priorities for biological research focus on mechanisms of response to climate, research on and experimental tests of conservation and climate adaptation strategies and their effectiveness are equally needed. Conservation and restoration biology lead the way in this regard. The urgency to intervene, conserve threatened species, and improve degraded systems should not preclude capitalizing

Stewardship, Conservation, and Restor ation   937

biodiversity, prioritizing physiographic landscape features versus habitat types; prioritizing for conservation areas with lower climate water deficit; improving habitat connectivity through increasing reserve size and number versus corridor conservation or diversifying moderately modified habitat surrounding reserves; and enhancing corridors versus facilitating broader movement through matrix lands.

Ecological Monitoring and Public Participation

FIGURE 41.6 Inland Mendocino California Naturalists monitor a nest cavity at the Hopland Research and Extension Center as part of NestWatch , a citizen science project. Photo: Brook Gamble.

on opportunities to conduct long-term studies and controlled, experimental research on intervention effectiveness. Truly interdisciplinary research, designed from the outset by collaborations of physical, biological, and social scientists and nonacademic partners, requires significant investments to build professional relationships and bridge intellectual domains. However, such collaborations hold great potential to address conservation and climate impacts; design and evaluate solutions; and identify the costs of, barriers to, and opportunities for implementation. Priorities for interdisciplinary conservation science include:

. Experimental study of how systems respond to inter-

. . .

. .

.

ventions such as invasive animal or plant removal, prescribed burning, and other actions that could enhance ecosystem recovery and resilience, where possible with controls, replication, and time depth. Retrospective analysis of outcomes of past actions, including restoration projects, effectiveness of mitigation banks and projects, and conservation easements. Study of recreation impacts, including risk analyses for threatened species in high-visitation areas, changes in animal movement patterns to avoid people, and effect zones around trails. Integrated, large-scale models examining relationships among development patterns, fire regime, and ecosystem resilience; effects of low-density residential housing on community composition and ecosystem processes; and biophysical dynamics such as through models that link climate and vegetation structure and dynamics to transfers of energy, water, carbon, and other nutrients. Risk assessment of the potential for and consequences of change for rare habitat types. Valuation of regulating, supporting, and amenity services flowing from natural ecosystems and design of markets to incentivize private landowners to conserve or increase provision of ecosystem services on their properties. Conservation planning and trade-off analyses of the implications of targeting ecosystem services versus

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The need has never been greater to implement broad networks of monitoring and early detection sites and to acquire baseline data in the short term. Increasingly, advances in remote sensing will provide more complete coverage with enhanced spatial and temporal resolution for many variables of conservation interest. Ecological monitoring efforts nationwide span the gamut from the multimillion-dollar National Environmental Observatory Network (NEON) to dispersed citizenscience initiatives such as the National Phenology Network and biodiversity observations collected by the public through systems such as iNaturalist.org. Hundreds of active citizenscience programs exist in California, with details for each provided through the UC California Naturalist Program website (UC California Naturalist Program 2013) (Figure 41.6). Trade-offs in investment, data consistency, and long-term reliability must be addressed in each case to identify the optimal balance of dispersed, low-cast networks and more concentrated sites with institutional support. Rapid data processing and availability are critical for adaptive management, especially given rapid rates of change. Local Cooperative Weed Management Areas also demonstrate the power of ecological monitoring by volunteer stewards to increase effectiveness of invasive species detection and control while reducing its cost (see Chapter 13, “Biological Invasions”). Preventing mass extinction will require comprehensive science, adaptive management, and a committed public. Priorities for advancing ecological monitoring and public participation include:

. A statewide, open-access database on restoration out. .

.

.

comes monitoring to foster evidence-based restoration and management actions. Networks of “sentinel” sites spanning bioregions and ecosystems across the state, providing continuous monitoring and early detection of changes. Participatory research through improved usability of existing mobile and online applications for crowdsourced data acquisition; improved detection of local species declines and extirpations; and new statistical approaches to weight-contributed data based on estimates of reliability (e.g., bird list length). Citizen monitoring of early detection of new arrivals and diseases; changes in flowering timing, seed set, and other life-cycle events in relation to environmental cues; citizen rephotography of historic photo points; and citizen supervision of monitoring instruments. Education of small-parcel landowners about sustainable land management practices, including ways to maintain appropriate grazing levels, enhance riparian vegetation and woody debris, control invasive species, manage pests, and minimize movement barriers such as fencing.

Summary Given continued rates of land use change, increasing influence of climate change, and the transient nature of California’s ecosystems, it is time to reconsider some of our current approaches to conservation, stewardship, and restoration. Time lags between our actions today and species and ecosystem responses in the future make even more critical efforts to anticipate and conserve now. We need renewed, immediate protection against habitat loss and fragmentation from continued urban and agricultural development throughout California. Given increased rates of change and the resulting widespread existence of novel communities, we must adopt and pursue novel approaches to restoration and increase investment in stewardship. Ecological restoration efforts should acknowledge ecosystem disequilibrium associated with directional global changes, take advantage of information on historical ecology and consider human needs to set achievable goals and produce realistic outcomes. A new stewardship framework founded in resilience to guide future interventions will help maintain ecosystem integrity over a large range of variability in climatic conditions and perturbation frequency and intensity. Many successful interventions described throughout this book make clear that we have substantial successes to build on, learn from, and advance further stewardship efforts. Based on the concerns and strategies raised in many of the ecosystem chapters, we emphasize the importance for California of managing fire and natural hydrologic flow regimes, tackling emerging challenges to open space stewardship, and ecological monitoring as part of this stewardship framework. Advancing our understanding of California’s ecosystems and their responses to environmental changes, primarily mediated through extreme physical events, is at the heart of advancing our stewardship goals. This requires a stronger state funding stream for targeted biophysical research on ecosystem dynamics and applied, interdisciplinary research on environmental disturbances, restoration, and management. Equally important, we must implement monitoring networks within and across California’s ecosystems and take advantage of new technologies as well as public participation to be successful in this endeavor. New stewardship strategies and research should not overshadow the need to protect California’s ecosystems. An emerging part of this is the growing need for climate change adaptation and resilience, particularly as efforts to prepare other systems for climate change—​ such as California’s water storage and delivery—​a ffect ecological systems. The state should set clear targets that meet or exceed international standards for conservation and restoration of every ecosystem type. This type of conservation planning and ecological restoration needs to be coupled with land use planning and considerations for carbon dioxide sequestration. In sum, coping with rapid environmental changes requires increased focus on strategic conservation planning, collaborative stewardship efforts, new approaches to restoration that improve the returns on our investment, and an integrated, targeted research and extension agenda.

Acknowledgments We would like to acknowledge Cris Benton, Dylan Chapple, Greg Damron, Frank Davis, Brook Gamble, Kerry Heise, and

Heather Rustigian-Romsos for assistance with some figures and photographs that help to bring this chapter alive.

Recommended Reading Barnosky, A. 2009. Heatstroke. Island Press, Washington, D.C. Chapin, F. S., III, G. Kofinas and C. Folke, editors. 2009. Principles of ecosystem stewardship. Springer. New York. Cole, D. and L. Yung, editors. 2010. Beyond naturalness: Rethinking park and wilderness stewardship in an era of rapid change. Island Press, Washington, D.C. De Nevers, G., D. S. Edelman, and A. M. Merenlender. 2013. The California naturalist handbook. University of California Press, Berkeley, California. Hobbs, R., and E. Higgs, editors. 2013. Novel ecosystems. Wiley– Blackwell: A John Wiley & Sons Ltd. Oxford, UK. Hobbs, R., and K. Suding, editors. 2009. New models for ecosystem dynamics. Island Press, Washington, D.C. Interdisciplinary Research Team Summaries. 2011. Ecosystem Services: Charting a Path to Sustainability. The National Academies Keck Futures Initiative. Arnold and Mabel Beckman Center, Irvine, California, November 10–11, 2011. Jensen, D. B., M. S. Torn, and J. Harte. 1993. In our own hands: A strategy for conserving California’s biological diversity. University of California Press, Berkeley, California.

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INDEX

A horizons, 52, 55, 56, 58, 65, 67 abalones, 318, 320–22, 326, 337, 343, 350, 374, 767, 769, 784, 797, 798, 800 black abalone, 320, 344, 769, 797 red abalone, 320, 350, 786, 790 white abalone, 195, 196, 769, 796, 804 Abatzoglou, John T., 17 Abies, 138, 482f Abies amabilis, 537t Abies bracteata, 221–22 Abies concolor, 32, 221, 537, 557, 580, 817 Abies grandis, 218, 536 Abies lasiocarpa, 222, 557, 580 Abies magnifica, 145, 554, 580 Abies magnifica var. shastensis, 537 Abies procera, 222 abiotic shift, 557 aboveground net primary production (ANPP), 456f Abronia latifolia, 409f, 414f Abronia maritima, 418 Abronia umbellata, 414 Acanthinucella, 344 Acanthomysis bowmanii. See Hyperacanthomysis longirostris Acanthomysis spp., 397 Acartia spp., 371 Acartiella sinensis, 371 Acaulospora, 484 Accipiter cooperii, 488t Accipiter gentilis, 37, 561 Accipiter striatus, 488t accretion. See erosion and accretion cycles Acer, 139, 536, 542f Acer macrophyllum, 237, 512, 541, 559 Acer macrophyllum Pursh, 721 Acer negundo, 512 achenes, 417 Achillea, 458 Achillea millefolium, 416 Achnatherum hymenoides, 648, 649f acid neutralizing capacity (ANC), 696–97 acidification. See ocean acidification Acipenser medirostris, 379t, 728 Acipenser transmontanus, 798 Acipenseridae, 727f acmispon, Nuttall’s, 414 Acmispon glaber, 433t, 434, 437, 439, 496 Acmispon prostratus, 414

acorn barnacles, 352 acorn caches, 175 acorn feeder, 515 acorn woodpecker, 514, 516 acorns, 514, 841 acridid grasshoppers, 191 Actinemys marmorata, 680, 727 actinomycetes, 644 actual evapotranspiration (AET), 10, 11f, 510 Adams, Ansel, 705 adaptive governance, 889 adaptive management, ecological monitoring for, 936 adder’s tongue, California fettid, 542 Adenostoma sp., 479, 484t Adenostoma fasciculatum, 56, 220, 479f, 480, 482f, 484t, 490f Adenostoma sparsifolium, 479f, 483 adiabatic heating, 110 adiabatic warming, 12 adsorption, 682 advection, 14 Aechmorphous clarkia, 372 Aegilops triuncialis, 454, 850 aerobic waters, 370 aerosols, 20, 108, 118–20, 413, 696 Aesculus californica, 218, 512 Aesculus spp., 133f Agabus sp., 698f Agarum fimbriatum, 314, 315f Agavaceae, 654 Agave deserti, 254, 647 agaves, 169, 175 desert agave, 254, 647 Agelaius tricolor, 680 aggregating anemones, 352 Agonidae, 322 agricultural and ecosystem regions of California, defining the, 875–76 agricultural development drivers of, 869, 882 waves of, 870, 871f agricultural ecosystem of California, 865, 882–83. See also agroecosystem economic malfunctions, 877–79 functions, 876–77 synergies and trade-offs, 881 agricultural interactions with natural ecosystems, 880

agricultural production in California, 271, 866 structure of organic, 877–78 agricultural services environmental services and, 881 pollination, pest control, and supporting, 271–75 agricultural valleys, 89–90 agriculture, California, 523, 865, 882–83 climate, agroecology, and, 866 constants that have driven, 869 factors that have contributed to the success of, 870–71 farm size and, 866–67 labor and, 877 land and, 866 output of various categories of, 868, 873 scope of, 867–68 transformations in, 872–73 water, irrigation, and, 866 agriculture history, California, 767, 868–69 agriculture at start of 20th century, 870 first half of 20th century Depression and World War II (19301949), 871–72 extensive to intensive agriculture (1890-1930), 870–71 pre-20th century agriculture following California Gold Rush, 521–22, 869 gold, cattle, statehood, and growth, 1848-1860s, 869–70 sheep, wheat, and early horticulture (1860s-1890s), 870 Spanish-Mexican period (1760-1848), 869 second half of 20th century to the present, 872 big water, growth, relocation, and diversification (1950-1970), 872–74 ups and downs, intensification, and internationalization (1970-2010), 874 into 21st century, 874–75 agroecosystem, California’s. See also agricultural ecosystem of California future of, 881–82 Aimophila ruficeps, 439, 488t 943

air pollutants, 117–18. See also specific pollutants affecting terrestrial ecosystems, 108, 109f human health effects, 120–21 air pollution, 122–23. See also nitrogen (N) air pollution and atmospheric deposition; pollution conservation, adaptation, and mitigation, 121–22 interactive effects of climate change and, 120 air pollution emissions, changes in, 109f air pollution transport, statewide atmospheric circulation patterns and, 110–12 air quality fire effects on, 119–20 temporal and spatial patterns of, 108–10 Air Quality Management Districts (AQMDs), 112 air quality standards, 121, 122t Aira, 461 airsheds, 694 Alameda whipsnake, 198 Alaska yellow-cedar, 222, 580, 593, 597 albacore tuna, 100, 292, 293, 790f, 791f, 794f, 795 albatrosses, 293 black-footed albatross, 293f, 294, 296 Laysan albatross, 294 albedo, 20, 119 alderflies, 698f, 723f alders, 139, 145, 187f, 588, 702 red alder, 541, 678f, 679 white alder, 512, 721 Aleochara spp., 399 Aleutian cackling goose, 200 Aleutian Low (AL), 95, 98–99 Alexornis, 199 alfalfa, 231, 876 Alfisols, 50, 55f, 57–60, 65, 66, 555 algae, 314, 315f, 317, 318, 349, 368–69, 377, 716–17, 721–22. See also specific topics defined, 716 types of, 314, 315f, 717. See also kelp; specific types of algae alien species, 229, 230. See also invasive species; species: non-native alkali fly, 194, 701 alkalinity, 703 Allen, Edith B., 435, 437 Allen’s hummingbird, 765 alligator lizards, 198, 459 Allium, 192 allocthonus, 320, 370, 396 Alloniscus perconvexus, 397f, 398 Allophyllum gilioides, 193 allozymes, 420 alluvial bottomlands, 675 alluvial fans, 49, 52, 642 weakly developing soils on broad, 52, 53f alluvium, 51–52 weakly developing soils in basin, 52, 53f almonds, 89, 272, 866, 867, 873, 876 Alnus, 139 Alnus spp., 588 Alnus incana tenuifolia, 702 Alnus rhombifolia, 512, 721 Alnus rubra, 541, 678f, 679 Alopias vulpinus, 293, 795 Alpert, Peter, 413 alpine chipmunk, 164, 206t, 594, 596, 627 alpine ecosystems, 613, 630 climate adaptation, 629–30 climate regimes and abiotic stress, 617, 619 conservation, 629

944  INDEX

defining, 613–14 geographic distribution, 614–17 geologic and geomorphic setting glacial and periglacial history, 619–20 historical geology of uplift and erosion (subsidence), 619 geomorphic settings and habitats, 620 mountain summits and upland alpine plateaus, 620 human interactions, 628 invasive species, 629 mammals species, 627t processes and ecosystem dynamics, 622, 624 upland slopes and basins, 620 broken rock habitats, 621 glaciers and permanent snowfields, 622 patterned ground, 621–22 wetlands, 622, 623f vegetation and flora, 624 evolution of flora, 625–26 fauna, 626–28 floristics diversity and phylogenetic branch and depth, 625 plant functional groups and adaptive traits, 624–25 alpine fellfields, 617f altitudinal shifts. See mountainous areas, altitudinal shifts in amanitas, 483 ambrosia beetle, oak, 237 Ambrosia chamissonis, 414f Ambrosia deltoidea, 642 Ambrosia dumosa, 142, 225, 642, 649f Ambystoma californiense, 198, 232, 465, 675, 848 Ambystoma macrodactylum croceum, 198 Ambystoma mavortium, 198, 232 Ambystoma spp., 459 Ambystomatidae, 198, 727f Ameiurus spp., 726 Amelanchier utahensis, 594 Ameletus edmundsi, 698f American avocet, 202, 685 American badger, 516. See also badgers American beach grass, 413, 422 American bison, 760t, 763 American black oystercatcher, 400 American bullfrog, 199, 680, 727, 732 American coot, 680 American dune grass, 412, 413 American kestrel, 488t American marten, 204, 628 American pika, 207, 594, 596, 621, 626–28 American pipit, 201, 400 American pronghorn, 657 American restart, 200 American searocket, 416, 419 American shad, 373 American white pelican, 680, 704 Ammondramus savannarum, 459 ammonia (NH3), 109t, 115 distribution in southern Sierra Nevada, 115, 116 Ammophila arenaria, 412, 413f, 419f Ammophila breviligulata, 413 Ammospermophilus leucurus, 657 Ammospermophilus nelsoni, 206t, 847 amphibians. See also specific topics biogeography, 197–98 conservation context, 198–99 in deserts, 653, 656 evolutionary diversification, 196–97 riverine food webs and, 726–27 wetlands and, 680 Amphipoda, 723f

amphipods, 269, 289f, 291, 317–18, 320, 322, 328, 343, 368, 371, 395, 398–401, 641, 717, 720, 722, 723f Amphispiza bilineata, 656 Ampithoe humeralis, 269 Ampithoe valida, 371 Amsinckia menziesii, 433t Anabaena, 720 anabatic winds, 110 Anacapa deer mouse, 240, 770 Anacapa Island, 240, 756, 758t–760t, 762– 63, 764t, 766t, 770, 771, 773 Anacardiaceae, 491 anadromous fish, 724 anadromous white trout, 792 anaerobic bacteria, 682 anaerobic decomposition, 373 Anarrhichthys, 319f Anarrhichthys ocellatus, 322 Anas acuta, 371, 680 Anas clypeata, 371, 680 Anas cyanoptera, 680 Anas platyrhynchos, 400, 680 Anas strepera, 680 Anaxyrus boreas. See Bufo boreas Anaxyrus californicus, 726 Anaxyrus canorus, 628, 853 anchovies, 96, 101, 292, 293, 296, 321, 371, 791f northern anchovy, 788, 790f, 793 Anderson, K., 179 Anderson, M. K., 179, 180 Anderson, R. S., 766 Anderson’s thornbush, 648, 649f andic properties, 50 Andisols, 50, 58–60, 555 Andrenidae, 192 Andrew’s clintonia, 542 Andricus quercuscalifornicus, 515 Aneides ferreus, 541 Aneides vagrans, 539 anemones, 320, 321, 337, 343, 344, 352 angel shark, 795–96 angiosperms, 132, 138, 139, 158 Anguidae, 198 animal recovery, fire and, 37 Anna’s hummingbird, 438f, 488t Annelida, 343, 723f annelids, 235f and grassland distribution, 458–59 annexation, 913 Anniella, 198 Anniella spp., 415 Anniella pulchra, 418, 514 Anniellidae, 198, 415 annual grasses, 220 annual grassland, California. See valley grassland annual primary productivity, 698 Anoplopoma fimbria, 787 Anota, 198 anoxic sediments, 682 anoxic waters, 370 ant species diversity in deserts, precipitation and, 651, 652f antelope, 455 antelope bitterbrush, 647 antelope ground squirrel, 654, 657 antelope squirrel, San Joaquin, 206t, 847 Antennaria, 626 Anthopleura, 343 Anthozoa, 343 Anthus rubescens, 201, 400 Anthus spinoletta alticola, 628 Antilocapra americana, 522, 657, 841. See also pronghorn

Antioch Dunes shieldback katydid, 189t, 195 antlions, 650f ants, 418, 458, 515, 541, 644, 650f, 651, 652f, 654 Argentine ant, 238, 239, 770 harvester ant, 239 Anza-Borrego Desert State Park, desert scrub vegetation in, 225f Aonidiella aurantii, 239 apex predators, 323 Aphelocoma californica, 200, 488t, 514, 516 Aphelocoma coerulescens, 200 Aphelocoma insularis, 200, 764–65 aphids, 239, 272, 418, 486 Aphis gossypii, 239 Apis mellifera, 195, 767 Aplodontia rufa, 595t aplodontid rodents, 157f, 160 Aplodontidae, 595t aplodontids, 160 Aplysia, 319f Aplysia californica, 318 Apodemia mormo langei, 195 Apostichopus parvimensis, 320 Apostichopus spp., 320 apparent competition, 419 apples, 272 oak apple, 515 apricots, 89 Aptostichus simus, 415 aquatic garter snake, 680, 721 aquic moisture regime, 57 aquifers, 391, 878 Aquila chrysaetos, 628, 773 Aquilapollenites, 134 arachnids, 193f, 195, 651 araucarian, 134 arboreal primates, 158 arborescent cacti, 225 Arborimus pomo, 206t Arborvitae, 139 arbuscular mycorrhizae (AM), 437, 440, 441f, 459–60, 483–84, 540 arbuscular mycorrhizal (AM) endophyte, 437 Arbutus, 139 Arbutus menziesii, 218, 237, 480, 512, 539, 546f, 559, 829, 830 Arceuthobium spp., 599 Archaea, 369 Archaeohippus, 157f, 161 archaeological sites and isolated finds, terminal Pleistocene/early Holocene, 172f archaeology, reactionary and resource depression, 178–79 Archeomysis spp., 397 Archoplites interruptus, 671 Arco Giant, 543 Arctocephalus townsendi, 767, 792 Arctostaphylos, 139, 192, 220, 513 Arctostaphylos sp., 479, 484t Arctostaphylos andersonii, 482f, 829–30 Arctostaphylos crustacea, 482f Arctostaphylos glauca, 56, 482f Arctostaphylos myrtifolia, 482f Arctostaphylos nevadensis, 482f, 598 Arctostaphylos patula, 482f Arctostaphylos pungens, 56 Arctostaphylos sensitiva, 482f Arctostaphylos spp., 480 Arctostaphylos viscida, 482f Ardea alba, 371 Ardea herodias, 323 areas of special biological significance (ASBS), 350 Arenaria paludicola, 188

Areniscythris brachypteris, 415 Arenivaga investigata, 651 arêtes, sculpted, 140 Arfia, 156, 157f Argentine ant, 238, 239, 770 argillic horizons, 56–58, 66, 68 aridic soil mixture regime, 55 Aridisols, 55, 61 Ariolimax columbianus, 541 Arizona cypress, 139 Arlington Canyon, 769f Arnica, 626 Arroyo toad, 726 Artedius corallines, 317f Artemia, 701f Artemia monica, 194, 701 Artemisia, 37, 139, 144f, 429 Artemisia spp., 224, 642 Artemisia californica, 432f, 433, 435–37, 437f, 439, 439f, 440, 480, 764, 769f Artemisia cana, 849 Artemisia pycnocephala, 416 Artemisia tridentata, 36, 141, 636 Artemisiospiza belli, 201, 488t Arthropoda, 343, 723f arthropods, 37, 194, 195, 268, 418–21, 436, 458, 486, 515, 601, 621, 644, 648, 651–53 predacious, 650f artichoke thistle, 433t artichokes, 876 artiodactyl, 156 Artiodactyla, 203, 205f, 596t, 759t artiodactyls, 156, 157, 160–63 Artocarpus, 138 Arundo donax, 205, 682, 770 Asarum caudatum, 542 asbestos, 55 Ascaphidae, 197 Ascaphus, 726 Ascaphus truei, 197, 198, 726 Ascophyllum nodosum, 380 ashes, 139. See also mountain ashes Oregon ash, 512, 721 ashy storm-petrel, 202, 769, 770 Asian clam, 699 Asiatic clam, 722 asilids, 650f Asio flammeus, 201, 459 Aslan, C. E., 260 asparagus, 875 aspens, 588, 593, 594, 598, 603 assisted migration, 932 aster, 146 Asteraceae, 139, 189t, 414, 485t, 625, 768 Asteroidea, 320 asters, 139, 144 Astragalus lentiginosus var. borreganus, 652 at-a-station hydraulic geometry, 714 Athene cunicularia, 443, 459, 465, 655, 847 Atherinop affinis, 371 Atlantic oyster, 798 atmosphere and livestock production, 854 atmospheric chemistry, 107–8 atmospheric circulation patterns and air pollution transport, 110–12 atmospheric convection, 16 atmospheric forcing, 95 atmospheric nitrogen deposition, 109. See also nitrogen deposition atmospheric rivers (ARs), 14–15, 365, 555, 586, 617 Atractoscion nobilis, 790 Atriplex spp., 657 Atriplex canescens, 641 Atriplex confertifolia, 141, 641

Atriplex tularensis, 189t Atyidae, 189t auditory masking, 298 auklets, 293 Cassin’s auklet, 202, 293, 294, 770, 771 rhinoceros auklet, 294 Aurelia spp., 371 Aurora draytonii, 848 Australian vedalia ladybeetle, 242 autotrophy, 697 avalanches, 587–88 Avena spp., 771 Avena barbata, 238, 436, 452, 513 Avena fatua, 238, 432f, 433t, 513 avocado, 138, 139, 159 avocet, American, 202, 685 Axelrod, D. I., 191, 482 Aythya americana, 202 azalea, western, 541 B horizons, 49, 57, 67 Babbitt, Bruce, 918 baby boom, post-World War II, 80–81 “baby boom echo,” 81 Baccharis pilularis, 417, 435, 439, 513, 523, 769f, 772, 844. See also coyote brush/ coyote bush Bacillariophyta, 716 backswimmers, 681 bacteria, 39, 119, 266, 289, 290, 345, 368, 369, 458, 459, 484, 496t, 640, 643–45, 716, 722, 732, 853, 890 bacterial infections and diseases, 120, 320 badgers, 204, 459, 487t, 516 Baeolophus inornatus, 514 Baja California Islands, 758t Bakersfield smallscale, 189t Bakun, A., 99 Balaenoptera musculus, 203, 294, 792 Balanus, 343, 346 Balanus crenatus, 321 Balanus nubilus, 319f, 321 bald eagle, 188, 772, 773 Baldocchi, D. D., 524 baleen whales, 294, 297, 298 Bales, R. C., 706 Ballona cinquefoil, 190t banana slugs, 195 Pacific banana slug, 541, 543 band-tailed pigeon, 514 Bangasternus orientalis, 233 bank swallow, 724 bankfull, 714 barbary sheep, 760t barbed goatgrass, 461, 466, 850 barberry, 139 Barbour, Michael G., 410, 415, 555, 567 Barbourofelis, 157f, 161 Barchyn, T. E., 412 bark beetles, 37, 114, 194, 567, 569, 570, 599, 603 oak bark beetles, 237 bark borers, 515 barley yellow dwarf virus, 461 barleys, 89, 461, 867 barn owl, 459 barn swallow, 256t, 400 barnacles, 320, 321, 337, 342, 343, 346–49, 351, 352 Barnosky, A. D., 261 barnyard grass, 684 barracudas California barracuda, 323 Pacific barracuda, 795 barred owl, 200, 828–29 barred sand bass, 795

INDEX  945

barred surfperch, 399 barred tiger salamander, 198, 232 barrel cactus, 648 Barth, J. A., 101 basal area, 517, 560 basal burls. See lignotubers basalt, 55 basaltic lava, 49 base cations, 116 base saturation, 57 Basgall, M. E., 179 Basin and Range Province geology and geomorphology, 60–61 mountain soils, 63 piedmont soils, 63–64 playa soils, 61–63 basin margins, weakly developing soils on, 52, 53f basking shark, 293, 796 Bassariscus astutus, 596t basses, 795. See also black basses; sea basses calico bass, 322 largemouth bass, 242, 699 spotted sand bass, 371, 795 striped bass, 373 basswood, 134, 139 Bastias, B. A., 489 bat ray, 322, 796 bat star, 320 batholith, 49 Batillaria attramentaria, 371, 380 Batrachochytrium dendrobatidis, 199, 628, 697 Batrachoseps, 197 Batrachoseps spp., 514 Batrachoseps major aridus, 198 Batrachoseps pacificus, 764 Batrachoseps simatus, 198 Batrachoseps stebbinsi, 198 bats, 159, 162, 203, 204, 594, 648, 657, 720, 723, 724, 731 western yellow bat, 654 Baumhoff, M. A., 178–79 Baxter, W. T., 543 Bay Area Conservation and Development Commission (BCDC), 910t, 913 Bay Area Greenbelt Alliance, 909t, 913 bay checkerspot butterfly, 234, 465. See also Edith’s Bay checkerspot butterfly Bay Delta Conservation Plan (BDCP), 379, 380 bay laurel, 138, 237, 512, 548f. See also California bay laurel bay mussel, 371 Bay shrimp, 371, 373 beach animals, 391–93 beach bur, 414 beach ecosystem attributes and food webs, 394–400 zonation and, 395 beach ecosystems functions, 400–401 human impacts and influences, 401–2 beach evening primrose, 415, 416 beach food webs, 395–96. See also beach ecosystem attributes and food webs beach grass American beach grass, 413, 422 European beach grass, 412, 413, 417, 419–22 beach hoppers, 397f, 398, 399f beach layia, 414f beach morning glory, 414f beach mouse, Perdido Key, 207 beach-nesting fishes, 399–400 beach pea, silky, 414f beach profiles and features, 391–93

946  INDEX

beach spectacle pod, 414 beaches, sandy, 389, 403 conservation and restoration strategies, 402–3 geomorphic properties and characteristics, 389–91 key drivers, processes, and patterns, 393–94 Beale Air Force Base (AFB), 838 Bean, L. J., 177 bean clams, 396f bear clover, 555 bear-dogs, 162 beargrass, bigelow, 146 bears, 160, 162. See also grizzly bears black bear, 627t short-faced bears, 157f, 163 Beatley, J. C., 647 beaver-tail prickly pear cactus, 655 beavers, 157f, 160–62, 203, 681, 727, 730 North American beaver, 603, 727 beech, 139 bees, 191, 271–72, 467, 541, 651, 652, 657 European honey bee, 195, 767 leafcutter bees, 417 beetle larva, 515 beetles, 193, 395, 398, 399, 415, 515, 541, 564, 569, 599, 644, 651, 721, 723f, 847, 934. See also bark beetles; diving beetles; pine beetles; tiger beetles Australian vedalia ladybeetle, 242 blister beetle, 652 fire beetles, 37 predacious beetles, 196 tenebrionid beetles, 650, 651 water beetles, 681, 698f wood borer (woodboring beetle), 114, 515 Bekker, M. F., 598 Belding’s ground squirrel, 594, 596, 627 Belding’s savannah sparrow, 379t, 400 Belkin’s dune tabanid fly, 415 Bell’s vireo, least, 201, 202 Bendire’s thrasher, 202 Bendix, J., 180 Benedict, M. R., 871 benthic algae, 316, 369, 694, 698f, 726 benthic invertebrates, 370, 371 benthic macroalgae, 349, 368–69 benthic macroinvertebrates, 628 benthic microalgae, 368 benthic-pelagic coupling, 344–45 coastal upwelling, 345 complex life histories, 345–46 conceptual diagram of, 345f Berberis, 139 Bergemann, S. E., 540 berm crest, 392 berms, 392 Betula, 139 Betula occidentalis, 593 Betulaceae, 134 Bewick’s wren, 486, 488t Big Basin State Redwoods Park, 548 big galleta, 647 big leaf maple, 237, 512, 541, 559, 721 big sage, 647, 648 big sagebrush, 146 bigberry manzanita, 482f, 490 bigcone Douglas-fir, 35, 218, 222, 483, 484t bigelow beargrass, 146 bigeye tuna, 795 bighorn sheep, 169f, 204, 840t desert bighorn sheep, 594, 626, 627t, 630, 652, 657 Sierra Nevada bighorn sheep, 594, 596, 627t, 628, 630

bigleaf maple. See big leaf maple bigpod ceanothus, 479–80f, 490, 494–96 billfishes, 292 bioaccumulation, 685, 732, 772 bioclimate, 218 bioclimate envelope models (BEMs), 257 biocontrol, 241–42 biocrusts, 643–44 biodiversity, 187–90, 207, 932. See also species biodiversity, ecosystem functioning, and ecosystem services in decision making, 275–78 climate change and, 261 ecosystem functions and, 267–68 facets of, 265–66 levels and losses of, 266–67 Biodiversity, International Convention on, 928–29 biodiversity goals, 932 biodiversity hotspots, 203, 204, 205f, 206t, 512 biodiversity measures, 187–88 biogenic habitat, 318 biogeochemical cycles, 599–600, 853. See also under chaparral biological communities. See estuarine biota and their roles in the ecosystem biological organic carbon pump, 297 biological soil crusts (BSC), 643–44 biomagnification, 732 biophilia hypothesis, 659 biophysical research, 937 bioretention, 888 bioswales, 888 biotic homogenization, 886, 926 biotic interactions, 260 birches, 134, 139 water birch, 593 birds, 199. See also specific topics bill size diversity, 375 biogeography, 200–202 chaparral, 486, 488t conservation context, 202–3 in deserts, 656–57 endangered/threatened species that require or use estuarine and tidal marsh habitats, 379t evolutionary diversification, 199–200 feeding on beach invertebrates, 399 in wetlands, 680 bishop pine, 67, 200, 215 bison, 157f, 163, 178, 204, 455, 522, 767, 772, 885 American bison, 760t, 763 Bison, 161, 369 Bison bison, 204 bitter cherry, 222, 594 bitterbrush, 145, 147 antelope bitterbrush, 647 bittern, least, 680 bivalve reefs, 367 bivalves, 343, 345, 368 Bivalvia, 723f bivoltinism, 599 black abalone, 320, 344, 769, 797 black-backed woodpecker, 562, 563f, 596 black basses, 726, 881 giant black sea bass, 795 black bear, 627t black-bellied plover, 398, 399f black-capped chickadee, 200 black carbon, 119 black-chinned sparrow, 488t black cottonwood, 593 black flies, 721, 722, 723f, 731

black-footed albatross, 293f, 294, 296 black-headed grosbeak, 256t black locust, 139 black mustard, 40, 433t, 457 black-necked stilt, 373f, 685 black oak, 237, 510, 512t, 514f, 515–17, 554t, 557, 559t, 560t, 563f, 565f. See also California black oak black oystercatchers, 344 American black oystercatcher, 400 black phoebe, 400 black rails, 201, 680 California black rail, 379t, 854 black rat, 240, 759t, 772 black rockfish, 321, 323, 327f black sage, 433t black stain fungus, 599 black storm petrel, 202 black-tailed deer, 417 black-tailed jackrabbit, 417, 418, 420, 654, 657 black tern, 202 black-throated gray warbler, 256t black-throated sparrow, 656–57 black-vented shearwater, 202 blackbirds Brewer’s blackbird, 400 tricolored blackbird, 680 yellow-headed blackbird, 202 blackbrush, 146, 148, 642, 647, 648, 657 blackbuck, 760t Blackburn, T. C., 179 blackbush, 225 Blackfordia virginica, 371 blackgill rockfish, 794f blacksmith, 320, 321 bladdered kelps, 343 bladderweed, 344 blades (kelp), 312 blennies, 322 Blenniidae, 322 Blennosperma, 192 Blepharipoda occidentalis, 398f, 399 blind snake, western, 198 blister beetle, 652 Blodgett Experimental Forest, 568f, 826f, 827 Blodgett Forest Research Station, 825, 827 blowouts, 412 blue-blossom ceanothus, 830 blue butterflies, 189t, 195, 415 blue dicks, 433t blue-gray gnatcatcher, 256t blue-green algae, 716. See also blue-green bacteria blue-green bacteria, 343. See also blue-green algae blue grouse, Mount Pinos, 201 blue oak, 11f, 49, 220, 509–11f, 512–15, 517– 19, 521–25, 528, 554t, 557, 849, 851, 932 blue oak savanna, 849, 851. See also oak savanna blue oak–foothill pine, 849. See also oak savanna blue rockfish, 321, 327f, 790, 795 blue shark, 292f, 293, 769, 795 blue whale, 203, 295–98, 300, 792 blue wild rye, 450 bluebirds mountain bluebird, 596 western bluebird, 514 bluebunch wheatgrass, 225, 845 bluefin tuna, 293 bluegill, 243 bluegrass, Douglas’, 414f blueprints, regional, 919

blunt-nosed leopard lizard, 198, 199, 847 boa kelp, feather, 312, 395f boas, 198 boatman, water, 698f bobcat, 487t, 516, 527, 658, 852, 935, 937 bocaccio, 786, 790, 795 Boechera, 626 Bograd, Steven J., 98 bogs, 675 boid snakes, 198 boime boundaries/species ranges, 258–59 Bolander beach pine/Bolander pine, 67, 214, 536, 592 Bolboshoenus fluviatilis, 684 bole, 543 boletes, 483 Bolger, D. T., 439 bollworm, pink, 242 bombyliids, 650f Bonasa umbellus, 200 bond cycles, 132 bone-crushing dogs, 163 bonito, Pacific, 795 booby, brown, 202 Bootettix argentatus, 651 Boraginaceae, 189t, 190t, 485t borers, 515 wood borer (woodboring beetle), 114, 515 Borophagus, 162 Borrego milkvetch, 652 Bos bison, 763 Bosmina longirostris, 696 Botryllus schlosseri, 328 bottom-up effects, 396 box elder, 512 box turtles, 727f boxthorn, 769 Bracero Program, 872 Brachyistius frenatus, 317f Brachyramphus marmoratus, 541 brackish habitats, 367, 669, 671, 673, 674, 684 Brahea edulis, 193 Bramidae, 795 Branchinecta conservatio, 681 Branchinecta longiantenna, 681 Branchinecta lynchi, 681 Branchinecta sandiegonensis, 681 branching bushes, 343 branchiopods, 681 Branchiosyllis exilis, 328 Brandt’s cormorant, 323, 371 Branta canadensis moffitti, 680 Branta hutchinsii leucopareia, 200 Brassica, 461 Brassica nigra, 40, 433t, 457 Brassica tournefortii, 659 Brassicaceae, 414, 625 brassicas, 272 Brazilian waterweed, 682 breakwaters, 393 Brennania belkini, 415 Brewer’s blackbird, 400 Brewer spruce, 139 brin, 702 brine shrimp, 194, 374, 701, 702 bristle snail, Trinity, 828 bristlecone pine, 146–48, 224, 582, 590, 596–99, 602, 605 bristlecone pine forests, 591 brittle stars, 320, 321, 343 brittlebush, 648 broad leaf filaree, 452 broadcast spawning, 397 broadleaf forest, 220 Brode, J. M., 726

Broitman, B. R., 316 bromes, 433t, 461 California brome, 450 foxtail brome, 452, 461 ripgut brome, 420, 455, 458, 468, 770–71 Bromus carinatus, 450, 513 Bromus diandrus, 238, 420, 513, 771 Bromus hordeaceus, 238, 433t, 452, 513 Bromus madritensis, 452, 513, 659 Bromus madritensis ssp. rubens, 433t Bromus rubens, 440, 442 Bromus spp., 461 Bromus tectorum, 37, 225, 233, 603, 644, 659 Brontotheres, 160 brook trout, 696 broom snakeweed, 646 Broughton, J. M., 178 Brower, David, 550 Brown, Jerry, 920 Brown, P. M., 543 Brown, Pat, 912–13, 917 brown algae, 317, 327, 349 brown booby, 202 brown Irish lord, 323 brown pelicans, 200, 202, 372, 772, 788 California brown pelican, 704 brown rockfish, 327f, 373 brown shrimp, 298 brown trout, 699 brush rabbit, 487t, 514, 543 Bryophyta, 716 Bryozoa, 343 bryozoans, 318, 320, 321, 324, 328 Bt horizons, 49, 50, 51f, 52 Bubo virginianus, 459, 488t buckbrush, 193, 484t, 642 buckeyes, 133f, 218 California buckeye, 512 buckthorn, California, 433t buckwheat, 458 California buckwheat, 433t Santa Cruz Island buckwheat, 764 wild buckwheat, 772 buffalo, 177 buffalo sculpin, 323 Bufo boreas, 438f, 721 Bufonidae, 727f Bugula, 321 Bugula neritina, 328 built environment, 885 bulk density, 852–53 of soil, 456–57, 852–53 bull kelp, 312, 313, 315–17, 323, 324, 328 bullfrogs, 199, 237–38 American bullfrog, 199, 680, 727, 732 bullheads, 726 bulrushes, 684 California bulrush, 679 hardstem bulrush, 678f bumblebees, 272 bunchgrasses, 220, 225, 458 bur, beach, 414 bur clover, 433t Bureau of Land Management (BLM), 84, 891 Bureau of Reclamation. See United States Bureau of Reclamation Burgess, S. O. O., 545 Burian, S. J., 889–90 burn severity, 29 burned area emergency response (BAER), 39 burning. See also fire controlled, 179–80 landscape patterns of, 33–34 plant nutrients released upon, 66 prescribed, 568, 684 season of, 31

INDEX  947

burro, 652, 657, 759t burro bush, 225 burrowing legless lizard, 418 burrowing owl, 443, 459, 465, 655, 847, 851 burrowing shrimp, 371 bursages, 647 triangleleaf bursage, 642, 647 white bursage, 142, 147, 148, 642, 647, 649f Bury, R. B., 726 bush-mallows, 189t bushtit, 438f, 488t bushy-tailed packrat, 133f, 596. See also bushy-tailed woodrat bushy-tailed woodrat, 627. See also bushytailed packrat Busing, R. T., 544 Buteo jamaicensis, 459, 488t, 628 Buteo regalis, 459 Buteo swainsoni, 459, 465 butterflies, 195, 254, 415, 541, 626, 651 Bw horizons, 49 bycatch, 786–87 Byrne, R., 180 Byrnes, J., 322 cabezon, 322, 326, 327f Cabrillo, Juan Rodriguez, 107 cackling goose, Aleutian, 200 cacti, 142, 147, 220, 647, 661 arborescent cacti, 225 barrel cactus, 648 beaver-tail prickly pear cactus, 655 cholla cacti, 636 cactus mouse, 657 caddisflies, 194, 698f, 719, 721, 723f Castle Lake caddisfly, 189t Cafius spp., 399 Cairney, J. W. G., 489 Cakile edentula, 416, 419 Cal Fire, 33 Calamagrostis, 626 Calanus, 290 calcic horizon, 63 calcium carbonate, 63, 64 calico bass, 322 calico flower, 679 calicoflowers, 192 calicotheres, 160 Calidris alba, 394f, 400 Calidris himantopus, 202 Calidris minutilla, 372 California. See also specific topics ecological features, 229–30 physiographic regions, 47, 48f California barracuda, 323 California bay laurel, 139, 218, 480, 515, 559, 829 California Beechey ground squirrel, 851 California black oak, 221, 559. See also black oak California black rail, 379t, 854 California Borderlands, 760, 761f California brittlebush, 433t California brome, 450 California brown pelican, 704 California buckeye, 512 California buckthorn, 433t California buckwheat, 433t California bulrush, 679 California Ridgway’s rail, 201, 372, 379t California closed-cone pines, 142, 145 California Coast Ranges. See Coast Ranges California Coastal Act of 1976, 83, 420–21 California Coastal Commission, 278, 914 California coastal dune beetle, 415

948  INDEX

California Coastal Prairie, 850 California coffeeberry, 513 California condor, 188, 200–203, 266, 400 California cone snail, 321 California Cooperative Fisheries Oceanography Investigation (CalCOFI), 793 California Cooperative Oceanic and Fisheries Investigations (CalCOFI), 287 California cordgrass, 237, 678f, 679 California Current (CC), 96, 97, 140, 141, 292 California Current Ecosystem (CCE), 295, 302 currents and oceanographic features, 288f fidelity and attraction to, 293f map of, 288f California Current System (CCS), 95, 100– 101, 287, 292 circulation, 96–97 future challenges in, 99–100 mesoscale structure, 97 offshore region, 287 temporal variability decadal/PDO-NPGO, 98–99. See also North Pacific Gyre Oscillation; Pacific Decadal Oscillation interannual/ENSO, 98. See also El Niño Southern Oscillation seasonal phenology, 97–98 California Department of Fish and Wildlife (CDFW) Invasive Species Program, 241 California Department of Water Resources (DWR), 909t, 912 California Desert Conservation Area Plan, 84 California Desert Protection Act of 1994, 84 California Endangered Species Act (CESA), 526, 936 California Environmental Quality Act of 1970 (CEQA), 910t, 911t, 914, 917 California Essential Habitat Connectivity Project, 930f, 931 California fan palm, 654 California fetid adder’s tongue, 542 California Fish and Game Commission (CFGC), 789 California fivespinned ips, 564 California freshwater shrimp, 681, 724 California Fur Rush, 83, 727 California giant salamander, 726 California Global Warming Solutions Act of 2006, 936. See also Global Warming Solutions Act of 2006 California gnatcatcher, 37, 201, 202–3, 434, 439, 488t California Gold Rush, 171, 732 agriculture following, 521–22, 869 and capitalist production, 89 livestock and, 522, 767, 837, 842, 849 lumber, logging, and, 548, 818, 820 mining and, 51, 79, 88, 837 non-native plant species and, 231 population growth during and after, 76, 521–22, 671, 730, 842 urbanization following, 891 and wildlife, 204 California golden beaver, 727 California golden trout, 237 California grizzly bear, 509, 926 California ground squirrel, 400, 514, 516 California grunion, 401 California gull, 202, 701 California halibut, 371, 372f, 373, 381, 769, 790, 795 California hazel, 541–42 California huckleberry, 542

California Islands, 755–56, 774. See also Channel Islands of California; island ecosystems map of, 757f patterns of variability among geographic distinctions and commonalities based on geologic history, 762–63 vegetation bearing the imprint of physical environment and past land use, 763 vegetation changes, 763–64 physical characteristics and land ownership, 755–56, 758t California juniper, 146, 147 California killifish, 371 California Land Conservation Act of 1965, 913. See also Williamson Act of 1965 California laurel. See California bay laurel California leaf-nosed bat, 657 California least tern, 201, 379t, 401 California legless lizard, 198 California lilac, 139, 192, 480, 484t California lizardfish, 323 California Marine Invasive Species Act, 380 California Marine Life Protection Act of 1999, 350 California moray eel, 323 California mountain ash, 594 California mouse, 438f California mussel, 342, 348f, 377 California Natural Resources Agency, 936 California Naturalist Program, 5 California newt, 726 California oak moth, 515 California oak woodlands, 509, 515, 517, 519, 527, 855, 935. See also oak woodlands California oaks, 513f, 515, 518, 522 California Oaks Foundation, 526 California oakworm, 515 California plantain, 465 California poppy, 416, 433t California prehistory. See also vegetation prehistory framework for, 171–77 California Protected Area Database (CPAD), 902 California quail, 465, 488t California Rangeland Conservation Coalition, 855 California Rare Bird Committee (CRBC), 199, 200 California red-legged frog, 199, 238, 726, 848 California red scale, 239 California red tree vole, 206t California sage/California sagebrush, 433t, 437f, 439, 764 California scorpionfish, 327f California sea hare, 318 California sea lion, 295, 401, 767, 769, 791 California sea mussel, 179 California sheephead, 322, 327f, 795 California small wrasses, 322 California spiny lobster, 321–22 California spotted owl, 201, 561, 563f California State Office of Planning and Research (OPR), 900 California thrasher, 201, 488t California tiger salamander, 198, 465, 675, 680, 848, 851 California Tomorrow, 909t, 913 California towhee, 201, 488t California Undercurrent (CUC), 97 California vole, 376t, 438f, 759t. See also voles

California walnut, 512 California Water Plan, 704, 909t, 912 California Wild Scenic Rivers Act of 1972, 910t, 912 Californios, 85 Callianax biplicata, 398f Calliostoma, 318, 319f Callipepla californica, 465, 488t Callippe silverspot butterfly, 195 Callitropsis nootkatensis, 580. See also Chamaecyparis nootkatensis Callophrys mossii bayensis, 195 Callorhinus ursinus, 204, 767 Callospermophilus lateralis, 595t, 627 Calocedrus, 144f Calocedrus decurrens, 32, 221, 536, 537t, 557, 817 Calochortus monanthus, 189t Calvin, Jack, 338 Calycadenia, 192 Calypte anna, 438f, 488t Calypte costae, 488t Calystegia soldanella, 414f camels, 157f, 160–63, 173, 455 Camissoniopsis cheiranthifolia, 415 Canada del Puerto Creek, 773 canary rockfish, 786, 795 Cancer, 372. See also Metacarcinus Cancer antennarius, 371 Cancer gracilis, 399 Cancer magister. See Metacarcinus magister Cancer productus, 371 Canidae, 206t, 595t Canis latrans, 442, 459, 487t, 514, 516, 595t, 764, 852, 935 Canis lupus, 204 canker rot fungi, 515 cannabis. See marijuana canopy. See overstory canopy closure, 561 canopy cover, 557, 566, 568 canopy greenness, 435 canyon live oak, 218, 510, 512t, 559 canyon towhee, 488t cape iny, 235, 238 capillary fringe, 61 Capra hircus, 763, 767 caprellid shrimp, 371 Caprimulgus arizonae, 202 carabid beetles, 721 Carabidae, 399 carbon (C), 854 black, 119 dissolved organic, 729f rivers and, 720–21 carbon carrying capacity, 567 carbon credits, 908 carbon cycle, 644–45 carbon losses due to fire, 39 carbon monoxide (CO), 108, 120–21, 122t carbon production in estuaries, 373–74 carbon pump, biological organic, 297 carbon sequestration (and storage), 274, 297, 374, 919–20. See also carbon storage carbon storage. See also carbon sequestration (and storage) fire effects on, 38–39 montane forests and, 566–67 oak woodlands and, 524 carbonate horizons, 640, 641 Carcharodon carcharias, 293, 323, 796 Carcinus maenas, 371 cardinal, northern, 202 Cardinalis cardinalis, 202 Carduelis pinus, 628 Caretta caretta, 294

Carex, 625 Carini, S., 700 Carnegiea gigantea, 636 Carnivora, 203, 206t, 595t carnivoran creodonts, 156, 157f carnivores, 37, 157f, 158–60, 162, 163, 203, 205f, 316, 329, 514, 516, 541, 594, 595t, 628, 653, 658, 728, 759t, 847, 935 beach ecosystem attributes, food webs, and, 399–400 in kelp forest ecosystems, 321–22 native, and livestock production, 852 in rocky intertidal ecosystems, 343–44 carnivorous gastropods, 322, 323 Carpobrotus chiloensis, 420 Carpobrotus edulis, 235, 418, 419f Carpodacus cassinii, 596 Carreiro, M. M., 889–90 Carroll, Lewis, 927 carrying capacity, 76–77 Carson, Kit, 875 Carson, Rachel, 702 Carson Range escarpments, 139 Carya, 138 Caryophyllaceae, 485t, 625 Cascade Range birds in, 200–201 geology and geomorphology, 58 soil landscape relationships, 58–59 volcanic soil properties, 59–60 Cascades frog, 199 Cassin’s auklet, 202, 293, 294, 770, 771 Cassin’s finch, 596 Castanea, 139 Castilleja, 454 Castilleja leschkeana, 189t Castilleja uliginosa, 189t Castle Lake, interannual fluctuations in primary productivity in, 702–3 Castle Lake caddisfly, 189t Castor canadensis, 203, 603, 681 Castor canadensis subauratus, 727 catadromous species, 371 Catalina eddy, 12 Catalina grass, 772 Catalina ironwood, 765 catch hyperstability, 792–93 caterpillar tractors, 548 catfish, 726 Cathartes aura, 400, 514, 658 Cathedral Peak, 223f cation exchange capacity, 50 cations, 496 Catostomidae, 727f Catostomus occidentalis lacusanserinus, 725 Catostomus platyrhynchus, 725 Catostomus santaanae, 725 Catotomus occidentalis, 725 cats, 160, 162, 199, 205, 344, 759t. See also saber-toothed cats domestic, 442, 459, 659, 759t feral, 459, 659, 769, 772 cat’s ear, 847 cattails, 684 cattle, 85, 89, 205, 231, 272, 455, 522, 523, 567, 603, 628, 672, 675, 718t, 730, 731, 756, 760t, 767, 836–37, 840, 842, 845, 848–50, 852, 853, 855, 857, 868–70, 871f, 875 range, 841 cattle grazing and oaks, 523 Caulacanthus ustulatus, 344 Caulerpa taxifolia, 235, 241, 380 cavitation, 492 cavitation resistance, 492, 493f cavity nesters, 563

Cayan, Daniel R., 564 ceanothus, 480, 483, 489, 491, 493 bigpod ceanothus, 479–80f, 490, 494–96 blue-blossom ceanothus, 830 Ceanothus, 192, 220, 513 Ceanothus sp., 484t, 494 Ceanothus spp., 480, 496t, 555, 598, 642 Ceanothus cordulatus, 482f Ceanothus crassifolius, 193 Ceanothus integerrimus, 222, 482f Ceanothus megacarpus, 435, 479f, 490f Ceanothus papillosus, 482f Ceanothus spp. subgenus Cerastes, 492t Ceanothus thyrsiflorus, 482f, 830 cedar, 145, 146. See also yellow-cedar incense cedar, 32, 144–48, 221, 536, 537t, 554t, 557, 559, 560t, 817, 821, 825 Port Orford cedar, 536, 537t Spanish-cedar, 138, 139 western red cedar, 146, 536, 537t Cedrela, 138, 139 Centaurea melitensis, 433t, 435f, 460, 513 Centaurea solstitialis, 205, 232–33, 454, 461, 513, 850 centipedes, 651 Central Coast, 875 Central Valley, 197, 198 agriculture, 880 geology and geomorphology, 51–52 lakes, 704–5 landforms, 53f last glacial period in, 145 marshes, 672f, 673 pesticides and nutrient and metal contamination, 733 wetlands, 672–75, 684, 687 Central Valley grasshopper, 189t Central Valley Project (CVP), 90, 730, 871– 72, 909t, 912 Central Valley soil landscapes, spatial extent of, 53f Central Valley soils soils of the basin and east side, 52, 53f highly developed soils on dissected high terraces on eastern edge, 52 highly developed soils on low terraces, 52 weakly developed soils in basin alluvium and recent floodplain deposits, 52 weakly developed soils on basin margins and broad alluvial fans, 52 soils of the delta, 54 soils of the west side, 54 vernal pool soils, 54 Central Valley Water Project, 871 Centrarchidae, 727f Centrocercus urophasianus, 37, 628 Centrostephanus, 319f Centrostephanus coronatus, 318 cephalopods, 293, 295 Cephaloscyllium ventriosum, 322 Cepphus columba, 294 Cerastium glomeratum, 513 Ceratomyxa shasta, 729 Ceratostoma, 319f, 344 Ceratostoma foliatum, 321 Ceratostoma nuttalli, 321 Cercis, 139 Cercocarpus, 139, 483 Cercocarpus sp., 484t Cercocarpus spp., 642 Cercocarpus betuloides, 484t, 496t Cercocarpus ledifolius, 580 Cercyon fimbriatus, 415

INDEX  949

Cercyonis sthenele sthenele, 195 cereal yellow dwarf virus, 461 Cerithidea californica, 371 Cerorhinca monocerata, 294 Cervidae, 596t Cervus canadensis nannodes, 188, 204, 522, 671, 841. See also tule elk Cervus canadensis roosevelti, 188, 673, 674f Cetacea, 203 cetaceans, 292, 294–95, 297, 299 Cetorhinus maximus, 293, 795 Chaetodipus sp., 487t Chaetodipus californicus, 487t Chaetodipus fallax, 487t Chaetodipus penicillatus, 657 Chaetorellia succinea, 233 chain migration, 81 Chalk Bluffs, 138 Chalybion, 650f Chamaea fasciata, 201, 438f, 488t Chamaebatia foliolosa, 555 Chamaecyparis, 139 Chamaecyparis lawsoniana, 536, 537t Chamaecyparis nootkatensis, 222. See also Callitropsis nootkatensis chamise, 56, 220, 479–80f, 480, 482f, 483, 484t, 489, 490f, 491, 498 chamisso bush lupine, 416, 417 Channel Islands National Marine Sanctuaries (NMS), 300, 301t Channel Islands of California, 296, 756, 761f, 933, 934. See also California Islands; islands in transition diversity, 757, 758 endemism, vulnerability, and resilience endemism among plants and animals, 764–65 paleoendemics and neoendemics, 765–66 factors contributing to resilience and recovery of, 774 human history. See also island ecosystems earliest North Americans, 766–67 early maritime hunting and fishing, 767 oil exploration, military uses, tourism and recreation, 767–68 ranching and agriculture, 767 terrestrial mammal introductions and eradications on, 756, 759–60t unique flora and fauna, 756 Channel Islands slender salamander, 764, 765 Channel Islands spotted skunk, 765 Chaoborus astictopus, 702 chaparral, 187f, 218, 219f, 220, 479–80, 499. See also specific topics biogeochemical and hydrological dynamics carbon exchange, 496 ecosystem services, 498–99 hydrology, 498 mineral nutrition, 496–98 compared with coastal sage scrub, 434–35 fire and, 32, 486–89, 494–96, 499. See also chaparral areas: postfire future of, 499 geography, 480 life histories and wildfire, 486–89 in the long absence of fire, 494–96 organisms found in animals, 486, 487t, 488t mycota and other microbiota, 483–86 plants, 483, 484t–486t origins, 482–83 chaparral animals, 486, 487t, 488t

950  INDEX

chaparral areas birds found in, 486, 488t mammals found in, 486, 487t postfire plants found in, 483, 485–86t predators found in, 486, 488t chaparral pea, 483, 484t chaparral shrubs, physiology of, 489 freezing and distribution, 493–94 growth and photosynthesis, 490 leaf function of deciduous and evergreen, 490, 491t life history type and, 494 water stress, 491–93 chaparral vegetation distribution, 480, 481f variation in, 480, 482f Chapin, F. S., 927 Charadriiformes, 200 Charadrius alexandrinus nivosus, 379t Charadrius montanus, 201 Charadrius nivosus, 398, 399f Charadrius nivosus nivosus, 201, 414 Charadrius vociferous, 400 charcoal, 179, 180 Chasmistes brevirostris, 725, 728 cheat grass, 37, 225, 603, 644 cheatgrass, exotic, 644 checkerspot butterflies, 195, 234, 254, 439, 465 Checkley, D. M., 96, 101 cheesebush, 648 cheeseweed, 771 Chelonia mydas, 199, 294 chemoautotrophic fixation of CO2, 700 chemocline, 702 Chendytes lawi, 179 Chenopodiaceae, 144, 189t Chenopodium littoreum, 414 chenopods, 648, 649f cherries, 139, 483 bitter cherry, 222, 594 holly-leaved cherry, 484t sweet cherry, 272 chess, soft, 452 Chester, M. V., 889 chestnut, 139 chewing insects, 486 chickadee, black-capped, 200 chickens, 868t, 875 Chilean sea fig, 420 chilipepper, 786, 795 China rockfish, 323, 327f, 794f Chinook salmon, 100, 175, 292, 379t, 680, 724, 728, 739, 786, 790f, 792, 798, 829 chinquapins giant chinquapin, 218 golden chinquapin, 559 Chionactis occipitalis, 653 chipmunks, 206t, 486, 487t, 628 alpine chipmunk, 164, 206t, 594, 596, 627 lodgepole chipmunk, 596 yellow-cheeked chipmunk, 206t, 540f Chironomidae, 722, 722f, 723f chironomids, 697, 723f Chiroptera, 203, 205f, 595t chisel-tooth kangaroo rat, 657 chitons, 318, 342, 343, 345, 352 Chlidonias niger, 202 chlorophyll, 695, 701, 702 chlorophyll-a, 396 mixed-layer, 701f Chlorophyta, 716 Chlorostoma, 343, 350 chlorotic mottle, 108 cholla cacti, 636

Chondestes grammacus, 459 Chondracanthus, 343 Chordata, 343 Chromis, 319f Chromis punctipinnis, 320, 321 Chromista, 343 chronosequences, 59, 67 Chrysolepis chrysophylla, 218, 559 Chrysolepis sempervirens, 482f, 598 Chrysolina, 242 Chrysothamnus, 225 Chrysothamnus spp., 647 Chrysothamnus viscidiflorus, 625 Chthamalus, 343 chubs, 725 thicktail chub, 189t chuckwalla, common, 653 chum salmon, 728 Chumash occupation of Channel Islands, 768 chytrid fungi, 237, 604, 628. See also fungal chytrid chytridiomycosis, 697 Cicindela, 417 Cicindela hirticollis gravida, 415 Cicindela ohlone, 195 Cicindelidae, 415 cienegas, 673 ciliates, 368, 369 Cincindela spp., 399 Cinnamomum, 138 cinnamon, 138 cinnamon teal, 680 cinquefoils, 190t Cirsium rhothophilum, 414 Circus cyaneus, 459 cirques, 140, 620 Cirripedia, 342 Cirsium occidentale, 417 Cirsium praeteriens, 189t Cistaceae, 485t Cistothorus palustris, 680 cities, California’s, 885, 892. See also urban areas of California biotic environment of, 886–87 integrating biotic and abiotic components of, 889–90 citrus, 872, 876 citrus crops, 238 citrus industry, 873 citrus virus, 886 cladocerans, 717 Cladophora, 719–21, 726 Cladophora glomerata, 719 clam shrimp, 681 clams, 173, 322, 371, 372, 374, 377, 379, 395, 402, 717 Asian clam, 699 Asiatic clam, 722 fingernail clams, 723f Pacific razor clam, 396 pea clams, 698f Pismo clam, 396–97, 399 Clarke, L. W., 886 Clarkia, 191, 192 clarkias, 191 Clark’s grebe, 372 Clark’s nutcracker, 589–91, 596, 597f, 605, 628 clasts, 64 claypin, 50 Clean Air Act (CAA), 108, 120 Clear Lake, 703f, 706 polymixis and mercury pollution, 702 Clear Lake gnat, 702 Clear Lake splittail, 189t

Clements, Frederick E., 837, 846 Clementsian equilibrium ecology, 837 Clemmys marmorata. See Actinemys marmorata clerid beetle, 650f Clethrionomys californicus, 595t Clethrionomys gapperi, 561 cliff swallow, 256t, 400, 724, 732 cliffs, 620, 620f Clifornia tiger salamander, 232 climate adaptation, 276–77 of alpine ecosystems, 629–30 climate change, 251, 262, 601, 685–86, 936. See also sea surface temperature; subalpine forest ecosystem services and biodiversity, 261 and biotic interactions, 260 in California, 251–52, 253f, 936 and coastal sage scrub, 439 and dunes, 422 and ecosystem management, 261–62 and fire, 40–41 and fisheries, 798 future challenges and, 920 impacts of future, 256 impacts to date, 253–56 and invasive species, 260–61 and land use planning, 920 in a multistressor context, 252–53 and range management, 856–57 research needs that emphasize climate change as an emerging driver, 937 species’ (in)ability to track and capitalize on, 257 and species ranges, 257–59 as threat multiplier, 259–60 velocity, 257, 258f and wetlands, 685–86 climate change projections, 100 climate-dependent surplus production (fisheries), 782–84 climate regimes, abiotic stress, and alpine ecosystems, 617, 618t, 619 climate data for various locations, 618t climate variability, 295–96. See also El Niño Southern Oscillation; Pacific Decadal Oscillation climate water deficit (CWD), 510, 514f climate zones, 214, 215f climate(s), California, 22–23, 132f. See also precipitation; specific topics and agriculture, 866 atmospheric rivers (ARs), 14–15, 365, 555, 586, 617 coastal influence, 17 cool season, 14 general climate features, 10–12 nature of, 9–10 ocean influence on, 14–15 range of, 16–17 regional features, 20–22 seasonal and diurnal temperature variation, 17–18 spatial variability of temperature, 17 warm season, 12, 13f climatic water deficit (CWD), 583 climax community, 845 climax state, 837, 846 climax theory, 846 clingfishes, 322 Clinidae, 322 clinids, 322 clintonia, Andrew’s, 542 Clintonia andrewsiana, 542 Clitellata, 723f clonal krummholz mats, 589 closed-cone pines, California, 142, 145

clotbur, spiny, 771 cloud albedo, 20 cloud condensation nuclei, 118 clouds, stratus, 12 clovers, 454, 458, 466 bear clover, 555 bur clover, 433t Clupea pallasi, 291 Cnidaria, 343 cnidarians, 235f Coachella Valley fringe-toed lizard, 199 coarse-grained environments, 30 coast, California. See coastal California coast horned lizard, 239, 459 coast live oak, 214f, 236–37, 483, 484t, 495, 510, 511f, 512, 514f, 515, 517, 518, 521– 23, 541, 548f, 849 Coast Ranges geology and geomorphology, 55 lakes in, 702–3 soils, 55–56 coast redwood, 139, 144, 146, 213–15, 217, 536, 537t, 675, 678f. See also coastal redwood coast redwood forest water balance, fog and, 544–45, 546f coast redwood forests, 535, 550–51. See also redwood forests future of, 550 geological history and features, 535–36 soils and unique plant associations, 68, 536 physiographic setting, 535–36 coast strawberry, 415, 416, 417f Coastal Act. See California Coastal Act of 1976 coastal California, 338 biogeographic patterns, 339–40 coastal geology and topology, 338–39 the contrast between terrestrial and aquatic environments, 341–42 dune field soils, 67 geology and geomorphology, 66–67 marine terrace soils, 67 physical environment, 340–41 physical features, 360, 362f, 363 topography, 362f wave exposure, 342 Coastal Commission. See California Coastal Commission coastal cutthroat trout, 728 coastal cypress, 142 coastal dune beetle, California, 415 coastal dune plant species, 414–15 introduced and invasive, 419–20 patterns of vegetation, 415 coastal dunes, 409, 423 community ecology effects of plants, 416–17 interactions between plants, animals, and microbes, 417–18 components of dune systems, 422–23 ecosystem services, 418 effects of humans, 418–20 formation and shape, 410, 412–13 future scenarios, 421–22 geography, 409–10, 411f macroorganisms, 414–15 management strategies, 420–21 native vegetation on, 410, 412f physical conditions for organisms, 413–14 population and evolutionary ecology, 415–18 restoration, 421 coastal fir, 35 coastal giant salamander, 726

coastal goosefoot, 414 coastal pelagic fishes, 793 Coastal Prairie, 850 coastal prickly pear, 433t coastal protection, flood mitigation and, 274–75 coastal ranges. See Coast Ranges coastal redwood, 237. See also coast redwood coastal sage scrub, 429, 444. See also specific topics chaparral compared with, 434–35 community interactions native and exotic species interactions, 435–36 plant-animal interactions, 436 soil biota, 436–37 ecosystem services provided by. See also coastal sage scrub ecosystem food and forage, 429–30 physical features and control over distribution, 430, 433 urban recreation and biodiversity services, 430 fragmentation by urban development, 442 future scenarios, 443–44 human impacts from postcolonial era to present change in fire frequency, 441–42 climate change, 439 impacts on plants and animals, 442 nitrogen deposition, 440–41 urbanization and agriculture, 439–40 management and restoration, 442–43 organisms, 433–34 protected area status, 429, 431f transitions between grassland and, 438–39 coastal sage scrub distribution, 429, 431f coastal sage scrub ecosystem. See also under coastal sage scrub characteristics, 434–35 variation in time and space, 438–39 temporal variation in plant community composition, 437 coastal sage scrub native and exotic plant species, 433t responses to fire, 433t coastal sagewort, 416 coastal strand vegetation, 392 coastal upwelling. See upwelling coastal wetland soils, 67–68 coastal wetlands, 673–74 coastlines land use, 82–84 and sculpting during Pleistocene Ice Ages, 762 cobweb thistle, 417 Coccinellidae, 415 cochineal insects, 768 cockeye salmon, landlocked, 699 cockroach, desert, 651 cod, 794 Coelopa spp., 398 Coelus ciliatus, 415 Coelus globosus, 415 Coenonympha tullia, 415 coffeeberries, 483, 484t California coffeeberry, 513 coho salmon, 175, 292, 379t, 724, 728, 792, 798, 829 cohort resonance, 785 Colaptes chrysoides, 202 cold desert, 224 Cole, D. N., 849 Coleogyne ramosissima, 146, 225, 642

INDEX  951

Coleoptera, 723f collembolans, 644, 651, 658 collinsia, hillside, 190 Collinsia sparsiflora var. collina, 190 Collinsia corymbosa, 414 colluvium, 55, 600t Cololabis saira, 291 colonial botryllid tunicates, 371 colonial moss animals, 343 colonial tunicate, 328 colonizations, 142, 144 Colorado Desert. See also deserts climate, 638 Coluber spp., 514 Colubridae, 198, 727f Columba livia, 488t commodity production, land use for, 84 common-pool resource, 878 compensatory plant growth, 845 competition, 234t apparent, 419 competitive exclusion, 347 condor, California, 188, 200–203, 266, 400 cone snail, California, 321 confined animal feeding operations (CAFOs), waste disposal from, 879, 880 congeners, 237, 399 Connell, Joe, 346–47 Conochilus unicornis, 696 Conozoa hyalina, 189t conservancy fairy shrimp, 681 conservation, 193. See also specific topics in an era of change, 927 managing fisheries for economic efficiency to support, 802–4 conservation challenges in 21st century, 916 mismatches between policy goals and legislative tools, 917–18 mixed messages to localities, and no additional resources, 916–17 tensions between local authority and regional planning, 916 conservation context, 203–7 conservation development, 891 conservation easements, 526, 856, 900, 901, 907, 915 conservation history, 902. See also land conservation history of California timeline of legislation and regulation affecting conservation, 909–11t conservation land acquisition history, 901, 906–7 timeline, 906f conservation planning, 275–76, 828 conservation-reliant species, 774 conservationists, 83 conspecifics, 348 continental background, 113 continental rim, creation and disruption of, 760, 761f continentality, 214–16 controlled burning, 179–80 conurbations, 82 Conus, 344 Conus californicus, 321 convection, atmospheric, 16 convergence hypothesis, 483 “conveyor belt,” warm-moisture, 14 Cooper, W. S., 33 Cooperative Weed Management Areas (CWMAs), 241 Cooper’s hawk, 488t, 514 Coordinating Council, 913 cootie, 138 coots, 728 American coot, 680

952  INDEX

copepods, 289, 290, 295, 302, 368, 370, 371, 377, 540f, 698f, 717 copper rockfish, 327f, 794f coppice management, 522 coprovores, 651 Corallina, 343 coralline algae, 317, 318, 325f, 343, 352 coralline sculpin, 317f Corbicula fluminea, 699, 722 corbina, 399 Corbula, 377 Corbula (Potamocorbula) amurensis, 368, 371, 377 Cordell Bank National Marine Sanctuary (NMS), 301 cordgrass, 268–69, 372 California cordgrass, 237, 678f, 679 smooth cordgrass, 237 coreopsis, giant, 764, 769 Corixidae, 698f cormorants, 401 Brandt’s cormorant, 323, 371 double-crested cormorant, 201 corms, 35 corn, 177, 855, 874 cornucopianism, 77 Cornus nuttallii, 557 correlative rights, 876 Cortaderia jubata, 229f Corvidae, 774 Corvus corax, 488t, 650 Corvus spp., 400 Corydalidae, 723f Corylus, 542f Corylus cornuta californica, 541–42 Corynactis californica, 321 Coryphaena hippurus, 795 Costaria costata, 314, 315f Costa’s hummingbird, 488t Cottidae, 322, 727f cotton, 871 cotton rat, 162 cottonwoods, 139, 647, 678f black cottonwood, 593 Fremont cottonwood, 512, 679, 935 cottony cushion scale, 242 Cottus beldingi, 725 cotyledons, 543 cougar, 658 Coulter pine, 56, 218, 222, 483, 559 councils of government (COGs), 914 cover, 218 cow-calf producers, 840t, 841, 855 Cowart, A., 180 cowcod, 786, 788, 795, 803 coyote, 164, 459, 487t, 516, 527, 655, 658, 659, 764, 851, 852, 935, 937 coyote brush/coyote bush, 221, 417, 439, 513, 523, 763, 772, 844, 850 coyote tobacco, 193 crabs, 318, 320–23, 343, 350, 371, 372, 374f, 376, 377, 399, 728, 788, 797 Dungeness crab, 371–73, 381, 784, 790, 791f, 797 hippid crabs, 395 northern kelp crab, 318 sand crabs, 396, 399, 400 craneflies, 723f cranes, 376. See also sandhill cranes Crangon, 371 Crangon nigricauda, 399 Crangon spp., 373 crappie, 726 Crassostrea gigas, 374, 798 Crassostrea sikamea, 798 Crassostrea virginicas, 798

Crataegus, 139 crawling water beetles, 681 crayfishes, 195, 717, 722, 724, 728 Louisiana red swamp crayfish, 722, 726 Shasta crayfish, 681, 724 sooty crayfish, 189t crenulate spits, 390 creodonts, 156, 157f, 159 creosote bush, 142, 147, 148, 213, 636, 642, 644–48, 649f, 651, 654–56, 662 creosote bush desert scrub, 218 creosote bush scrub, 215f, 225 creosote grasshopper, 651–52 creosote roots, 646 creosote shrublands, uptake values from, 644 Crepis, 626 Crespí, Juan, 359, 376 crevice-dwelling sea cucumbers, 343 Cricetidae, 206t, 595t Cricetinae, 595t Crisia, 321 criteria pollutants, 119 critical loads (CL) of nitrogen (N) deposition, 117, 118f, 121t, 122f, 440, 441f. See also nitrogen (N) air pollution and atmospheric deposition critical zones, 644 croakers spotfin croaker, 399 yellowfin croaker, 373, 399 Crocanthemum greenei, 772 crocodiles, 159 crocodilians, 160 Cronartium ribicola, 564, 599 Crooks, K. R., 442 Crosby, Alfred, 78 cross-shore transport, 393 Crotalus cerastes, 653 Crotalus oreganus, 514 Crotalus ruber, 653 Crotalus scutulatus, 653 Crotalus viridis, 459, 486 Crotophytidae, 198 Croucher, P. J. P., 236 crown fire regimes, 32 crown fires, 28, 29 active vs. passive, 28 independent, 28 crown rust, 460 crows, 400, 589 cruise liners, 297 Crustacea, 723f crustacean fisheries, 797 crustaceans, 193f, 195, 235f, 288–90, 311, 318, 320–24, 371, 394, 397, 399, 675, 681, 696, 717, 720, 722, 791f demersal, 788 crypsis, 401 Crypsis schoenoides, 684 Cryptantha hooveri, 189t Cryptantha intermedia, 433t cryptanthas common cryptantha, 433t Hoover’s cryptantha, 189t Cryptochiton, 319f Cryptochiton stelleri, 318 cryptograms, 415 Cryptosporidium, 853 crystalline iceplant, 770 Cthamalus, 346 Cucumaria, 343 Cucumaria minitata, 321 cucumbers. See also sea cucumbers wild cucumber, 433t Culex spp., 698f

cultivation, 179 cultivation effect, 788, 796 cultural diversity, historical, 187–88 cultural resources management (CRM), 172 Cummins, P. F., 99 Cunningham, Laura, 5, 680 Cunningham Marsh cinquefoil, 190t Cupressaceae, 139 Cupressus. See Hesperocyparis Cupressus arizonica, 139 Cupressus goveniana ssp. pygmaea, 67 Cupressus guadalupensis guadalupensis, 193 Cupressus macnabiana, 484t Cupressus sargentii, 484t Cupressus spp., 142 Curculio sp., 515 Curculionidae, 189t curl-leaf mountain mahogany, 580, 593–94 curlew, long-billed, 201, 372, 373f Cuscuta salina, 679 cushion scale, cottony, 242 Cushman, J. H., 413 cut-off lows, 19 cutthroat trout, 725, 829 coastal cutthroat trout, 728 Lahontan cutthroat trout, 699 Paiute cutthroat trout, 237 cuttle-fish, 311 cyanobacteria, 288–90, 343, 349, 643–45, 694, 700, 702, 705, 716, 717, 719, 720, 739 cyanolichens, 539 Cyanophyta, 716 Cycadaceae, 134 cycads, 134, 138 cyclonic storms, 20 Cydia latiferreana, 515 Cylindropuntia spp., 636 Cylindrospermum, 720 Cynara cardunculus, 433t cynipid wasps, 516 Cynipidae, 515 Cynoscion nobilis, 323 Cyperaceae, 145, 625, 849 cypresses, 36, 139, 144–46, 193, 483, 484t, 489, 536 coastal cypress, 142 Mendocino cypress, 67 Monterey cypress, 215 Santa Cruz cypress, 215 Tecate cypress, 40 cyprinid splittail, 739 Cyprinidae, 189t, 727f Cyprinodon spp., 725 Cyprinodon macularius, 656, 704 Cyprinodon radiosus, 188 Cyprinodontidae, 727f Cystoseira. See Stephanocystis Cytisus scopartus, 452 dace, 725 Dactylis glomerata, 461 Dactylopius sp., 768 Dahlgren, R. A., 545, 844 dairy industry, 872–73 transformation of, 874 daisies Mariposa daisy, 189t seaside daisy, 409–10f Dama dama, 231 dams, 738–39 damselflies, 194, 681, 722, 723f Dana Plateau, 214f Danaus plexippus, 195 Dance House, 175f Daphnia, 699

Daphnia middendorffiana, 698f Daphnia rosea, 696, 698f dark-eyed junco, 628 Darwin, Charles, 311, 312 Dasmann, Ray, 5 dates, 876 Davidson Current, 97 Davis, W. M., 694 dawn redwood, 139 Dawson, Todd E., 545 Dawson, William, 199 DDT (dichlorodiphenyltrichloroethane), 396, 772 DeBano, L. F., 38 decapods, 295, 722 deciduous chaparral shrubs, 490, 491t deciduous subalpine forests, 593–94 decomposers, 372, 467, 644, 648 decomposition, 497 in deserts, 644 ways fauna accelerate, 644 deep scattering layer, 290, 291 Deepwater Horizon oil spill, 299 deer, 37, 160, 162, 164, 169, 177, 178, 180, 203, 452, 455, 522, 759t, 767, 772, 852, 885 black-tailed deer, 417 fallow deer, 231, 759t mule deer, 225, 420, 487t, 626, 627t, 630, 657, 724, 760t, 763, 764, 773 deer-brush, 222 deer mice, 162, 164, 417, 419, 487t, 514, 543, 627–28, 765 Anacapa deer mouse, 240, 770 deer weed, 433t, 496 deflation plains, 412 Delairea odorata, 235, 238 Delhi Sands flower-loving fly, 66 Delphinus delphis, 292f Delta Breeze, 22 delta smelt, 371, 372f, 377, 379t, 738 Delta Wetlands Project, 704 Deltistes luxatus, 725, 728 demersal crustaceans, 788 demersal fishes, 800 demersal fishing practices, 299 demersal gillnets, 785, 796 demersal habitats, 779, 787, 797 demersal marine fish, 371 demersal species, 371 Dendragapus fuliginosus howardi, 201 dendritic tidal creeks, 673 Dendroctonus brevicomis, 564 Dendroctonus jeffreyi, 564 Dendroctonus ponderosae, 564, 599 Dendroctonus spp., 37 Dendroctonus valens, 599 Dendromecon, 139 dengue, 880 denitrifying bacteria, 682 dense false gilia, 193 dentition, 158 Department of Water Resources (DWR), 909t, 912 deposit feeders, 398 depositional chutes, 620 derived characteristics, 415 Dermochelys coriacea, 290, 294 Descurainia pinnata, 193 desert agave, 254, 647 desert bees, 651 desert bighorn sheep, 594, 626, 627t, 630, 652, 657 desert cockroach, 651 Desert Conservation Area Plan. See California Desert Conservation Area Plan

desert ecosystem components and processes, 642–43 biological soil crusts, 643–44 carbon cycle, 644–45 decomposition, 644 nitrogen cycle, 645 nutrient transfers and distribution, 645–46 desert ecosystem processes, drivers of, 636 climate, 636, 638–40 geology, 640 geomorphic settings and associated vegetation, 641–42 soils, 640–41 desert ecosystem services, 658 provisioning services of water, food, and fiber, 658 cultural services, 659 energy, 658–59 desert icons, California, 654–55 desert iguana, 653 desert ironwood. See ironwood desert life, animal adaptations to, 648, 650–51 invertebrates, 651–52 vertebrates, 652–53, 656–58 desert mistletoe, 656 desert packrat, 133f desert pavement, 62 desert plants. See also vascular plants adaptation to extreme conditions, 646–47 invasive plants and fire regimes, 660 plant-plant interactions and vegetation dynamics, 647–48, 649f desert pocket mouse, 657 Desert Protection Act. See California Desert Protection Act of 1994 desert pupfish, 656, 704 desert range, forage in hot and cold, 845 desert regions, 635, 637f desert shrew, 657 desert slender salamander, 198 desert-thorn, 142, 147 San Nicolas Island desert-thorn, 189t desert tortoises, 198, 199, 653, 655, 659, 920 desert units, California areas of, 635, 636t desert woodrat, 654 deserts, 84, 635–36, 662. See also specific deserts biogeography, 636, 637f birds in, 202 climate, 17 food webs, trophic pyramids, and key consumers, 648, 650–51 future scenarios, 660–61 incorporating into land management, 661–62 impacts of humans during postcolonial period, 659–60 invertebrates and biodiversity in, 194–95 lakes in, 703–4 land use, 84–85 restoration strategies, 661, 662 DeSimone, S. A., 439 Desmarestia, 315, 317 Desolation Wilderness, 223f detrital food web, 436 detrital macrophytes, 345 detrital particles, 343, 345, 398 detrital pools, 290, 373 detrital processing, 389 detritivores, 193, 317, 318, 319f, 320–23, 329, 486, 651, 658, 681, 843f, 857 in deserts, 658

INDEX  953

detritus, 289f, 290, 318, 320, 370, 375f, 398, 400, 650f, 682, 698f, 716, 720–24, 731, 787 Diablo winds. See Santa Ana winds Diacodexis, 157f ,156 diameter distribution, 561 diameter limits, 567–68 Diamond, N. K., 524 diamond rattlesnake, red, 653 diamond turbot, 373 Diaptomus signicauda, 696 diatom epiphytes, 720. See also epiphytic diatoms diatoms, 117, 288, 344, 395, 696–97, 716 Dicamptodon tenebrosus, 726 Dichelostemma capitatum, 433t dichlorodiphenyltrichloroethane (DDT), 396, 772 dicks, blue, 433t Dicosmoecus, 719 Dicosmoecus gilvipes, 719 Dictyoneurum californicum, 314, 315f Dictyota, 317, 319f didelphid marsupials, 157f, 160 Didelphimorpha, 203 Didelphis virginiana, 231, 442 Didemnum, 321 Didymosphenia, 717 Didymosphenia geminata, 717 Diegan sage scrub, 432f, 434, 438, 441f, 443 Diffendorfer, J. E., 437, 442 differenced normalized burn ratio (dNBR), 29 Diguetia, 650f Diguetia mohavea, 650f dimictic periods, 694 dinoflagellates, 288–99, 344 dinosaurs, 135, 156, 203 Dinothrombium pandorae, 195 Diopatra ornata, 320 diplurans, 651 Dipodomys sp., 203, 438f, 487t Dipodomys spp., 851 Dipodomys agilis, 206t, 487t Dipodomys heermanni, 206t, 457, 487t Dipodomys ingens, 206t, 851 Dipodomys microps, 657 Dipodomys nitratoides, 206t Dipodomys nitratoides exilis, 205 Dipodomys nitratoides nitratoides, 205 Dipodomys stephensi, 206t, 207, 443 Dipodomys venustus, 206t, 487t Dipsosaurus dorsalis, 653 Diptera, 717, 723f dire wolf, 157f, 163 Disporum smithii, 542 Dissanthelium californicum, 772 dissipative beaches, 391 dissolved organic carbon (DOC), 729f distinct population segments (DPSs), 724, 735 disturbance, 234t, 848, 928 defined, 714 disturbance regimes, 234 Dithyrea maritima, 414 diurnal temperature variation, 17–18 divergent selection, 375 diversity. See also biodiversity functional, 932 diversity hotspots. See biodiversity hotspots diving beetles Mono Lake diving beetle, 189t, 194 predaceous diving beetle, 681, 698f dock, 458 Dodonaea viscosa, 139

954  INDEX

“dog-bears,” 161, 369 dogfish, 289f dogs, 157f, 160–63, 421, 659, 720, 759t domestic, 659, 759t, 772, 841 feral, 659 dogwood, 557 Doliosaurus, 198 dolphins, 157f, 163, 292f, 297, 376 Domagalski, J. L., 700 Donax gouldii, 396, 396f donkey, 759t dorid nudibranchs, 321, 349 dormancy, 489 Doryteuthis opalescens, 291, 784. See also Loligo opalescens; market squid Dosidicus gigas, 291, 293, 798 double-crested cormorant, 201 Douglas-fir, 145–47, 218, 221, 237, 536–37, 539–41, 548, 554f, 555, 557, 559, 560f, 817, 818, 819f, 820f, 821, 822, 830 bigcone Douglas-fir, 35, 218, 222, 483, 484t Douglas’s bluegrass, 414f Douglas’s squirrel, 596 Dover sole, 790f doves, 488t dowitchers, 373f long-billed dowitcher, 399f downcoast, 390 Downingia, 192 downingia, Hoover’s, 677f Downingia bella, 677f downstream, 714 Draba, 625, 626 dragon kelp, 312 dragonfly, 194, 681, 722, 723f drainage areas, 714 Drake, Francis, 177 Dreissena polymorpha, 722 Dreissena rostriformis bugensis, 722 drought-deciduous leaves, 217 dry meadow, 600t dry playas, 641 Dryas integrifolia, 626 Duane, T. P., 523 ducks, 371, 372, 376, 675, 685, 728 duckweed, 679 Dugan, Jenifer E., 401 Dull, R. A., 853 dune beetle, 415 dune field soils, 67 dune grass American dune grass, 412, 413 Eureka dune grass, 193 dune localities, 410 dune rush, 417 dune sheet, 412 dune tabanid fly, Belkin’s, 415 dunes, 393, 642. See also coastal dunes Dungeness crab, 371–73, 381, 784, 790, 791f, 797 Dunne, J. P., 99–100 Durrenberger, R., 872 dusky-footed woodrat, 514, 828–29 dust, 118–19 dwarf island fox, 400 dwarf mistletoe, 599 dwarf viruses, yellow, 461 dwarf wild flax, 192 Dyksterhuis, E. J., 846 dynamic general vegetation models (DGVM), 525 dynamic global vegetation models (DGVMs), 257–58 dynamic management areas (DMAs), 302 dynamic ocean management, 300

Dyschirius marinus, 399 Dytiscidae, 189t E horizons, 67 eagles, 656, 658 bald eagle, 188, 772, 773 golden eagle, 514, 628, 655, 658, 773 eared grebes, 701 earthworms, 459, 467, 515 aquatic, 717, 723f East Mojave Scenic Area, 84 eastern gray squirrel, 231 eastside pine, 558f, 560–61 Ebeling, A. W., 325 Echinochloa crus-galli, 684 echinoderm fisheries, 796–97 Echinodermata, 343 echinoderms, 318, 323, 345, 350, 791f Echinoidea, 345 ecofunctionalism, 177 ecological footprint, 886 ecological interactions, 268–69 ecological monitoring for adaptive management, 936 public participation and, 938 ecological restoration, 41, 821 ecological sites, 846 classifications, 845–46 ecological speciation, 375 ecological staircase, 536 ecology in California early development of, 2 writings on, 2 economic growth, 872, 873 ecoregion map, 203f ecosystem and agricultural regions of California, defining the, 875–76 ecosystem-based fisheries management (EBFM), 798–99 enriching traditional managing with, 799–800 ecosystem disservices, 889 ecosystem ecology, 267 ecosystem engineers, 268, 516 ecosystem functions, 267–69 ecosystem services and, 269 examples, 271 ecosystem management in the changing future, 261–62 ecosystem service initiatives, range management and, 855–56 ecosystem services defined, 269, 566 examples, 271 natural capital and, 269 niche markets for, 856 supply, service, and value, 269–71 types of, 269 ecosystem stewardship. See stewardship ecosystems, 888 in California. See also specific topics emergent patterns, 2–4 recommended reading, 5 healthy, 780, 799 9ecosystems in California patterns emerging across, 3–4 ecotone, 579 ectomycorrhizae (ECM), 483, 484, 540 ectomycorrhizal fungi, 237 ectomycorrhizal species, 495 ectoprocts, 235f ectotherms, 653 edaphic conditions, 557 edaphic endemics, 255 edaphic gradients, 216

eddy covariance, 519 edge effects, 442, 444, 522 Edith’s Bay checkerspot butterfly, 195. See also bay checkerspot butterfly Edith’s checkerspot butterfly, 254 Edmondson, W. T., 694 Eel River, 175, 363, 364 characteristics, 718t natural food webs, 719–22 Eel River biota, seasonal phenology of, 719–20 eelgrass, 371, 378, 397, 679 Japanese eelgrass, 380 eelpouts, 322 eels California moray eel, 323 wolf eel, 322 efflorescent crust, 119 Egeria densa, 682 Egerton-Warburton, L. M., 437 Egregia, 313f, 317, 343 Egregia menziesii, 312f, 314t, 398 egrets, 372 great egret, 371 snowy egret, 323, 399f Egretta thula, 323, 399f Ehrharta calycina, 417, 419f Ehrlich, Paul, 76 Eisenia, 313f, 319f, 343 Eisenia arborea, 314, 314t, 315f Ekman layer, 96 El Niño Southern Oscillation (ENSO), 22, 98–101, 131–32, 735 alpine ecosystems and, 619 beaches and, 394, 402 coastal sage scrub and, 439 estuaries and, 360, 365, 371 kelp forests and, 316, 325, 328, 329 montane forests and, 555, 569 offshore ecosystems and, 294–96 overview, 15–16, 98 subalpine forest ecosystems and, 586 El Segundo blue butterfly, 415 El Segundo flower-loving fly, 415 Elanus leucurus, 459 elasmobranchs, 791f, 795 elder, box, 512 electrical conductivity (EC), 54 elephant, 162 elephant seal, 293f, 296, 772, 792 northern elephant seal, 292f, 295, 401, 414, 767, 792 elevation and climate, 17 elevation shifts, climate change and, 254, 255f elf owls, 202 elfin butterfly, San Bruno, 195 Elgaria, 198 Elgaria spp., 459, 514 Elgaria panamintina, 198 elk, 169, 178, 180, 455, 760t, 885 elk kelp, 312, 313, 317 elm, 134, 139 eluvium. See E horizons Elymus caput-medusae, 850 Elymus elymoides, 225 Elymus glaucus, 450, 513 Elymus mollis, 412, 413f Elymus spicatus, 845 Elymus triticoides, 450 Embiotocidae, 322, 727f, 790 emboli, 492 Emerald Lake, 695 Emerald Lake watershed, 599–600 Emergency Wetlands Resources Act of 1986, 686

Emphyastes fucicola, 398 Empidonax traillii, 200 Empidonax traillii brewsteri, 200 empty niche, 234t Emydidae, 727f Emys marmorata. See Actinemys marmorata Enantiornithines, 199 Encelia californica, 433t, 437 Encelia farinosa, 225, 648 Encelia virginenses, 142 Endangered Species Act, California. See California Endangered Species Act Endangered Species Act of 1973 (ESA), 195, 196, 827, 855, 900, 910t, 918 endangered species that require or use estuarine and tidal marsh habitats, 379 endangered stream invertebrates, 724 endemics, edaphic, 255 endemism, 191, 192, 625, 626, 636, 722 Endocladia, 343 endophytes in redwood forest, 539–40 endorheic lakes, 694 endosymbionts, 729 Engelmann oak, 510, 511f, 512t, 932 Engelmann spruce, 222, 537t, 557, 593 English sole, 371–73, 381 Engraulis mordax, 291, 788 Enhydra, 319f Enhydra lutris, 204, 268, 372, 788 Enhydra lutris nereis, 322, 379t Enophrys bison, 323 Ensatina, 197 Ensatina spp., 514 entelodonts, 160, 161 Enteromorphora, 722 Entisols, 49, 52, 55, 56, 58–61, 62f, 65–67 environmental change, 925–26. See also stewardship, conservation, and restoration ecological responses to, 925–27 future research, education, and outreach for addressing, 937–38 environmental impact assessment and management, 277–78 environmental law, 277, 379, 733, 789, 908, 909–11t. See also policy environmental protection strategies, development of, 188 Environmental Quality Act. See California Environmental Quality Act of 1970 Eocene, 137, 138f mammals in, 159–60 Eocene-Oligocene extinction event, 137 Eopsetta jordani, 802 Epantius obscurus, 398 Ephedra, 225 Ephedra nevadensis, 649f Ephemeroptera, 194, 723 Ephydra hians, 194, 701 epibionts, 318 epicormic buds, 543 Epigaulus, 159 epipelic algae, 717 epiphyte communities, 68, 121, 123 epiphytes, 318, 343, 719 in redwood forest, 539–40 epiphytic algae, 318, 717 epiphytic algal production, 721 epiphytic bacteria, 679 epiphytic diatoms, 719, 726. See also diatom epiphytes epiphytic lichens, 116, 117, 121, 123, 509– 10f, 513 epiphytic oak parasites, 514 epiphytic plants, 539

epiphytic species, 539 epiphyton, 716, 717 Epithemia, 719–20 Epithemia sorex, 729 epithilic algae, 717 equids, 658 Equisetum, 134 Equus, 162 Equus asinus, 652 Equus caballus, 657 erect coralline algae, 319f Eremophila alpestris, 438f, 459, 656 Eremophila alpestris insularis, 201, 765 Erethizon, 162 Erethizon dorsatum, 595t Erethizontidae, 595t Ericaceae, 540 Ericameria ericoides, 416 Ericameria linearifolia, 147 Ericameria spp., 141 ericoid mycorrhizae, 540 Erigeron glaucus, 409f Erigeron mariposanus, 189t Eriogonum, 513 Eriogonum spp., 458, 772 Eriogonum arborescens, 764 Eriogonum fasciculatum, 432f, 433, 433t, 435, 439, 485t Eriophyllum lanatum var. grandiflorum, 193 Erlandson, J. M., 179, 181 Erodium spp., 443 Erodium botrys, 452 Erodium cicutarium, 433t, 452, 513, 845 erosion, 641 coastal, 393–94. See also beaches erosion and accretion cycles, 392, 393 Erysimum ammophilum, 414 Erysimum menziesii, 414 Erythrobalanus, 515 escapement, 782 Escherichia coli, 272, 629, 853 Eschrichtius robustus, 792 Eschscholzia californica, 416, 432f, 433t Esox lucius, 188, 241 Essential Habitat Connectivity Project. See California Essential Habitat Connectivity Project estuaries, 359–63, 371, 381, 881 classic, 365 climate system: precipitation and runoff, 363. See also runoff: from estuaries extreme flows, 365 interannual patterns, 364–65 mean annual patterns, 363–64 multidecadal patterns, 365 complex patterns of ecosystem variability, 367 patterns of variability over time, 367–68 spatial patterns, 367 El Niño and, 360, 365, 371 habitat types and diversity within, 359– 60, 365–67 the human dimension and forms of human disturbance, 376 habitat losses, 378 human health advisories, 378 landscape transformation, 376–77 pollution, 377–78 species introductions, 377, 378f species losses, 378–79 what has been lost and what is at risk, 378–79 large spatial gradients, 367 a long view of, 380–81 lowland channels to, 721–22

INDEX  955

estuaries (continued) plans for recovery and protection controlling species introductions, 380 habitat restoration, 379–80 public policies, 379 principles of how they function in ecosystems, 368–73 steps to lessen the threats to, 381 watersheds, 363t what is at risk, 380–81 what will drive future changes, 380 estuarine biota and their habitats, services provided by, 372–73 fish nursery and migration corridors, 373 flood buffering, 374 food production, 374 opportunities to understand processes of evolution, 374–75 organic carbon production, export, and storage, 373–74 waste assimilation, 374 estuarine biota and their roles in the ecosystem (biological communities), 368–72 birds, 371–72, 373f, 373t fish, 371, 372f interactions between, 372–75 invertebrates, 370–71 microbes, 369–70 primary producers, 368–69 estuarine endemics, 380 estuarine fisheries, loss of, 379 eucalyptus, 235, 773 Eucalyptus camaldulensis, 773 Eucalyptus globulus, 235, 773 Euceratherium, 162 euchalon, Pacific, 379t Eucyclogobius newberryi, 379t, 703 Eugster, H. P., 704 eulachon, 728 Eularia, 313f Eularia fistulosa, 314t Eumeces spp., 459 Eupentacta, 320 Euphagus cyanocephalus, 400 Euphausia pacifica, 290, 296 euphausiids, 289–90, 295, 296, 302 Euphilotes battoides allyni, 415 euphotic zone, 288. See also photic zone Euphydryas editha, 254 Euphydryas editha bayensis, 195, 234, 465 Eurasian watermilfoil, 682, 699 Eureka dune grass, 193 Eureka Valley evening primrose, 193 Eurhynchium oreganum, 542 European beach grass, 412, 413, 417, 419–22 European flat oyster, 798 European hare, 759t European honey bee, 195, 767 European mouflon sheep, 756, 760t, 767 European rabbit, 759t, 763, 769 European searocket, 419 Eurytemora affinis, 371 Euspira, 323 Eustenopus villosus, 233 Euthynnus spp., 795 Euthynnus pelamis, 292 eutrophic status, 694, 703–4 eutrophication, 116, 498, 698, 699 Euzonus, 398–99. See also Thoracophelia evaporative cooling, 342 evapotranspiration, 363, 465, 498. See also actual evapotranspiration (AET); potential evapotranspiration evening primroses beach evening primrose, 415, 416

956  INDEX

Eureka Valley evening primrose, 193 evergreen chaparral shrubs, 490, 491t evergreen huckleberry, 539, 540f, 829 evolutionary biology, possibilities for advancing, 375 ex-urban development, 886 exchangeable sodium percentage (ESP), 54 Excirolana spp., 399 Exclusive Economic Zone (EEZ), 789 exotic cheatgrass, 644 exotic species, 229, 230, 603. See also invasive species; species: non-native exotic vascular plants, 231 number of, 231f extinct species, 188, 189–90t extinctions, megafaunal, 177–78 extirpations, 142, 144 Extriplex californica, 225 exurban refugees, 81 Fabaceae, 414, 416, 485t facultative methylotrophic bacteria, pinkpigmented, 437 facultative seeders, 35, 488 Fagus, 139 fairy lantern, 542 fairy shrimp, 195, 641, 675, 681 Falco peregrinus, 400, 628, 772 Falco sparverius, 488t falcon, peregrine, 400, 628, 772 fall transition, 97 fallow deer, 231, 759t false-cypress, 139 false saber-toothed cats, 157f, 161 fan palm, California, 654 Farallones, Gulf of the. See Gulf of the Farallones farewell-to-spring genus, 191 farm gate value, 866 farm size and agriculture, 866–67 farmland in California, 866, 867f. See also agriculture Farmland Protection Act of 2001, 911t, 917 farmland values, index of, 871 Farris, G., 179 feather boa kelp, 312, 395f fecundity, 454 Federal Aid Highway Act of 1956, 83 Federal Highway Acts, 909t, 911–12 Federal Wilderness Areas, 602 fee title ownership, 902 Felidae, 595t Felis catus, 199, 769 Felis concolor, 658 Felis domesticus, 459 Felis rufus, 595t Felis sylvestris catus, 442 fellfield, 600t, 621, 624 fellfield community, 590 felsenmeer, 620 fence lizard, western, 438f, 459, 486 fens, 675 feral pigs, 231, 238 feral populations, 231 ferns, 359, 361f, 548f sword fern, 218, 542 Ferocactus eastwoodii, 648 Ferranto, S., 856 ferruginous hawk, 459 fertility, 80–81 replacement-level. See replacement level fescues Idaho fescue, 225 rat tail fescue, 433t tall fescue, 461 Festuca arundinacea, 461

Festuca idahoensis, 225 Festuca perennis, 458 fetid adder’s tongue, California, 542 Ficus, 138 fiddleneck, common, 433t fiddler crabs, 371 Field, J. C., 791 figs, 89, 138, 139, 159 figwort, 458 filamentous algae, 681 filarees, 454, 458, 461, 466, 661, 847 broad leaf filaree, 452 redstem filaree, 433t, 452, 845 filbert weevil, 515 filbertworm, 515 fin whale, 297, 298, 300, 792 finches Cassin’s finch, 596 gray-crowned rosy-finch, 201, 628 fine-grained patchiness, 30 fine particulate matter (PM2.5). See particulate matter finfishes, 179, 326, 379, 791 fingernail clams, 723f Finlay, J. C., 721 fir-spruce, 539 fire, 21, 27–28, 41. See also burning animal communities impacted by, 437 chaparral and, 32, 486–89, 494–96, 499 global changes, 40–41 in historical context, 32–33, 34f human impacts on, 32–33 intensity/severity, 29, 30f as management tool, 841–42, 928 patterns of spread, 30 postfire recovery of plant communities, 35 colonization from unburned metapopulations, 36 delayed seedling recruitment from in situ surviving parent plants, 36 delayed seedling recruitment from postfire resprouts, 36 endogenous regeneration, 35–36 restoration, 41 rivers and, 731 in southern California, 21 in subalpine forests, recreation and, 603–4 in Transverse and Peninsular Ranges, 65–66 types of fuels consumed, 28–29. See also fire regime types fire beetles, 37 fire departure index, 34–35 fire effects on air quality, 119–20 on animal communities, 36–38, 437, 438f on soils, hydrology, and carbon storage, 38–39 fire frequency, 30–31 in coastal sage scrub, change in, 441–42 fire management, 39–40, 604 fire “management” vs. suppression strategies, 34, 35 future directions, 934 postfire, 40 prefire, 39–40 fire policy, 34 fire regime, 28 mixed-conifer conditions produced by active, 565f fire regime types, 28–29, 32 fire return interval departure (FRID), 34, 35f fire return intervals (FRIs), 30, 218, 564 fire rotation period, 30

Firestone, R. B., 178 firewood, tree cutting for, 521–22 Firman, J., 523, 526 firs, 138, 139, 144–48, 221–22, 537, 557, 560f, 562f, 564–65, 580, 598, 821. See also Douglas-fir; red fir coastal fir, 35 grand fir, 218, 536 subalpine fir, 222, 557, 580, 593 white fir, 32, 221, 222, 224, 537, 554f, 557, 558f, 559, 560f, 561, 562f, 566, 580, 817, 821, 825 fiscalization of land use, 916–17 Fish and Game Commission. See California Fish and Game Commission fish nursery and migration corridors, 373 fish stock(s), 779, 781–805. See also fisheries consequences of depleting forage, 788 management of volatile, short-lived, 784 mixed-stock fisheries, 786, 799 stock assessments and management advice, 782 fisher (mustelid), 37, 204, 541, 561 fisher collection trap sites, 828f fisher reintroduction project, 827–28 fisheries, 287, 779–80, 790f, 804–5 aquaculture and, 273 as competitors, 788 continuing pressures on, 797–98 crustacean, 797 definition and nature of, 780–81 depensation, 796 historical landings and impacts, 791–97 increasing use of, 797 indirect ecosystem effects, 787 consequences of depleting forage stocks, 788 consequences of predator removal, 788 subsidizing scavengers and redirected trophic pathways, 788 kelp forests and, 326 loss of estuarine, 379 marine protected areas and, 800–801 mixed-stock, 786, 799 outlook for marine ecosystems and, 804 pathways to sustainable, resilient, 798–804 recent landings and value, 789–91 windfall harvests in developing, 782 Fisheries Ecosystem Plan, 799 fisheries ecosystem(s), 780–81. See also under fisheries conceptual view of a, 780–81 laws, agencies, and management, 789 fisheries management, 780–81, 783f, 799. See also ecosystem-based fisheries management collaboratively transforming, 804 laws and agencies, 789 management of volatile, short-lived stocks, 784 management structure and guiding legislation, 788–89 managing for economic efficiency to support conservation, 802–4 phases in the history of, 788–89 sources of uncertainty and bias data limitations, 784 recruitment variability and climatedependent surplus production, 782–84 stock assessments and management advice, 782 surplus production, sustainable yields, and stability, 781–82

Fisheries Oceanography Investigation. See California Cooperative Fisheries Oceanography Investigation fisheries science, 780. See also fisheries management Fishery Conservation and Management Act of 1976 (FCMA), 789 fishery management plans (FMPs), 789 fishes, 292–93, 317f, 322, 323, 327, 371– 73. See also crayfishes; groundfishes; pelagic fishes; pupfishes; rockfishes; starfish; sunfishes; trout; specific topics anadromous fish, 724 beach-nesting, 399–400 California killifish, 371 coastal pelagic, 792–93 in deserts, 656 endangered/threatened species that require or use estuarine and tidal marsh habitats, 379t extinct, 189t finfishes, 179, 326, 379, 791 forage fish, 289f, 290–96, 302, 788, 799 freshwater fishes, 187, 188f, 191 highly migratory species, 794 jellyfish, 99, 371 mountain whitefish, 725 nearshore, 399–400 planktivorous fishes, 320, 321, 377, 788 riverine food webs and, 724–26 starfish, 268, 311 tunny fish, 376 fishing early maritime, 767 how development changed the functioning of, 769–70 recreational, 705 fishing, ecological effects of, 784 population responses and conservationyield trade-offs, 784–85 age-structure truncation, maternal effects, and fisheries-induced evolution, 785–86 “fishing a stock down,” 781 fishing practices, demersal. See demersal fishing practices fishing quotas, individual, 803 fishing techniques, 785 five-needled pines, 599 fivespinned ips, California, 564 flagellates, 369 flannel bush, 139 flat abalone, 320 flat-topped krummholz, 253 flatfishes, 371, 791f, 794 flatworms, 343, 717, 723f flax, dwarf wild, 192 flea, water, 698f fleshy blades, 343 flies, 233, 415, 515, 541, 717, 723f. See also caddisflies alderflies, 698f, 723f alkali fly, 194, 701 black flies, 721, 722, 723f, 731 damselflies, 194, 681, 722, 723f Delhi Sands flower-loving fly, 66 fruit fly, 238 kelp flies, 399 mayflies, 194, 681, 698f, 723, 731 seaweed flies, 398 stiletto flies, 415, 417 stoneflies, 194, 722, 723 Voluntina Sonemyia fly, 189t flightless beetle, 415 flightless duck, 179, 181 flood buffering, 374

flood control districts, 876–77 flood mitigation and coastal protection, 274–75 flooding, 681, 714, 716 floodplain deposits, weakly developing soils in, 52, 53f floodplains, 369, 881 flora. See also specific topics biodiversity, 190–93 biogeography, 191–92 evolution, 625–26 evolutionary diversification, 191 Florida scrub-jay, 200 flounder, starry, 373 flow-regime adaptations, 726 flower-loving fly, Delhi Sands, 66 flycatchers, 200–202, 256t silky flycatcher, 656 flying squirrel, northern, 561, 563f Foehn wind, 21 fog water, 544–47 foliar feeder, 515 foliose algae, 318 foliose red algae, 315, 317, 319f folivores, 486, 648 food web(s). See also offshore ecosystems; riverine food webs beach ecosystem attributes and, 394–400 of California Current Ecosystem, 288, 289f desert, 648, 650–51 of Eel River, 719–22 of high-elevation Sierran lakes with introduced trout, 697, 698f foothill oak, 514f foothill oak woodlands, 511f foothill pine, 510, 512, 517f, 517t, 849 foothill woodlands, 512t. See also oak woodland types, widespread foothill; Sierra foothills foothill yellow-legged frog, 726, 727 foothills. See Sierra foothills forage, 836 forage fish, 289f, 290–96, 302, 788, 799 forage production, 272–73, 466 forage productivity, 837, 839f foraminifera, 133, 159, 254–55 forbs, 415, 450, 454, 466 foredunes, 392, 412–13 forest-alpine ecotone, 615f forest fires. See fire forest fishes, kelp, 317f forest habitat conservation plans (FHCPs), 828 forest management regimes global ecosystem services under different, 822 local ecosystem services under different, 822–30 expanding fishers into their historical range, 827–28 experimental managed forests, 825–27. See also Blodgett Experimental Forest Forest Practice Act of 1973, 550 forest resilience, 568 Forest Service. See United States Forest Service forest stand age, 823, 824f forest types, 817–18, 819f, 820t. See also montane forest types and ownership, 817, 818t, 820t and site productivity, 818, 820t forests, 817–19, 830. See also montane forests; timberlands defined, 579

INDEX  957

forests (continued) experimental managed, 825–27 family, 827 managing sun-demanding plants within shade-tolerant, 829–30 fork-tailed storm-petrel, 200 Formica obscuripes, 418 Fort Ross weevil, 189t foundation species, 516 defined, 516 fountain grass, 659 Fouquieria splendens, 148 four wing saltbush, 641 fowl, 376. See also waterfowl foxes, 344, 459. See also island foxes; kit foxes; red foxes foxtail barley, 458, 461 foxtail brome, 452, 461 foxtail pine, 146, 222, 224, 557, 559t, 580, 582, 592, 593, 596–98, 605 foxtail pine forests, 589–90 foxtail pine subspecies austrina, 589 foxtail pine subspecies balfouriana, 589 fracking, 920 Fragaria chiloensis, 415, 417f fragmentation. See habitat fragmentation Francis, R. C., 791 Franciscan complex, 338 Frangula, 483 Frangula sp., 484t Frangula californica, 484t, 830 Frangula rubra, 484t Frankia spp., 555 Fraxinus, 139, 536 Fraxinus latifolia, 512 Fraxinus latifolia Benth., 721 free-tailed bat, pocketed, 657 Freeland, H. J., 99 freeze-thaw cycles, 620–22, 623f freezing of chaparral shrubs, 493–94 Fremont cottonwood, 512, 679, 935 Fremontodendron, 139 Fremont’s goldfields, 677f French broom, 452 freshwater fishes, 187, 188f, 191 freshwater landscapes, future directions for management of, 934–35 freshwater systems, invasive species in, 240–41 freshwater wetlands, California’s, 669–71, 687. See also freshwater landscapes; wetlands area within ecoregions of California, 669, 670t biogeochemical processes, 682–83 drivers and patterns of variability, 677–78 variability in hydrology, 677–78 earliest and most extensively drained, 671 future scenarios, 685–87 geographic distribution, 673–77 habitat types, 678 habitats, biological communities, and their interactions, 678–81 historical ecology, 671–73 historical perspective on, 880 restored, 933f. See also wetlands restoration Fresno kangaroo rat, 206t Friends of the Los Angeles River (FoLAR ), 737 fringe-toed lizard, Coachella Valley, 199 frogs, 196, 198–99, 654, 697, 721, 723, 726, 727f, 733. See also bullfrogs; red-legged frogs; yellow-legged frogs Pacific tree frog, 438f, 459, 721 tailed frog, 197–98, 726

958  INDEX

fronds, 315 frost weathering. See freeze-thaw cycles frugivores, 648, 656, 657 fruit fly, 238 fruits, 417 Fucellia, 397f, 398 Fucellia spp., 398 Fucus, 343, 346 fuel management and woody plant encroachment, 850 Fujimori, T., 544 Fulica americana, 680 functional diversity, 266 functional groups, 932 functional redundancy, 268 fungal chytrid, 198, 199, 237. See also chytrid fungi fungal communities, soil, 436–37 fungal filaments, 643 fungal pathogens, 237, 543, 697 fungi, 39, 190, 236, 266, 318, 417, 457– 59, 483–85, 489, 495, 539–40, 599, 640, 643–45, 682, 694, 716, 722, 845, 853. See also arbuscular mycorrhizae; mycorrhizal canker rot fungi, 515 frog-trout-fungal interactions, 696 macrofungi, 414 sulfur fungus, 515 fur rush. See California Fur Rush fur seals, 293f, 767 Guadalupe fur seal, 792 northern fur seal, 204, 401 gadwall, 680 Gaertner, M., 419 Gaines, David, 705 Galeorhinus zyopterus, 794 gall wasp, 515 gallers, leaf, 486 galleta, big, 647 Gallinula chloropus, 680 Gaman, T., 523, 526 Gambelia sila, 198, 847 Gambusia affinis, 726 game fishes, 790, 791f, 794, 794f gametophyte, 315 gamma diversity, 266 Garboletto, M., 540 Gardali, T., 200, 202 Garibaldi, 322 Garrya sp., 220, 483, 484t Garrya elliptica, 484t Garrya fremontii, 484t Garrya veatchii, 484t garter snakes, 459, 727f aquatic garter snake, 680, 721 common garter snake, 721 giant garter snake, 198, 680 mountain garter snake, 697, 698f San Francisco garter snake, 198 gasoline production, 297 Gasterosteidae, 727f Gasterosteus aculeatus subspp., 725 Gastropoda, 723f gastropods, 195, 255, 289, 290, 318, 322–24, 350, 371, 457f, 458–59, 791f gastrotrichs, 681 Gaultheria, 540 Gaultheria shallon, 546f, 829 geese, 376, 675 Aleutian cackling goose, 200 gelastocorid bugs, 721 gelatinous zooplankton, 289 Gelidium, 343 Gelochelidon nilotica vanrossemi, 202

Gempylus serpens, 795 general circulation models (GCMs), 39 General Plans, 900, 907, 908, 909t, 910t, 914, 917, 919 General Sherman giant sequoia, 553 genetic diversity, 266 Genista monspessulana, 452 Geococcyx californianus, 201, 488t, 656 geoduck, 371 geologic processes. See also specific topics overview, 47–48 geomorphic surfaces, 52 geomorphology, 380 Geomyidae, 595t, 764 Geomys, 162 geophytes, 36, 37, 457, 624, 630 geostrophic balance, 96, 110 Geothlypis trichas sinuosa, 201, 376t Geukensia demissa, 371 ghost pine, 483 giant black sea bass, 795 giant chinquapin, 218 giant coreopsis, 764, 769 giant garter snake, 198, 680 giant green anemone, 352 giant kangaroo rat, 206t giant kelp, 311–18, 323–29, 395f, 397f, 768, 769 giant kelpfish, 317f giant keyhole limpet, 318 giant marmot, 163 giant red sea cucumber, 797 giant red velvet mites, 195 giant redwood, 192 giant reed, 205, 682, 770 giant sagebrush, 141, 145 giant salamanders, 726 giant sea bass, 323, 769, 795, 804 giant sequoia, 139, 142, 145, 192, 222, 223f, 554t, 558f, 559t, 560, 566, 597–98 Giant sequoias, 146 giardia bacteria, 603 Giardia intestinalis, 603, 853 Gigaspora, 484 Gila bicolor thallassina, 725 Gila crassicauda, 189t Gila monster, 198, 653 Gila orcutti, 725 gilded flicker, 202 gilia dense false gilia, 193 Hoffmann’s slender-flowered gilia, 764, 765f Gilia, 415 Gilia tenuiflora ssp. hoffmannii, 764, 765f Gill, J. L., 179 gillnets, 785–87 demersal gillnets, 785, 796 Gilmer, D. S., 680 ginger, wild, 542 Ginkgoaceae, 134 ginkgos, 134, 139 Girella nigricans, 318, 795 Giri, B. J., 700 glacial cirque lakes, 695 glacials, 256 glaciated valley slopes, 620 glaciers, 619–22 rock, 614, 621 Glass Mountain, 592f Glaucomys sabrinus, 561 Glaucopsyche xerces, 189t, 195 gliding mammals, 157f, 158 global change models (GCMs), 629, 630 Global Warming Solutions Act of 2006, 911t, 917. See also California Global Warming Solutions Act of 2006

globose dune beetle, 415 Glomus, 484 GLORIA (Global Observation Research Initiative in Alpine Environments), 629 Glossosoma, 721 Glossosoma penitum, 721 Glossotherium, 162 Glycera spp., 399 glyptodonts, 162 Glyptotherium, 162 gnat, Clear Lake, 702 gnatcatchers blue-gray gnatcatcher, 256t California gnatcatcher, 37, 201–3, 434, 439, 488t goatgrass, 453f, 454, 455, 469, 470 barbed goatgrass, 461, 466, 850 goats, 205, 456, 760t, 763, 767, 840, 841 gobies, 322, 371, 727f Gobiesocidae, 322 Gobiidae, 322, 727f godwit, marbled, 399f gold mining, 86. See also under California Gold Rush Gold Rush. See California Gold Rush golden chinquapin, 559 golden eagle, 514, 628, 655, 658, 773 golden-mantled squirrel, 627 golden trout, California, 237 goldenbush, 147, 148 goldfields, 192 Fremont’s goldfields, 677f Goldman, C. R., 694, 698, 702–3 goldspotted oak borer, 515 gomphotheres, 157f, 163 Goodman, D., 703 “goose pens,” 543 gooseberry, Parish’s, 190t goosefoot, 144, 146 coastal goosefoot, 414 gooseneck barnacles, 352 gopher rockfish, 323, 327f gopher snake, 459 gophers, 160, 162, 321, 323, 457, 459, 514, 764 Gopherus agassizii, 198, 653, 920 gorse, 452 goshawks, 37 northern goshawk, 561, 563f Goulden, M. L., 259 governance, adaptive, 889 government, councils of, 914 graben lake, 697 Grace, J., 535 graceful rock crab, 399 gram-positive and gram-negative bacteria, 455 graminoids, 624, 823t, 824 grand fir, 218, 536 Grandidierella japonica, 371 granitic terrain, soils of, 49–50 granivores, 37, 436, 457–59, 463, 486, 648, 651, 656, 851 Grant, Madison, 549 grape farming, 767 grapes, 89, 866 grass rockfish, 326, 327f grasshopper mouse, southern, 657 grasshopper sparrow, 459 grasshoppers, 418, 458, 459, 461, 651, 655 acridid grasshoppers, 191 Central Valley grasshopper, 189t creosote grasshopper, 651–52 tetrigid grasshoppers, 721 grassland distribution, factors controlling, 450, 452–55

biota, 455 birds, 459 insects, annelids, and gastropods, 458–59 large herbivores, 455–57 microbes, 459–60 nonavian predators, 459 small mammals, 457–58 interactions among biotic and abiotic factors, 460–61 grassland distribution, map of, 451f grassland ecology and management of ecosystem services, 465 biodiversity, 465–66 carbon sequestration, 467 fire control, 468 forage production, 466 pollination, 466–67 water quality and supply, 467 grassland ecosystem functioning, 461 decomposition, 462–63 net primary production (NPP), 461–63 nitrogen cycling, 463–64 water balance, 464–65 grassland structure and function interactions among biotic and abiotic factors determining, 460–61 local controls over fire, 454–55 topography and soils, 454 weather variations, 454, 455f, 468, 469f grassland subtypes. See also grassland types coastal, 452 determined by unique soils, 453 interior, 450, 452 grassland sustainability, 840 grassland systems on continuum from equilibrium to nonequilibrium, 838–39 grassland types, 220. See also grassland subtypes diversity of, 450, 453f variations within, 454–55 grasslands, 220–21, 449–50, 470–71. See also rangelands; specific topics controls over structure and function of, 450, 452f local, 454–55 human impacts on, 468–69 seasonal variations in temperature and precipitation driving plant growth, 450, 452f Grasslands Ecological Area, 672 grasslands management under future conditions, 469 interaction of changing precipitation patterns and non-native vs. native plants, 469 managing for grassland resilience in the face of multiple environmental changes, 470 managing fragmented grasslands for diversity, 470 gray-crowned rosy finch, 201, 628 gray fox, 400, 442, 487t, 514, 658, 724 gray jay, 200 gray owl, great, 199, 201 gray pines, 56, 218, 220 gray squirrel, eastern, 231 gray vireo, 202 gray whale, 207, 297, 298, 300, 792 gray wolf, 204. See also grey wolf Grayia spinosa, 147, 648, 649f grazers, 318–20, 343 grazing. See also livestock grazing in coastal sage scrub, 440 grazing distribution, 845

grazing effects, factors that influence, 851 grazing impacts on vegetation, ways to manipulate, 851 grazing period, 845 grazing phenomena affecting ecosystems, 845 grazing practices, traditional cautions and implementation of, 851 grazing pressure, 845 greasewood, 145, 148, 646 Great Basin, 194, 202, 601f lakes, 696–702 Great Basin bristlecone pine, 224 Great Basin Desert, 638. See also deserts Great Basin desert scrub, 661 Great Basin sage, 636, 642 great blue heron, 323 great egret, 371 great gray owl, 199, 201 great horned owl, 459, 488t, 514 great white shark, 323 greater roadrunner, 201 greater sandhill crane, 201, 680 grebes, 680, 701, 702 Clark’s grebe, 372 green abalone, 320 green algae, 317, 318, 349, 643, 644, 716, 717, 719, 721, 722, 733. See also yellowgreen algae marine, 722 green anemone, giant, 352 green crab, 371 Green Diamond Resource Company, 828–29 green infrastructure, 888 green sea turtle, 199 green sturgeon, 379t, 728 green sunfish, 242 Greenbelt Alliance. See Bay Area Greenbelt Alliance Greenberg, R., 375 Greene, E. L., 2 greenhouse gases (GHGs), 911t, 917 greenlings kelp greenling, 322, 326, 327f painted greenling, 322 Greenprints, 919, 920 grey fox. See gray fox grey pine. See gray pine grey whale. See gray whale grey wolf, 164 Griffin, J. R., 515, 851 Griggs, G., 422 Grinnell, George, 603 Grinnell, Joseph, 2, 254 grizzly bears, 164, 204, 514, 541, 680, 724, 852 California grizzly bear, 509, 926 groins, 393 grosbeaks black-headed grosbeak, 256t pine grosbeak, 201, 596 Grossulariaceae, 190t ground-dwelling squirrels, 160 ground fires, 28–29 ground sloths, 157f, 162, 163, 173 Shasta ground sloth, 654 ground squirrels, 163, 164, 457f, 458, 459, 487t, 628, 851 antelope ground squirrel, 654, 657 Belding’s ground squirrel, 594, 596, 627 California ground squirrel, 400, 514, 516 Mohave ground squirrel, 206t groundfish fishery, catch shares in U.S. West Coast, 803 groundfishes, 290, 779, 787, 788, 790f, 791f, 794–95, 802, 804

INDEX  959

groundwater contamination by nitrogen, 879 groundwater overdraft, 878 group selection, 825 grouses Mount Pinos blue grouse, 201 ruffed grouse, 200 sage grouse, 37, 225, 628 sharp-tailed grouse, 202 growing seasons, 434, 457f, 594, 875 chaparral and, 434, 490, 495 fire and, 35, 36, 222, 487, 495, 521, 569, 598 ice and, 696, 703 length, 221, 222, 316, 430, 442, 454, 461, 467–69, 519, 569, 579, 583, 588, 593, 598, 605, 614, 624, 630, 703, 849, 857 nitrogen and, 463, 467, 469, 696 seed production, seedlings, and, 462, 487 snow and, 253, 554, 555, 583, 598, 624, 677 temperature and, 161, 454, 490, 510, 579, 580, 600, 614, 622, 624, 630 topography and, 454 water and, 50, 215, 221, 450, 452, 454, 461, 462, 464, 465, 467–69, 490, 510, 520f, 554, 555, 569, 583, 587, 588, 594, 600, 619, 622, 624, 670, 674, 721 grubs, 459 grunions, 779 California grunion, 401 Grus canadensis canadensis, 201 Grus canadensis tabida, 201, 680 Guadalupe fur seal, 792 Guadalupe Island Cypress, 193 Guadalupe Island palm, 193 Guadalupe Island pines, 193 Guadalupe savroy, 193 guillemot, pigeon, 294 Gulf of the Farallones, 296 Gulf of the Farallones National Marine Sanctuary (NMS), 301 gull-billed tern, 202 gulls, 202, 399–401 California gull, 202, 701 Gulo gulo, 596t, 628 gumboot chiton, 318 gunnels, 322 Gutierrezia sarothrae, 646 Gymnogyps californianus, 188, 200, 266, 400 gymnosperm, 132, 138 Gymnothorax mordax, 323 Haagen-Smit, Ariel, 108 habitat, defined, 213 habitat connectivity, 444 Habitat Conservation Plans (HCPs), 195, 442, 443, 828–30, 903, 911t, 918–19 habitat fragmentation, 931 habitat protection, 84–85. See also conservation Habropoda pallida, 652 Hadrotes crassus, 398f, 399 Haematopus bachmani, 344, 400 hairless popcorn-flower, 190t hairy tiger beetle, 415 hake, 293 Haliaeetus leucocephalus, 188, 772 halibut, California, 371, 372f, 373, 381, 769, 790, 795 Halichoeres semicinctus, 322 Haliotis, 326, 350 Haliotis sp., 784 Haliotis spp., 767

960  INDEX

Haliotis corrugata, 320 Haliotis cracherodii, 320, 344, 769, 797 Haliotis fulgens, 319f, 320 Haliotis kamschatkana, 319f, 320 Haliotis rufescens, 319f, 320, 786 Haliotis sorenseni, 196, 319f, 769, 796 Haliotis walallensis, 320 Hall, D. C., 338 halophile invertebrates, 194 halophyte, 733 Halpern, B. S., 299–300 hamamelid, 134 Hamerlynck, E. P., 647 hand-harvest methods (fishing), 785 Hannah, L., 439 Hapke, C. J., 422 harbor seals, 323, 401 hardhead, 726 Hardie, L. A., 704 Harding grass, 461 hardstem bulrush, 678f hardwood lands, private, 526 historical and projected development of, 526 hardwood rangelands, land use in, 85–86 hares, 162, 163, 455. See also sea hares European hare, 759t white-tailed hare, 589 harrier, northern, 459 Harrison, Benjamin, 901 Harter, T., 879 harvest mice, 486, 487t salt marsh harvest mouse, 205, 206t, 376t, 379t, 680–81 western harvest mouse, 376t, 759t, 765 harvester ant, 239 hawks, 37, 459, 488t, 514, 656 red-tailed hawk, 459, 488t, 514, 628, 724 Swainson’s hawk, 459, 465 hawthorne, 139 Hazardia squarrosa, 432f, 435, 437 hazels California hazel, 541–42 witch hazel, 134 heat capacity of water, 14 heat island effect, 498 heat waves, 20–21 Heermann’s kangaroo rat, 206t Heizer, R. F., 171 Helianthemum scoparium, 437 Heliodinidae, 415 Heliscomys, 157f hellgrammite, 723f Heloderma suspectum, 198, 653 Helodermatidae, 198 Hemiauchenia, 162 Hemicyoninae, 161, 369 Hemigrapsus oregonensis, 371 Hemilepidotus, 319f Hemilepidotus hemilepidotus, 323 Hemilepidotus spinosus, 323 Hemipodia spp., 399 Hemizonia, 450 Hemizonia congesta, 232 hemlock, 139, 147, 148, 218, 598 Hepatic tanager, 202 herbivores. See also specific topics fire and, 37 large, in deserts, 657–58 large native, and livestock production, 852 Hereford, R., 647 hermit crab, 343, 399 herons, 372 herpetofauna, 188, 196–99, 437, 541, 726 herrings, 293, 371, 379

Pacific herring, 373, 790f Hershfeldia incana, 435f Hesp, P. A., 412 Hesperocyparis, 36, 483, 484t Hesperocyparis abramsiana, 215 Hesperocyparis forbesii, 40 Hesperocyparis macnabiana, 536 Hesperocyparis macrocarpa, 215 Hesperocyparis pygmaea, 536 Hesperocyparis sargentii, 536 Hesperodiaptomus shoshone, 698f Hesperolinon, 192 Heterobasidion spp., 564 Heterodontus francisci, 322 Heteromeles arbutifolia, 433t, 483, 484t, 492t, 830 Heteromyidae, 206t Heterostichus rostratus, 317f heterotrophic bacteria, 629 heterotrophic ecosystems, 370, 697, 888 Hewatt, Willis G., 338 Hexagrammos decagrammus, 319f, 322 Hexagrammos superciliosus, 319f, 322 Hexapoda, 723f Hickey, B. M., 101 Hickman, Jim, 5 hickory, 138, 139 high tide strand lines (HTS), 392 highway iceplant, 418–21 highways. See transportation legislation hillside collinsia, 190 Himantopus mexicanus, 685 Hipparion, 157f Hipparion, 161 hippid crabs, 395 Hippodiplosia, 321 hippopotamus, 157f, 161 Hirschfeldia incana, 433t Hirudinea, 723f Hirundo pyrrhonata, 724 Hirundo rustica, 400 Hispanic settlers, 85 historic mean fire return intervals (HFRI), 557 historical ecology. See also vegetation prehistory; specific topics methods of, 133–34 historical vegetation, diversity of proxies used in, 133 Histosols, 50, 54, 55f, 57, 68 Hobbs, R. J., 851, 932 Hoerling, M. P., 19–20 Hoffmann’s slender-flowered gilia, 764, 765f hogs, 868 Holbrook, S. J., 316 Holcus lanatus, 461, 850 holdfast, 312 holly-leaved cherry, 484t Holocarpha, 192 Holocene, 54 early Holocene, 146–47 archaeological sites and isolated finds, 172f. See also Holocene paleontological sites late Holocene, 147–48 mammals in, 163–64 mid-Holocene, 33 Holocene paleontological sites, 158f. See also Holocene: early Holocene Holton, B., Jr., 415 Holuthuriae, 311 Homestead Act of 1862, 87, 901 Homo sapiens, 164 homopterans, 651 Honckenya peploides, 417 honeybees, 271 European honey bee, 195, 767

hook-and-line methods (fishing), 785 Hoover’s cryptantha, 189t Hoover’s downingia, 677f hopbush, 139 hoppers, beach, 397f, 398, 399f hopsage, spiny, 147, 648, 649f Hordeum murinum, 238, 458, 513 Hordeum spp., 461 horn shark, 322 Horne, A. J., 694 horned larks, 201, 438f, 459, 656 Island horned lark, 765 horned lizard, coast, 239, 459 horned owl, great, 459, 488t, 514 horned rodents, 159 hornmouth, leafy, 321 horse clams, 371 horses, 157f, 159–63, 173, 233, 455, 456, 522, 759t, 840–42, 852 feral horses, 657 horsetail, 134 horticulture. See also agriculture early, 870 house mouse, 759t house wren, 256t Howland Flat, 87 Hubbard, D. M., 401 huckleberries, 540f California huckleberry, 542 evergreen huckleberry, 539, 540f, 829 Huff, M. H., 37 Hugenholtz, C. H., 412 human population. See population, human Humboldt County, 542f Humboldt lily, 772 Humboldt marten, 541 Humboldt Redwoods State Park, 219f forest floor at, 542f Humboldt squid, 293, 295 Humboldt State Park, 545, 547f hummingbirds, 648 Allen’s hummingbird, 765 Anna’s hummingbird, 438f, 488t Costa’s hummingbird, 488t humpback whale, 292f, 297, 298, 300, 792 hunting, 178 maritime, 767, 769 Huntsinger, Lynn, 524 Huyer, A., 101 hyacinth, water, 682 hybridization, 237 hydric soil, 670, 674 hydrilla, 682 Hydrilla verticillata, 682 hydrocorals, 320 hydrodynamics, 363 hydrographs, 714, 716 Hydroides elegans, 328 hydroids, 320, 324, 343, 371 hydrologic regimes in rivers, 714, 716 hydrologic regions of California, 715f hydrologically closed basins, 694 hydrology. See also under chaparral fire effects on, 38–39 freshwater wetlands, variability in, 677–78 oak woodlands, 524 subalpine forest ecosystem, 599–600 wetlands, 670, 673 Hydromantes, 197 Hydromantes brunus, 198 Hydromantes shastae, 198 hydroperiod, 670, 673 hydrophytic vegetation, 670, 674 Hydroporoni, 698f Hydropsychidae, 723f hydrozoans, 371

hyenas, 157f, 162, 163 Hygrotus artus, 189t, 194 Hymenoclea, 648 Hyperacanthomysis longirostris, 371 Hyperaspis annexa, 415 Hyperbaena, 138 Hypericum perforatum, 242 hyperparasitoids, 650f hypersaline estuarine habitat mosaics, 376 hypersaline waters, 694 Hypertragulus, 157f, 160 Hypochaeris spp., 847 Hypolagus, 157f, 161 hypolimnion, 698 Hypomesus pretiosus, 399f Hypomesus transpacificus, 371, 372f, 379t hyporheic habitats, 714 hypoxia, 100, 703 Hypsypops, 319f Hypsypops rubicundus, 322 ibis, white-faced, 201 Icaricia icariodes missionensis, 195 Icaricia icariodes pheres, 195 Ice Age (last glacial period), retreat from the, 145–48 Ice Age rollercoaster, 139–42 Ice Ages, Pleistocene. See Pleistocene Ice Ages ice-off, 703 iceplants, 235 Icerya purchasi, 242 ichthyoplankton, 290 Ichthyornis, 199 Ictalurus spp., 726 Idaho fescue, 225 Idotea resecata, 372 iguana, desert, 653 Iguanidae, 198 illuvial clay, 49 Ilyanassa obsoleta, 371 Important Bird Areas (IBAs), 200–202 incense cedar, 32, 144–48, 221, 536, 537t, 554t, 557, 559, 560t, 817, 821, 825 Inceptisols, 50, 55, 57–60, 66–68, 555 incorporation, 913 index of farmland values, 871 Indian ricegrass, 648, 649f indigenous California, 78–79, 169–70, 173– 74, 180–81. See also specific topics history of research, 170–71 indigenous ecology, alternate views of, 177 controlled burning, 179–80 emergent hypermanagement, 177 furthering ecomanagement, 179 megafaunal extinctions, 177–78 tempered overkill, 178–79 indigo bush, 147 indirect effects, 347–48, 773 individual fishing quotas (IFQs), 803 individual trees, clumps of trees, and openings or gaps (ICO), 561, 568 individually transferable quota (ITQ), 802 indricotheres, 160 infiltration, 449 infrared radiation, 18 infrastructure, land use for, 84 inhalable particulate matter (PM10). See particulate matter Inonotus andersonii, 515 Inonotus dryophilus, 515 insect gall, 515 insect pests, 238 Insecta, 723f Insectivora, 595t insectivores, 157f, 158, 160, 162, 203, 459, 486, 541, 657, 720, 724, 726, 732

insects, 599. See also specific topics and grassland distribution, 458–59 Inshore Countercurrent (IC), 97 insolation, 110 integrated water vapor (IWV), 15f intentional introductions, 231 interdisciplinary conservation science, 937–38 priorities for, 938 interglacials, 256 interior grassland, 450, 452. See also valley grassland interior live oak, 510, 512, 517t, 559 international agreements, 928–29 International Convention on Biodiversity (ICB), 928–29 interspaces, 63 intertidal ecology, 337–38, 346, 351–52. See also intertidal ecosystems, rocky climate change and, 350 future scenarios, 350–51 interactions between stressors, 351 community dynamics, 346–48 community regulation, 348–49 human impacts, 349 management, 350 recreational human visitation and exploitation, 350 water pollution, 349 patterns on the shore, 346 recruitment dynamics, 348 intertidal ecosystems, rocky, 337–38, 346, 351–52. See also intertidal ecology organisms, 343 carnivores, 343–44 grazers, 343 primary producers, 343 species of special concern, 344 suspension feeders, 343 intertidal habitats, 371 intertidal isopods, 398 intertidal research, rocky significance and history of, 338 intertidal weevil, 398 intertidal zones, 346, 347f introduced species, 230–31, 235, 377, 378f, 380, 419, 756, 759–60t, 768–69. See also invasive species; trout introduced vascular plants, origin of, 229–30 invader management biocontrol, 241–42 challenges and success stories, 238, 240–42 early detection, rapid response, and eradication, 240–41 federal agencies, 238, 240 long-term control and management, 241 prevention, 238, 240 invasion ecology, 230–31 invasion history, California’s, 231–32 invasions, biological, 229 degree of invasion in ecosystems, 235–36 ecological and anthropogenic drivers of, 232–33 anthropogenic drivers and pathways, 233–35 future of, 242–43 invasive grasses and forbs, 850 invasive plant species in wetlands, 681 Invasive Spartina Project, 682 invasive species, 229, 419, 603, 629 climate change and, 260–61 harmful, 230 impacts of, 236–38 terminology, 229, 230 Invasive Species Program of CDFW, 241

INDEX  961

invasiveness and ecosystem invasibility, mechanisms governing, 232, 234t inverse stratification, 696 inverse texture hypothesis, 640 invertebrates, 193 benthic, 370, 371 biogeography, 194–95 birds feeding on beach invertebrates, 399 conservation context, 195–96 in deserts, 194–95 endangered stream, 724 in estuaries, 370–71 evolutionary diversification, 193–94 in oak woodlands, 514, 515 pelagic, 370, 371 in redwood forest, 541 riverine food webs and, 717, 722–24 in wetlands, 681 ion exchange capacity, 38 ips, California fivespinned, 564 Ips paraconfusus, 564 Ips spp., 37 Irish lords, 323 ironwood, 139, 148, 636, 646 Catalina ironwood, 765 irrigation, 866, 871, 880 irrigation water damage, 878–79 Irvine, I. C., 437 Irvingtonian Land Mammal Age, 163 Ishi, 176f Island Chumash, 768 island ecosystems. See also California Islands; Channel Islands of California how development changed the functioning of Chumash era, 768 fishing, 769–70 introduced species, 768–69 maritime hunting, 769 ranching, 768 island foxes, 204, 206t, 764, 773 dwarf island fox, 400 island gray fox, 206t Island horned lark, 765, 773 island invasions, 240 island Jepsonia, 772 island loggerhead shrike, 769, 773 island night lizard, 199, 765 island oaks, 764, 765f island rush-rose, 772 island scrub-jay, 200, 201, 764–65, 773–74 measures to prevent extinction of, 774 islands in transition, 770 conservation management actions active restoration, 772–73 establishment of marine reserves, 771 infrastructure removal and amelioration of contaminants, 772 legislation for protection of at-risk species, 771–72 passive recovery, 772 recovery in an uncertain future, 773–74 reestablishing fog, 772 removal of non-native species, 770–71 principles and predictions, 770 “islands of fertility,” 645, 646 Isocheles pilosus, 398f, 399 isopods, 343, 371, 372, 395, 397f, 398, 399, 401, 658, 720, 722 isotherms, 257 Isurus oxyrinchus, 293, 795 Italian ryegrass, 458, 466 ivy, cape, 235, 238

962  INDEX

jack mackerel, 321 jackfruit, 138 jackrabbits, 487t, 659 black-tailed jackrabbit, 417, 418, 654, 657 white-tailed jackrabbit, 627t Jacobs, D. F., 541 Jacobson, L. D., 793 jaguar, 164, 204 Japanese eelgrass, 380 jays, 200. See also scrub-jays pinyon jay, 656 Stellar’s jay, 540f Jefferies, R. L., 413 Jeffrey pine, 36, 56, 109t, 113, 114, 117, 216f, 221, 224, 536, 554t, 557, 559t, 560–61, 565f, 566, 589, 592 Jeffrey pine beetle, 564 Jelinek, L. J., 870–71 Jellison, R., 706 jellyfish, 99, 371 Jenny, H., 67 Jepson, Willis Lynn, 2 Jepson Manual, The (Baldwin et al.), 190 Jepsonia, island, 772 Jepsonia malvifolia, 772 jet stream, 14 jewelflower, 192 Johanneson, K. H., 700 Johnson, A. F., 410, 415 Johnston, W. E., 882 Joint Working Group on Vessel Strikes and Acoustic Impacts (JWG), 301–2 Jones, Terry L., 177 Jordan, David Starr, 1 Joshua tree, 145–47, 225, 636, 642, 647, 654 Juglandaceae, 134 Juglans, 138 Juglans californica, 512 jujube, 493 jumbo squid, 798 Junco hyemalus, 628 juncos, 543 dark-eyed junco, 628 Juncus lescurii, 417 June beetles, 193 juniper-cypress, 145 junipers, 36, 144f, 145–48, 224, 642, 647, 675. See also pinyon-juniper one-seeded juniper, 636 Sierra juniper, 222, 580, 589, 591, 592, 605 Utah juniper, 141, 145, 146 western juniper, 147, 560t Juniperus, 144f Juniperus spp., 36, 642 Juniperus californica, 146 Juniperus grandis, 222, 580 Juniperus monosperma, 636 Juniperus occidentalis, 147 Juniperus occidentalis var. australis. See Juniperus grandis Juniperus osteosperma, 141, 224 Kajikia audax, 292 Kalmia polifolia, 594 kangaroo mice, 162, 163 kangaroo rats, 162, 163, 203, 205, 206t, 438f, 457f, 458, 487t, 847, 851 chisel-tooth kangaroo rat, 657 Stephen’s kangaroo rat, 206t, 207, 443 kaolin, 50 Katsuwonus pelamis, 795 katydid, Antioch Dunes shieldback, 189t, 195 Keckiella antirrhinoides, 432f

Keeley, Jon E., 180 Kelletia, 319f, 321 Kelletia kelletii, 321, 345, 797 Kellet’s whelk, 321, 797 Kellicottia spp., 698f Kelly, A., 259 kelp bass, 795 kelp crab, northern, 318 kelp distributions (within their geographic range), environmental determinants of, 314–16 kelp flies, 399 kelp forest community structure abiotic determinants of, 323–25 biotic determinants of, 324–25 disturbance, forest dynamics, and shifts in, 325 kelp forest ecosystem services, 325–26 kelp forest ecosystems, trophic structure and functional attributes of, 316 apex predators, 323 detritivores, 320 grazers, 318–20 planktivores, 320–21 primary producers, 317–18 small mobile carnivores, 321–22 tertiary consumers, 322–23 kelp forest fishes, cryptic coloration of, 317f kelp forests, 311–12, 329–30 beaches and, 395, 397, 398 ecosystem-based management and marine protected areas, 328–29 impending challenges, 329 El Niño and, 316, 325, 328, 329 geographic distribution, 313–16 human impacts to, 326–27 invasive species and, 327–28 phenology, 316 restoration, 326–27 kelp greenling, 322, 326, 327f kelp harvesting, 325–26 “kelp highway,” 766 kelp perch, 317f “kelp rafts,” 318 kelp rockfish, 323, 327f kelp species, 311–12, 314, 315f geographic range, 313, 314t giant kelp, 311–18, 323–29, 395f, 397f, 768, 769 subsurface canopy-forming, 314 surface canopy-forming, 313 kelp species composition, geographic variation in, 317 kelp wracks, 397, 398, 399f kelpfish, giant, 317f kelps, 312. See also specific topics bladdered kelps, 343 bull kelp, 312, 313, 315–17, 323, 324, 328 giant kelp, 311–18, 323–29, 395f, 397f, 768, 769 Kelvin waves, 98 Kennedy, C. A., 888 Kennedy, P. G., 540 Kennett, D. J., 173 Kennett, J. P., 173 Kent, William, 89, 549 Keratella quadrata, 698f Keratella spp., 698f Keratella taurocephala, 696 Kesterson Lake, 879 kestrel, American, 488t keystone ecosystems, 589 keystone predators, 516 keystone species, 268, 347, 509, 769 killdeer, 400, 401

“killer alga,” 241 killer whales, 203 transient killer whale, 292f killifish, California, 371 King, C., 179 King, J. R., 101 kingbirds, 256t, 400 Kinlan, B. P., 316 Kinney, Abbot, 83 Kinosternidae, 727f kit foxes, 658, 847 kite, white-tailed, 459 kiwi, 272 Klamath Basin, 672, 680, 728 Klamath Lake, 672 Klamath mixed conifer, 557, 558f Klamath Mountain soils, 57–58 of metamorphic terrain, 57 of ultramafic terrain, 56–57 valley soils, 58 Klamath Mountains, 535 elemental composition of rock types in, 56, 57f geology and geomorphology, 56 Klamath Region, 828 Klamath River, changes in water quality from headwaters to mouth, 728–29 Klamath weed, 242 Knapp, R. A., 697 knobcone pine, 220, 483, 536 Koeleria macrantha, 513 kokanee salmon, 699 Kostyack, J., 918 Krascheninnikovia lanata, 147, 648, 649f Krausman, P. R., 852 Kremen, C., 271 krill, 294–96, 299, 799 Kroeber, Alfred L., 171, 177, 181 Kroeger, T., 524 krummholz, 222–23, 253, 579, 589, 590, 614, 615f Kueppers, L. M., 525 kumamoto oyster, 798 kyphosids, 325 La Niña, 16, 98, 131–32, 619 La Perouse, Jean Francois, 177 labrids, 325 Lacey Act of 1990, 238 Lactuca, 450 lacustrine deposits, 60 lacustrine systems, 670 ladybeetle, Australian vedalia, 242 Laetiporus gilbertsonii, 515 Lafferty, K. D., 421 Lagomorpha, 203, 205f, 595t, 759t lagomorphs, 157f, 162, 204, 626, 628, 653 lagoons, 365–66 coastal, 703 Lagopus leucurus, 629 Lahontan cutthroat trout, 699 Lahontan Tui chubs, 725 Lake Tahoe, 699f, 916 ecological services provided by, 705 vertical mixing, clarity, and responses to invasive species, 697–99, 700f lake trout, 699 lakes, California, 693–94, 706 ecological services provided by, 705 future scenarios as function of climate and land use changes, 705–6 high-elevation, 696–97 limnological concepts and, 694 map of, 695f regional characteristics, 696

Central Valley, 704–5 coastal ranges, 702–3 deserts, 703–4 Sierra Nevada and western Great Basin, 696–702 runoff and, 693, 696, 698, 702, 703, 705, 706 lamb, 841, 868t Lamiaceae, 190t, 485t, 486t laminar, 391 Laminaria, 315, 316, 343 Laminaria farlowii, 314, 315f, 317, 319f Laminaria setchellii, 313f, 314, 315f, 317, 319f Laminariales, 312f Lamna ditropis, 293 Lampetra tridentata, 725, 728 Lampetra tridentata ssp., 725 lampreys, 725, 727f Pacific lamprey, 724, 725, 728 Lampris guttatus, 786 Lampropeltis getula, 514 Land Conservation Act. See California Land Conservation Act of 1965 land conservation history of California, 903–7 reconstructing, 902–3 land cover change, 77–78 land leveling, 52–53 land management and the future of deserts, 661–62 in montane forests, sensitive species that affect, 561, 563f of public lands, 900 Land Management, Bureau of. See Bureau of Land Management land use, 75, 90–91 agricultural valleys, 89–90 categories of, 84 coastlines, 82–84 deserts, 84–85 fiscalization of, 916–17 hardwood rangelands, 85–86 meaning and terminology, 77–78 mountains, 86–88 northwest forests, 88–89 land use planning, 899–901, 921 future challenges, 920–21 innovations toward a change, 918–20 key concepts in, 899–901 regulations, 899–901, 909–11t timeline of legislation, 909–11t land use planning and policy in California, evolution of, 907–8 attempts to regulate urban growth: 1940s1980, 911–15 Progressive Era legacy: pre-1940s, 908, 911 1980s and 1990s, 915 landlocked cockeye salmon, 699 landscape-level planning, 918–19 landslide topography, 57 Lange’s metalmark butterfly, 195 language group migrations into California over late Pleistocene and Holocene, 174 language territories of indigenous California, 169, 170f Lanius ludovicianus, 656 Lanius ludovicianus anthonyi, 773 Lanius ludovicianus mearnsi, 769 lapse rate, 254 large white anemone, 321 largemouth bass, 242, 699 lark sparrow, 459 larks. See horned larks Larrea tridentata, 142, 225, 636, 649f, 654, 655f

Larus californicus, 701 Larus delawarensis, 399f Larus livens, 202 larval insects, 515 larval oysters, 796 Lasiurus blossevillii, 724 Lasiurus xanthinus, 654 Lassen Volcanic National Park, wildfire sunrise in, 3f Last Glacial Maximum (LGM), 140, 145, 146, 173, 257, 619 Lasthenia, 192 Lasthenia fremontii, 677f Laterallus jamaicensis, 201, 680 Laterallus jamaicensis coturniculus, 379t, 854 Lathyrus littoralis, 414f Latinos. See Hispanic settlers latitudinal shifts of distribution ranges in low-relief areas, 141–42 Latrodectus, 650f Latrodectus hesperus, 650f laughing gull, 202 laurel sumac, 433t, 483, 491, 493 laurels, 138. See also bay laurel; California bay laurel Laurus, 138 lava. See volcanic rock Lawton, H., 177 Layia, 192 layia, beach, 414f Layia carnosa, 192, 414 Layia discoidea, 192 Layia gaillardiodes, 192 Layia glandulosa, 192 Laysan albatross, 294 lazuli bunting, 256t, 438f, 488t Le Conte’s thrasher, 199 Leach’s storm petrel, 202, 294, 296 lead, 118, 122t leaf area index (LAI), 518 leaf gallers, 486 leaf miners, 486 leaf-nosed bat, California, 657 leaf stress traits, 490, 491t leafcutter bees, 417 leaflike blades. See blades leafy hornmouth, 321 least Bell’s vireo, 201, 202 least bittern, 680 least sandpiper, 202, 372, 373f least tern, California, 201, 379t, 401 leather fern, 539 leather oak, 55 leatherback sea turtle, 293f, 294, 298 leaves, phases in the decomposition of, 497 LeConte, John, 694 LeConte, Joseph, 2 Ledum glandulosum. See Rhododendron columbianum leeches, 681, 717, 723f Leet, W. S., 791, 797 leeward side, 18 legacy trees, 541 legless lizards, 196, 415 burrowing legless lizard, 418 California legless lizard, 198 legumes, 415, 453, 454–58, 462, 463, 466– 68, 496t, 645 leguminous plants, 66 leguminous trees, 636 Leiopelma, 198 Leiopelma sp., 197 Lemna minor, 679 lemonade bush, 483, 484t, 491 lemonadeberry, 433t

INDEX  963

lemons, 89 Lenihan, J. M., 41, 525, 559 leopard lizard, blunt-nosed, 198, 199, 847 leopard shark, 373, 796 Lepidium latifolium, 682 Lepidium nitidum, 458 Lepidopa californica, 398f, 399 Lepidopteran pests, 272 Lepidurus packardi, 681, 681f Lepomis spp., 726 Leporidae, 595t, 764 leporids, 157f, 160 Leptochiton spp., 342 Leptocottus armatus, 371 Leptodiaptomus signicauda, 698f Leptographium wageneri, 599 Leptostyne gigantea, 764 Leptotyphlopidae, 198 Lepus sp., 487t Lepus americanus, 595t Lepus californicus, 417, 457, 487t, 595t, 657 Lepus townsendii, 589, 595t, 627t lesser sandhill crane, 201 lettuces miner’s lettuce, 542 sea lettuce, 352 wild lettuce, 450, 458 Leucophaeus atricilla, 202 Leucosticte tephrocotis, 201, 628 Leuresthes tenuis, 399f, 401, 779 Levine, J. M., 436, 764 Lewis, Henry, 177–79 lichens, 644, 645 cyanolichens, 539 epiphytic lichens, 116, 117, 121, 123, 509– 10f, 513 life cycle assessment (LCA), 888 life history of species, 781 light-footed clapper rail, 201, 379t Lightfoot, K. G., 180 lightning, 22 lightning-ignited fires, 31 lignotubers, 35, 486, 543 lilac, California, 139, 192, 480, 484t Liliaceae, 189t lilies Humboldt lily, 772 single-flowered mariposa lily, 189t yellow pond lily, 678f Lilium humboldtii, 772 limber pine, 146, 147, 224, 580, 582, 586, 588, 589, 593, 596, 598, 599, 605 limber pine forests, 590–91 liminology, 694 Limnanthes, 192 Limnodromus scolopaceus, 399f Limnoithona tetraspina, 371 Limosa fedoa, 399f limpets, 318, 337, 338, 342, 343, 350 Lin, J.-L., 700 Linanthus nuttallii subsp. pubescens, 625 Lincoln, Abraham, 901, 902 Lincoln’s sparrow, 201 Lindegren, M., 793 line fishing. See hook-and-line methods Linepithema humile, 239, 770 lingcod, 323, 327f, 786, 790, 795 lion, mountain, 203, 487t, 527, 626, 627t, 658, 852, 937 Liparididae, 322 Liquidambar, 138 Lissocrangon stylirostris, 399 Lithariapteryx, 415

964  INDEX

Lithobates catesbeianus, 199, 237–38, 727 Lithocarpus densiflorus, 145, 236, 546f, 829. See also Notholithocarpus densiflorus lithologies, 47 litter, 222, 434 Little Ice Age, 132, 173, 620 littoral cells, 360, 391, 393 littoral currents, 391, 394, 403 littoral transport, 391, 393 Littorina, 343 Littorina littorea, 380 Littorina spp., 342 littorine snails, 342 live oaks, 220 canyon live oak, 218, 510, 512t, 559 coast live oak, 214f, 236–37, 483, 484t, 495, 510, 511f, 512, 514f, 515, 517, 518, 521–23, 541, 548f, 849 interior live oak, 510, 512, 517t, 559 liverworts, 716, 848, 853 livestock. See also rangelands California Gold Rush and, 522, 767, 837, 842, 849 characteristics and distribution of, 840–41 livestock confined animal feeding operations (CAFOs), waste disposal from, 879, 880 livestock grazing. See also grazing dunes impacted by, 418–19 geography of, 836–37, 839–41 regional differences, 841 in subalpine forests, 603 livestock operators and ecosystem services, 854 livestock production atmosphere and, 854 ecosystem services and, 846–54 native carnivores and, 852 native herbivores and, 851–52 native plants and, 847–51 soil system and, 852–53 water quality and, 853–54 livestock products, relative shares of, 874 livestock ranching, 767 land use for, 84 lizardfish, California, 323 lizards, 196, 198, 199, 322, 459, 647, 652, 653, 656, 657, 659, 720, 721, 723, 724, 765. See also legless lizards; sideblotched lizards blunt-nosed leopard lizard, 198, 199, 847 coast horned lizard, 239, 459 island night lizard, 199, 765 western fence lizard, 438f, 459, 486 llamas, 522, 840 Loarie, S. R., 258, 570 lobsters, 320–22, 343, 756. See also spiny lobsters local agency formation commissions (LAFCOs), 913 localities, divestment of responsibility to, 908–9 locust, black, 139 lodgepole chipmunk, 596 lodgepole pine, 147, 216f, 222, 253, 536, 557, 560t, 566, 580, 582, 586, 589, 592, 597, 598, 603, 604 lodgepole pine forests, 592–93 loess, 60 log spiral, 390 loggerhead sea turtle, 294, 300 loggerhead shrikes, 656 island loggerhead shrike, 769, 773

logging, 547. See also timber harvest California Gold Rush, lumber, and, 548, 818, 820 in northwest forest during late 19th century, 88, 89f logging railroad, 821f Loligo opalescens, 769. See also Doryteuthis opalescens London, Jack, 5 long-billed curlew, 201, 372, 373f long-billed dowitcher, 399f long-range transport, 112 long-tailed weasel, 627t long-toed salamander, Santa Cruz, 198 longfin smelt, 379t longline fishing, 785 longshore sediment transport, 391 Lontra canadensis, 681, 728 Lophogorgia, 319f lords, Irish, 323 Lorimer, C. G., 542 Los Angeles, population density and greenhouse gases emissions in, 889t Los Angeles County imports, local sources, and outputs of energy, water, and food in, 888t Los Angeles sprawl, 80, 81f Lost River sucker, 728 lost thistle, 189t Lottia, 343, 350 Lottia gigantea, 343 Lottia spp., 342 Lotus nuttallii. See Acmispon prostratus Lotus scoparius. See Acmispon glaber Louisiana red swamp crayfish, 722, 726 Lower Arm (Clear Lake), 702, 703f lower montane zone, 221 Loxia curvirostra, 596 Luck, G. W., 889 Ludwigia hexapetala, 682 lumber, logging, and California Gold Rush, 548, 818, 820 lumber industry, 549. See also timberlands lupines, 466 chamisso bush lupine, 416, 417 Tidestrom’s lupine, 409–10f, 414f, 419 Lupinus, 625 Lupinus arboreus, 416, 417f Lupinus chamissonis, 416 Lupinus nipomensis, 414 Lupinus tidestromii, 409f, 414 Lxobrychus exilis, 680 Lycaenidae, 189t, 415 Lycium andersonii, 142, 648, 649f Lycium californicum, 769 Lycium verrucosum, 189t lycosid spiders, 721 Lyme disease, 515 Lyngbya, 720 Lynn, R. J., 96–97, 101 Lynx rufus, 487t, 514, 516, 658, 852, 935 Lyonothamnus, 139 Lyonothamnus floribundus, 765 Lyons, W. B., 700 lysimeter, 498 Lytechinus, 319f Lytechinus anamesus, 318 MacCall, A. D., 793 MacGillivray’s warbler, 256t MacGinitie, George, 338 MacKay, M. D., 705 mackerels, 791f jack mackerel, 321

Pacific mackerel, 790f, 793 snake mackerel, 795 Macnab’s cypress, 536 macroalgae, 312, 313f, 316–18, 322–24, 325f, 345f, 349, 368–69, 370f, 377, 381, 392, 397, 717, 726 beach-cast, 395f green, 392, 397, 719, 721 macroalgae species, 314, 315f, 324 macroalgal wrack, 392 macroarthropods, 658 macroclimatic patterns, 216 Macrocystis, 313, 313f, 317, 319f, 400 Macrocystis sp., 268 Macrocystis angustifolia, 313 Macrocystis integrifollia, 313 Macrocystis pyrifera, 311, 312f, 313, 314t, 316f, 397f, 398, 766 macrofauna, 515 macrofungi, 414 macrophyte wrack, 392, 395, 397, 400 macrophytes, 345, 346, 349, 395, 398, 400, 679, 683f, 702, 716, 723 Macrotis californicus, 657 Mad River, 549f Madia, 192, 450 madrones, 139, 218, 480, 512, 539f, 548f, 560t, 830 Pacific madrone, 237, 539, 540, 559, 829 Maeotias marginata, 371 mafic rocks, 49 magnolia, 138 Magnolia, 138 Magnuson-Stevens Fishery Conservation and Management Act (MSA), 789 magpie, yellow-billed, 200, 201, 514 Mahall, B. E., 518 mahimahi, 795 mahogany. See mountain mahogany Majidae, 318 Major, Jack, 192 Malacostraca, 723f Malacothamnus fasciculatus, 432f Malacothamnus mendocinensis, 189t Malacothamnus parishii, 189t Malakoff Diggins, 732f malaria, 880 mallards, 400, 680, 685 Malosma laurina, 432f, 433t, 483, 484t, 492t maltese star-thistle, 433t Malthus, Thomas Robert, 76 Malthusianism, 77 Malva parviflora, 771 Malvaceae, 189t Malvastrum rotundifolium, 646 mammal species in alpine ecosystems, 627t endemic to California, 204, 206t diversity hotspots of, 204, 206t mammalian wildlife, diversity hotspots of terrestrial, 203, 204, 205f mammals in California. See also marine mammals biogeography, 203–4 in deserts, 657 endangered/threatened species that require or use estuarine and tidal marsh habitats, 379t fossil history of, 156–64 large, 487t marine mammals, 792 small, 487t, 657 timeline of global events relevant for history of, 156, 157f

wetlands and, 680–81 mammoths, 157f, 161, 163, 173, 178, 455 Mammuthus, 369 Mammuthus sp., 204 Mammuthus exilis, 768 manacled sculpin, 317f management reference points, 782 management training, 877 Manayunkia speciosa, 729 Manila clam, 371 mantids, 650f manzanitas, 55–56, 139, 192, 220, 479, 480, 483, 484t, 486, 495 bigberry manzanita, 482f, 490 pinemat manzanita, 594 Santa Cruz manzanita, 829–30 maples, 139, 216, 237 bigleaf maple, 237, 512, 541, 559, 721 MAPSS-CENTURY 1 model (MC1), 259 Marah macrocarpus, 433t marbled godwit, 399f marbled murrelet, 202, 540f, 541 Margaritifera falcata, 725 Margolin, M., 177 mariculture, 798 marijuana, 718t, 875 Marina del Rey, 83 marine fisheries. See fisheries marine habitats, 367 marine hotspots, 296 Marine Invasive Species Act. See California Marine Invasive Species Act marine latitudinal shifts, climate change and, 254–55 Marine Life Management Act of 1999 (MLMA), 789 Marine Life Protection Act of 1999 (MLPA), 84, 789, 800, 929 marine mammals, fisheries and historical landings of, 791–92 marine protected areas (MPAs), 83, 300, 350, 380, 760, 800–802, 933 beaches, 402 establishment of, 771 goal of MPA network, 771 interactions and trade-offs with fisheries, 800–801 kelp forests and, 328–29 map of statewide network of, 328f marine regions, birds in, 202 marine stratus clouds, 12 marine terraces, 339 Mariposa daisy, 189t mariposa lily, single-flowered, 189t market squid, 769, 784, 789–90, 796 marlin, striped, 292, 293 Marmota flaviventris, 594, 595t, 627 marmots, 596 giant marmot, 163 yellow-bellied marmot, 594, 627, 630 Maron, J. L., 413 Marsh, George Perkins, 76 marsh bird races, tidal, 375 marsh birds, secretive, 680 marsh sandwort, 188 marsh wren, 680 marshes of Central Valley, 672f, 673 marsupials, 157f, 158, 160, 203 martens, 204, 561 American marten, 204, 628 Humboldt marten, 541 Sierra Nevada marten, 627t Martes americana, 204, 596t, 628 Martes americana spp. humboldtensis, 541

Martes americana sierra, 627t Martes martes, 561 Martes pennanti, 37, 204, 541, 561, 596t Martin, Paul S., 177–78 Marty, J. T., 848 masking, auditory, 298 Mason, J. E., 791 Masticophis lateralis euryxanthus, 198 masting, 514 mastodons, 157f, 161, 163 matorral, 483 matterhorn topography, 140 Matthews, K. R., 697 maximum sustainable yield (MSY), 781 maximum tree diameter removed (diameter limits), 567–68 Mayacamas popcorn-flower, 190t mayflies, 194, 681, 698f, 723, 731 Mazaella, 343 Mazurek, M. J., 541 McAuliffe, J. R., 647 McBride, J. R., 416, 541 McCalla, Alex F., 882 McCarthy, H., 179 McCaskie, Guy, 199 McCauley, C. K., 645 McClatchie, S., 101 McEvoy, A. F., 791 McLachlan, A., 421 McLaughlin, B. C., 525 McPhee, John, 5 McPherson, B. A., 237 McWilliams, Carey, 78, 90 meadow wetlands, mountain, 675–77 meadowfoam, 192 meadowlark, western, 459 meadows dry, 600t montane, 848–49 subalpine, 594 wet, 56, 57, 147, 215f, 594, 600t, 601, 603, 622, 624, 625, 628, 675–77, 696 meadowy surfgrass, 343 mealybugs, 239 mechanized agriculture, 871 Medicago, 454, 458 Medicago polymorpha, 433t Medieval Climatic Anomaly (MCA), 132, 173 Mediterranean annual grassland, inland, 844. See also valley grassland and interior grassland with oak canopy, 844 Mediterranean grassland, coastal, 844–45. See also coastal prairie Mediterranean-type climate (MTC), 536 native plants and livestock production in MTC inland and coastal grasslands, 847–48 seasonality, 489 medusa head, 453f, 454, 455, 461, 466, 469, 470, 850 medusae, 490 mega-cusp embayments, 391 Megachilidae, 191 megachilids, 191 megafaunal extinctions, 177–78 megalopa/postlarvae, 396 Megaloptera, 723f Megalorchestia spp., 397f, 398 Megalorchestia californiana, 397f Megalorchestia corniculata, 397f Megaptera novaeangliae, 292f, 792 Megastrea, 319f Megastrea undosa, 318, 797

INDEX  965

Megathura crenulata, 318 Melack, John M., 696, 705 Melanerpes formicivorus, 514 Melanophila species, 37 Melilotus, 454 Meliosma, 138 Meloe franciscanus, 652 Melolonthinae, 193 melons, 272 Melospiza lincolnii, 201 Melospiza melodia, 680 Melospiza melodia graminea, 769 Melospiza melodia maxillaris, 376t Melospiza melodia pusillula, 376t Melospiza melodia samuelis, 376t Melozone crissalis, 201, 488t Melozone fuscus, 488t Membranipora, 318 Mendocino, 67 Mendocino bush-mallow, 189t Mendocino cypress, 67 menodora, spiny, 648 Meniscomys, 157f, 160 Menodora spinescens, 648 Mensing, S. A., 180 Menzies’ wallflower, 414f Mephitis sp., 658 Mephitis mephitis, 459, 596t Meral, Gerald, 915 Merced Grove, 223f merced monardella, 190t mercury, 86, 685, 702, 706, 732, 733 Merenlender, Adina M., 935 Merluccius productus, 291 meromixis, 694, 706 meroplankton, 344 Merriam, John Campbell, 155f, 549 Merychippus, 157f, 161 Mesembryanthemum crystallinum, 235, 770 mesic, 159, 215 mesic shrub, 600t mesocarnivores, 37, 627t mesocosms, 322 mesocrustaceans, 312 mesofauna, 515 mesopelagic zone, 288 mesopelagics, 289f, 291–92, 294, 295 mesopredators, 442, 658 Mesozoic ecosystems, 134–36 Mesozoic toothed birds, 199 mesozooplankton, 288, 370 mesquite, 169, 175, 175f, 642, 646 storage basket for, 175f Metacarcinus, 323. See also Cancer Metacarcinus antennarius, 323 Metacarcinus magister, 323, 371, 784 Metacarcinus productus, 323 metalmark butterfly, Lange’s, 195 metaphyton, 717 Metasequoia, 139 methane emission from wetland soils, 682, 683f methanotrophic, 683f methylotrophic bacteria, pink-pigmented facultative, 437 Metridium, 319f Metridium farcimen, 321 metropolitan planning organizations (MPOs), 919 metropolitan statistical areas (MSAs), 886 Metropolitan Water District (MWD), 730 Mexican independence, 85, 869 Mexican pinyon pine, 139 Mexican whip-poor-will, 202 Mexicans, 85, 231, 842 Mexico, 203, 758t, 767

966  INDEX

Meyer, M. D., 571 mice, 160, 162, 206t, 457f, 458, 480, 487t, 514, 657, 759t. See also deer mice; harvest mice kangaroo mice, 162, 163 Perdido Key beach mouse, 207 Micrathene whitneyi, 202 microarthropods, 458, 643, 651, 658 microbes and grassland distribution, 459–60 microbial loop, 290 microclimate, 17 microclimatic gradients, 216 microcrustaceans, 312, 717 microphyll, 225 Micropterus sp., 881 Micropterus spp., 726 Micropterus salmoides, 699 Microtonae, 595t Microtus, 162 Microtus sp., 487t Microtus californicus, 438f, 457, 487t Microtus californicus halophilus, 376t Microtus californicus paludicola, 376t Microtus californicus sanpabloensis, 376t Microtus californicus stephensi, 376t Microtus longicaudus, 595t Microtus montanus, 595t Microtus oregoni, 595t microzooplankton, 289–90 midden, shell, 767f midges, 681, 698f, 717, 722, 723f, 731 phantom midge, 697 midtrophic organisms, 295 migration timing, 255–56 migratory eared grebe, 701 military uses of Channel Islands, 767–68 milkvetch, Borrego, 652 Millennium Ecosystem Assessment (MA), 269, 658 Miller, Char, 107 Miller, G. R., 524 Miller, P. C., 435 Miller, Paul, 108 millipedes, 651, 658 mills, 86, 88 mimicry, 652 Mimomys, 162 Mimulus, 191 Mimulus aurantiacus, 485t Mimulus brandegeei, 189t Mimulus guttatus, 416 Mimulus traskiae, 190t Mimulus whipplei, 190t mineral nutrition, 496–98 mineralization, 413 miners, leaf, 486 miner’s lettuce, 542 minimum residual canopy cover, 568 mining, 86. See also under California Gold Rush mink, 724, 730, 733 Minnich, R. A., 430 minnows, 674, 725, 726, 727f Miocene, mammals in, 160–62 Miotylopus, 157f, 160 Miriti, M. N., 647 Mirounga angustirostris, 292f, 295, 414, 767, 792 Mission Blue butterfly, 195 mistletoes, 114, 514 desert mistletoe, 656 dwarf mistletoe, 599 mites, 238, 515, 644, 651, 658, 698f giant red velvet mites, 195 rhagidiid mite, 621 water mites, 681, 717

mitigation banks, 908 mitigation easements, 856 mitigation offsets, 277–78 mixed conifer forest, 221–22, 557, 558f, 565f. See also specific topics generalized successional pathways for historic, 561, 562f management regime and tree density, 826f old-growth, 223f mixed-evergreen forest, 217f, 218 mixed-stock fisheries, 786 mixotrophic plankton, 344 mock heather, 416, 420 Modoc Plateau geology and geomorphology, 60 upland soils, 60 valley soils, 60 Moerisia sp., 371 Mohave ground squirrel, 206t Mojave Desert. See also deserts biogeography, 636 climate, 638 geology, 640 geomorphic and soil features in, 61 geomorphic surfaces in, 641f map of, 638f mountain-piedmont-basin floor interactions in, 61 range production, 840t subregions, 638f Mojave Desert ecosystem, conceptual model of, 639f Mojave Desert grasslands, 848 Mojave National Preserve, 84 Mojave rattlesnake, 653 Mojave sage, 146, 148 Mola mola, 290 mole salamanders, 727f moles, 457f, 458, 764 Molinari, N., 420 Mollisols, 50, 52, 55–61, 65, 66 Mollusca, 343, 723f molluscs/mollusks, 235f, 369, 371, 394, 717, 722, 791f, 796–97 Molossidae, 657 Monadenia setosa, 828 monarch butterfly, 195 monardella, merced, 190t Monardella leucocephala, 190t Monardella pringlei, 190t Monarthrum dentigerum, 237 Monarthrum scutellare, 237 monkey-flowers, 189t, 190t, 191 common monkey-flower, 416 monkey puzzle tree, 134 Mono Lake, 194, 672, 673 ecological services provided by, 705 mixing and plankton dynamics in a lowdiversity ecosystem, 698–702 Mono Lake diving beetle, 189t, 194 monocarpic plants, 416 Monocorophium insidiosum, 328 monomictic periods, 694 monsoon, southwest, 12 montane ecology, 221. See also mountains; specific topics montane forest types, 555, 557, 558f, 559–60 distribution, 556f, 557, 559f giant sequoia, 560 ponderosa, Jeffrey, and “eastside” pine, 560–61 white fir, 557, 559 montane forests, 553, 570 characteristics of mature trees for major species at turn of 19th century, 554t

ecological tolerances of common tree species in, 557, 560t ecosystem characteristics drought, pests, and pathogens, 563–64 fire, 564, 565 forest turnover, 566 topography’s influence, 564–65 wind, 565–66 ecosystem services, 566 carbon storage, 566–67 human impacts, 567 water, 566 fauna, 561–63 forest structure and function, 561 future scenarios drought and bark beetles, 569, 570f fire, 569 species distribution, 569–70 management strategies fuels treatment, 567–68 increasing forest heterogeneity and resilience, 568–69 physiographic setting climate, 554–55 soils, 555 restoration successes, 569 sensitive species that affect land management in, 561, 563f montane hardwood-conifer, 559–60 montane meadows, 848–49. See also meadows montane woodlands, 512t Monterey Bay National Marine Sanctuary (NMS), 301t Monterey cypress, 215 Monterey pine, 187f, 215 Montia perfoliata, 542 Montiaceae, 485t moon snails, 323 Mooney, Harold A., 413, 851 moorhen, common, 680 moraines, 140, 619 moray eel, California, 323 Mordecai, E. A., 436–37 Moreno-Mateos, D., 687 Mormon tea, 147, 646, 649f morning glory, beach, 414f morning sun star, 323 morphodynamic scale, 391 Morrison, S. A., 439 mosquitoes, 681, 698f, 717 mosquitofish, western, 726 mosses, 343, 644 aquatic, 716 Oregon eurhynchium moss, 542 moth larva, 515, 719 moths, 541, 651 California oak moth, 515 yucca moths, 194–95 mouflon sheep, European, 756, 760t, 767 Mougeotia spp., 720 Mount Lyell shrew, 206t, 627t, 628 Mount Pinos blue grouse, 201 mountain ashes, 139 California mountain ash, 594 mountain bluebird, 596 mountain garter snake, 697, 698f mountain hemlock, 132, 145, 147, 148, 215f, 221, 597, 598, 604, 605 mountain hemlock forests, 222, 591–92 mountain lion, 203, 487t, 527, 626, 627t, 658, 852, 937 mountain mahogany, 139, 147, 483, 484t, 593–94, 642 curl-leaf mountain mahogany, 580, 593–94

mountain meadow wetlands, 675–77 mountain pine beetle, 564, 569, 599 mountain plover, 201 mountain quail, 628 mountain soils, 63, 65–66, 555. See also Klamath Mountain soils mountain sucker, 725 mountain whitefish, 725 mountain yellow-legged frog, 199, 238, 680, 697, 698f mountainous areas, altitudinal shifts in, 142–44 changes in community composition, 144–45 mountains. See also montane ecology land use, 86–88 mourning dove, 488t mouse. See mice Moyle, Peter B., 694 mud turtles, 727f mudsnails, 371 New Zealand mud snail, 722 Muhlenbergia richardsonis, 625 Muir, John, 5, 561, 603, 613, 672, 842, 848 mule, 759t mule deer, 225, 420, 487t, 626, 627t, 630, 657, 724, 760t, 763, 764, 773 Multiple Species Habitat Conservation Planning (MSHCP), 918 multiple-use, 900 Multiple Use Sustained Yield Act (MUSYA), 900 multiplier effects, 273 multituberculates, 157f, 158, 160 multivoltine taxa, 722 muricids, 321 murrelets marbled murrelet, 202, 540f, 541 Scripps’s murrelet, 201–3, 769–71 murres, 293 common murre, 200, 294, 792 Mus musculus, 457 mussels, 322, 337, 338, 343–47, 349–52, 359, 371, 374, 376, 717, 722, 798 California mussel, 342, 348f, 377 California sea mussel, 179 western pearl shell mussel, 725 mustards, 433t, 461 black mustard, 40, 433t, 457 Sahara mustard, 659 short-podded mustard, 433t Mustela erminea, 596t Mustela frenata, 514, 596t, 627t Mustelidae, 596t mustelids, 204 mutilid wasps, 650f Mya arenaria, 371 mycorrhizae, 65, 117, 121t, 418, 421, 480, 483–84, 496, 497, 540–41, 645. See also arbuscular mycorrhizae mycorrhizal innoculum, 440 mycorrhizal mutualisms, 495 mycorrhizal networks, 542 mycota, 483 Mycteria americana, 202 myctophids, 799 Myliobatis californica, 322, 796 Mylopharadon conocephalus, 726 Myotis, 657 Myotis lucifugus, 595t, 724 Myotis yumanensis, 724 Myriophyllum spicatum, 682, 699 mysids, 289, 290, 322, 371, 377, 395, 397, 399 Mysis relicta, 699 Mytilus, 346–48 Mytilus californianus, 179, 342, 348f

Mytilus galloprovincialis, 371 Mytilus trossulus, 371 myxozoan parasites, 729 Nannipus, 162 Napa starthistle, 460, 461 narrow-faced kangaroo rat, 206t Nashville warbler, 256t Nassella. See Stipa National Forests, 821, 830 National Marine Fisheries Service (NMFS), 735–37, 789 National Marine Sanctuaries (NMS), 300, 301 National Marine Sanctuaries Program (NMSP), 300 National Oceanic and Atmospheric Administration (NOAA), 760 National Park Service (NPS), 549, 758 national parks, 84, 87, 549–50. See also specific parks National Reclamation Act, 876 National Redwood National Park Act, 550 National Vegetation Classification (NVC), U.S., 215f, 217, 218, 220–22 Native Americans, 173–74, 177–78. See also indigenous California fire regimes and, 33 natric horizons, 64, 67 natural capital, 269 Natural Communities Conservation Planning (NCCP) process, 911t, 916, 918–19 Natural Communities Conservation Plans (NCCP), 443, 904, 919 natural enemies, 234t Natural Resource Conservation Service (NRCS), 686 Natural Resources Agency. See California Natural Resources Agency Naturalist Program. See California Naturalist Program naturalized species, 230, 231, 850 Navarretia, 192 Navy, U.S., 758 neap high tides, 392 Neary, D. G., 489 necromass, 122 necrovores, 651 nectarines, 875 Neduba extincta, 189t, 195 Needle Dieback, 108 Nematoda, 193 nematodes, 458, 460, 515, 644, 658 Neocalanus, 290 Neoclinus uninotatus, 317f neoendemics, 156, 192, 764 Neogene, 136 recognizable taxa with nonanalog associations, 136–39 Neognath, 200 Neohipparion, 157f, 161 Neomysis mercedis, 371 Neopluvial, 148 Neotamias alpinus, 594 Neotamias speciosus, 596 Neotoma, 162 Neotoma sp., 487t Neotoma spp., 764 Neotoma cinerea, 133f, 595t, 596, 627 Neotoma fuscipes, 487t, 514, 543, 829 Neotoma lepida, 133f, 487t, 654 Neotrypaea californiensis, 371 Nephtys californicus, 398f, 399 Nephtys spp., 399 Nereocystis, 313f, 317, 319f

INDEX  967

Nereocystis luetkeana, 312f, 314t, 316, 398 net ecosystem production (NEP), 697 net primary production (NPP), 456f, 461–63 net-spinning caddisfly, 723f network effect, 800 Nevada National Security Site (NNSS), 648, 649f, 653, 656 Neviusia cliftonii, 557 New Chicago Marsh, 933f New Zealand mud snail, 722 newts, 723, 726 Ng, C., 721–22 Ngo, N. S., 888 Nicolaides, B. M., 916 Nicotiana attenuata, 193 night lizard, island, 199, 765 nimravids, 160, 161 nitrates, 367, 547, 600, 682, 696, 879 nitric acid vapor (HNO3), 108, 109t distribution in southern Sierra Nevada, 115, 116 nitrification, 682 nitrification-denitrification, 374 nitrogen (N), 682, 696, 729f cropland N sources and fates, 879 on dune soils, 413 groundwater contamination by, 879 growing seasons and, 463, 467, 469, 696 photosynthesis and leaf N, 518 in subalpine watershed, 599–600 total nitrogen (TN), 728, 729f nitrogen (N) air pollution and atmospheric deposition, 114–16. See also nitrogen (N) deposition critical loads of anthropogenic N deposition, 117, 121t, 440, 441f ecological effects, 116–17 exceedance values for critical loads of N deposition, 117, 118f forms and spatial trends, 114 nitrogen (N) cycle, 645 nitrogen (N) cycling in estuaries, 374, 375f in grasslands, 463–64 nitrogen (N) deposition, 109–10, 440–41, 462. See also nitrogen (N) air pollution and atmospheric deposition critical loads (CL), 117, 118f, 122f, 440, 441f nitrogen dioxide (NO2), 120, 122t nitrogen (N) fixation, 694 nitrogen-fixing bacteria, 66, 497, 555 nitrogen oxides (NOx), 108, 109t nitrogenous air pollutants in southern California, 116f nival regions, 614 noble fir, 222 noise impacts on marine organisms, 297–98 Nolina bigelovii, 146 Nollan v. California Coastal Commission, 900–901 nonequilibrium ecosystems, 836–39 nongovernmental organizations (NGOs), 526, 915 in San Diego, timeline of, 907f nordihydroguaiaretic acid (NDGA), 652, 655 Nordstrom, K. F., 421 Normalized Difference Vegetation Index (NDVI), 435 Norrisia, 319f Norrisia norrisii, 318 North American beaver, 603, 727 North American Land Mammal Ages (NALMA), 161, 369 North American river otter, 728 North Pacific albacore, 292

968  INDEX

North Pacific Gyre Oscillation (NPGO), 99, 295, 296, 365 North Pacific High (NPH), 10, 12, 13f, 14, 17, 20, 22, 95, 96, 99 North Palisade Glacier, 623f north-south distinctions in dynamics, threats, and perspectives on ecology, 4 northern anchovy, 788, 790f, 793 northern cardinal, 202 northern coastal grassland. See Coastal Prairie northern coastal scrub, 434 northern elephant seal, 292f, 295, 401, 414, 767, 792 northern flying squirrel, 561, 563f northern fur seal, 204, 401 northern goshawk, 561, 563f northern harrier, 459 northern kelp crab, 318 northern pike, 241 northern pintail, 371, 680 northern raccoon, 770. See also raccoons northern rough-winged swallow, 256t northern sea palm, 314–17 northern shoveler, 371, 680 northern spotted owl, 541, 561, 828 Northwest California: A Natural History (Sawyer), 535 Northwest Forest Plan, 89 northwest forests, California. See also Pacific coastal forest age and dimensions of conifers of, 536, 537t land use, 88–89 northwesterly winds, 12 Norway rat, 759t Notholithocarpus densiflorus, 218, 480, 512t, 515, 536, 559 Nothrotheriops, 162 Nothrotheriops shastensis, 654 Notiosorex crawfordi, 239, 657 novel ecosystems, 885, 932 novel weapons, 234t Nowacki, G. J., 179–80 Nucifraga columbiana, 589 nuclear power plant, 327 nudibranchs, 344 dorid nudibranchs, 321, 349 Numenius americanus, 201, 372 Nuphar polysepala, 678f nurse plants, 647 nutcracker, Clark’s, 589–91, 596, 597f, 605, 628 nuthatches, 563 nutmeg, 139 nutrient-use efficiency (NUE), 434 nutrients, 694. See also specific nutrients Nuttalina, 343 Nuttall’s acmispon, 414 Nuttall’s woodpecker, 514, 516 Nyctinomops femorosaccus, 657 Nymphalidae, 415 oak ambrosia beetle, 237 oak apple, 515 oak bark beetles, 237 oak diseases, 515 oak grasslands, 50, 51f oak moth, California, 515 oak parasites, epiphytic, 514 oak physiology, 518 oak savanna, 85, 201, 234, 236, 450, 451f, 480, 509, 510, 513, 516, 520f, 521–23, 536, 553, 718t, 836, 844, 849, 851–53, 875 blue oak savanna, 849, 851

oak species, 142 oak titmouse, 514 oak woodland climate, 510, 513f, 514f oak woodland types, widespread foothill structural characteristics of, 516–17 oak woodlands, 220, 509, 528, 840, 844. See also California oak woodlands; specific topics areal extent, 510, 512t biodiversity and characteristic species, 510, 512 foundation species, keystones, and ecosystem engineers, 515–16 invertebrates, 514, 515 plants, 512–13 vertebrates, 514–15 defined, 510 distribution, 509–10, 511f disturbance and ecosystem resilience, 521 cattle grazing, 523 fire, 521 grazing and woodland understory vegetation, 523 livestock grazing, 521–23 tree cutting for firewood and range management, 521–22 ecosystem services, 524 carbon storage, hydrology, and climate regulation, 524 cultural services, 524–25 ecosystem structure and processes, 516 canopy architecture and phenology, 518 carbon, water, and energy exchange, 519, 520f stand structure, 516–17 synthesis, 519, 521 land use conversion and ecosystem fragmentation agriculture conversion, 523 urban and residential development, 523–24 in mid-20th century, scenarios of, 525 climate change, 525–26 land use change, 526 management and adaptation strategies, 526–27 research needs and priorities, 527–28 threat of development to, 527f oaks, 513–15. See also black oak; blue oak; California oaks; live oaks; tanbark oak; tanoak; specific topics Engelmann oak, 510, 511f, 512t, 932 as foundation species, 515–16 island oak, 765f leather oak, 55 Oregon white oak/Oregon oak, 510, 511f, 512t, 559 relictual island oak, 764 scrub oak, 220, 484t shrub oak, 483, 484t, 488, 495 sudden oak death (SOD), 86, 233, 236–37, 515, 542, 559, 829, 849 valley oak, 214f, 510, 511f, 512, 514, 515, 517, 518, 521, 522, 525, 932 Oaks Arm (Clear Lake), 702 oakworm, California, 515 oats, 139, 771 obligate resprouters, 35, 36, 437, 487–89, 495 obligate seeders, 35–36, 488 ocean acidification, 100, 299, 380 ocean current. See California Current System ocean management, dynamic, 300 ocean shrimp, 784, 790f, 791f

Oceanic and Fisheries Investigations. See California Cooperative Oceanic and Fisheries Investigations oceanic response, 96–99 Oceanodroma furcata, 200 Oceanodroma homochroa, 769 Oceanodroma leucorhoa, 202, 294 Oceanodroma melania, 202 Ochotona princeps, 164, 207, 257, 594, 595t, 621, 626, 627t Ochotonidae, 595t ochre sea star, 344, 347f, 352 ocotillo, 148 Octopus, 344, 350 octopuses, 320, 322, 323, 343, 350 odentocetes, 294 Odocoileus sp., 203 Odocoileus hemionus, 420, 487t, 514, 516, 596t, 626, 627t, 657, 763, 852 Odocoileus hemionus columbianus, 417 Odonata, 723f Odonates, 194 odontocetes, 294 Oenothera avita eurekensis, 193 Office of Planning and Research (OPR), 914, 919 offshore California Current conservation issues and management structure, 299–300 ecosystem services, 297 ecosystem threats, 297–99 integrating science and management, 300–302 offshore ecosystems (and food webs of California Current System), 287–88, 302 management of California offshore ecosystem, 300 mesopelagics, 291–92 primary consumers, 289–91 primary producers, 288–89 services provided by offshore ecosystems, 297 small pelagics, 291 spatial distributions, 296–97 top predators, 292–95 trophic interactions and ecosystem functioning, 295–96 offshore region of California Current System, defined, 287 Ogle, K., 639–40 Ohlone tiger beetle, 195, 196f oil exploration in Channel Islands, 767–68 oil production, 297 oil spills, 298–99, 349 Oithonia davisae, 371 old-growth chaparral, 494–96 old-growth forests, 219f, 223, 592f old-growth redwood forests, 219f, 538f, 541– 44, 547–49, 547t Oligocene, mammals in, 160 Oligochaeta, 723f Oligochaeta, 698f oligochaete worms, 681 oligomictic periods, 694 oligotrophic species, 116 oligotrophic status, 694 olive rockfish, 321, 327f olive-sided flycatcher, 256t Olivella, 766 olives, 89, 875 Olmstead, A. L., 870 Olmsted, Frederick Law, Jr., 901 Olneya tesota, 636 Olympic Coast National Marine Sanctuary (NMS), 301t

omnivores, 161, 316, 318, 459, 514, 648, 653, 656, 657, 681, 722 “omnivorous” fisheries, effects and tradeoffs of, 786–87 habitat disturbance, 787 omnivorous predators, 658 Onagraceae, 485t Oncorhynchus clarki, 829 Oncorhynchus clarki henshaw, 725 Oncorhynchus clarki seleniris, 237, 725 Oncorhynchus clarkii clarkii, 728 Oncorhynchus clarkii henshawi, 699 Oncorhynchus keta, 728 Oncorhynchus kisutch, 291f, 292, 379t, 728, 786, 829 Oncorhynchus mykiss, 237, 379t, 680, 699, 725, 726, 728, 735, 792, 829 Oncorhynchus mykiss aguabonita, 237, 725 Oncorhynchus mykiss newberrii, 725, 728 Oncorhynchus mykiss subspp., 725 Oncorhynchus nerka, 699 Oncorhynchus tshawytscha, 292, 379t, 680, 728, 786, 829 Ondatra, 162 one-seeded juniper, 636 onespot fringehead, 317f onion, wild, 192 ontogenetic niche shifts, 725 Onychomys torridus, 657 opah, 786 opalescent inshore squid. See market squid opaleye, 318, 795 open space, 900, 901, 912–15, 935 oak woodland, 525 recreation and, 275 timeline of legislation and regulation affecting, 909–11t Open Space, 901–3, 905f, 906–8 Open Space acquisition and establishment, timeline of, 903f, 904f open space area in urban counties, 908t open space plans sustainable societies strategies, conservation, and, 919–20 Ophiodon, 319f Ophiodon elongatus, 323, 786 Ophioroidea, 320 Ophiothrix, 321 opossum, Virginia, 203, 231, 442 Opuntia, 433, 648 Opuntia spp., 768 Opuntia basilaris, 655 Opuntia littoralis, 432f, 433 orange-crowned warbler, 256t, 765 orange puffball sponge, 321 oranges, 873f orchard grass, 461 Orcinus orca, 203, 292f ordinations, 415 Oregon ash, 512, 721 Oregon eurhynchium moss, 542 Oregon oak. See Oregon white oak/Oregon oak Oregon vesper sparrow, 200 Oregon white oak/Oregon oak, 510, 511f, 512t, 559 Oremland, R. S., 704 Oreochromis niloticus, 704 oreodonts, 157f, 160 Oreortyx pictus, 628 Oreothlypis celata sordida, 765 Organic Acts, 900 organic agricultural production, structure of California’s, 877–78 Orme, A. R., 393 ornate shrew, 376t, 680–81

Orobanchaceae, 189t orogenic processes, 339 orographic effects, 583 orographic uplift, 18 Orthocarpus, 454 orthopterans, 651 Oryctolagus cuniculus, 763 Osborn, Henry Fairfield, 549 Osmeridae, 291, 727f ostracods, 717 Ostrea conchaphila, 796 Ostrea edulis, 798 Ostrea lurida, 371. See also Ostrea conchaphila ostriches, 840 Otospermophilus beecheyi, 400, 514, 595t. See also Spermophilus beecheyi otters, 311, 730, 733, 794. See also river otters; sea otters overexploitation, 178–81. See also overfishing; overharvesting overfishing, 769, 782–84, 786, 792, 794, 795, 800–804. See also fisheries overharvesting, 207, 350, 351. See also overexploitation overkill hypothesis, 178. See also overexploitation overstory, 217, 220, 547, 559 overyielding, 267 Ovis canadensis, 594, 596t Ovis canadensis californiana, 204 Ovis canadensis cremnobates, 204 Ovis canadensis nelsoni, 626, 627t, 652 Ovis canadensis sierrae, 626, 627t ovoviviparous reproduction, 702 Owens Lake, 704 Owens pupfish, 188 owl limpet, 343, 350 owls, 37, 201, 459. See also spotted owls barred owl, 200, 828–29 burrowing owl, 443, 459, 465, 655, 847, 851 elf owls, 202 great gray owl, 199, 201 great horned owl, 459, 488t, 514 Oxalis, 542f Oxalis oregana, 539, 546f oxisols, 482f oxygen minimum zone (OMZ), 291, 293 shoaling, 100 Oxyjulis, 319f Oxyjulis californica, 321 Oxylebius, 319f Oxylebius pictus, 322 Oxyura jamaicensis, 371 oyster drill, 371 oystercatchers. See black oystercatchers oysters, 173, 273, 371, 374, 796, 798 native oyster, 796 ozone (O3), 109t, 120, 122f, 602, 890 biological effects, 113–14 distribution in space and time, 112–13 Pachygrapsus, 343, 350 Pachythyone, 321 Pacifastacus fortis, 681, 724 Pacifastacus leniusculus, 724 Pacifastacus nigrescens, 189t Pacific banana slug, 541, 543 Pacific barracuda, 795 Pacific bluefin tuna, 292, 795 Pacific bonito, 795 Pacific coastal forest. See also northwest forests general vegetation extent and history of, 536–37

INDEX  969

Pacific Decadal Oscillation (PDO), 98–99, 132, 253, 295, 296, 365, 394, 735 Pacific electric ray, 323 Pacific euchalon, 379t Pacific fisher, 37, 541. See also fisher Pacific Fishery Management Council (PFMC), 789, 799 Pacific Gas and Electric Company, 704–5 Pacific herring, 373, 790f Pacific High. See North Pacific High Pacific kangaroo rat, 206t Pacific lamprey, 724, 725, 728 Pacific mackerel, 790f, 793 Pacific madrone, 237, 539, 540, 559, 829 Pacific Northwest subalpine tree species, 593 Pacific oyster, 374, 798 Pacific pond turtle, 680 Pacific rattlesnake, 486 Pacific razor clam, 396 Pacific rhododendron, 829 Pacific salmon, 792 Pacific sardine, 292, 784, 788, 790f, 793 Pacific saury, 292 Pacific shrew, 680–81 Pacific slope flycatcher, 256t Pacific staghorn sculpin, 371 Pacific tree frog, 438f, 459, 721 packrats/pack rats, 543. See also woodrats bushy-tailed packrat, 133f, 596 desert packrat, 133f Padgett, P. E., 435 Pagurus, 343 Paine, R. T., 268, 347–48 painted greenling, 322 Paiute cutthroat trout, 237 Paiute sculpin, 725 Paleocene, mammals in, 156–59 Paleocene-Eocene Thermal Maximum (PETM), 137, 159 paleoendemics, 192, 765 Paleogene, 136 recognizable taxa with nonanalog associations, 136–39 paleontological localities, map of, 159f Paleoparadoxia, 157f, 163 palm trees, 159 palms, 138, 159, 225 California fan palm, 654 Guadalupe Island palm, 193 paloverde, 636, 647, 656 palps, 397 palustrine wetlands, 678 palynological records, 258 Pandalus jordani, 784 Pandalus platyceros, 790 pannes, salt, 360 Panopeus generosa, 371 Panthera onca, 204 Panularis, 319f Panulirus interruptus, 321, 323, 326, 371, 787 Papaveraceae, 485t parabolic dunes, 412 Paralabrax, 319f Paralabrax clathratus, 322, 795 Paralabrax maculatofasciatus, 371, 795 Paralabrax nebulifer, 795 Paralichthys californicus, 371, 372f, 769, 790 parasites epiphytic oak parasites, 514 myxozoan parasites, 729 parasitic arthropods, 194 parasitic wasps, 196 parasitoids, 486, 650f Parastichopus californicus, 797 Parastichopus parvimensis, 319f, 797 Paratettix aztecus, 721

970  INDEX

Paratettix mexicanus, 721 parent material, 640 Parish’s bush-mallow, 189t Parish’s gooseberry, 190t Parker, A. J., 180 Parkinsonia microphylla, 636 Paromomys, 157f, 158 Parophrys vetulus, 371, 372f Parrish, O., 180 particulate matter (PM), 107, 108, 109f, 122f particulate organic carbon (POC), 729f partridge, 177 Parus rufescens, 559 Parvicapsula minibicornis, 729 Pasadena freshwater shrimp, 189t Passerculus rostratus, 400. See also Passerculus sandwichensis Passerculus sandwichensis, 459 Passerculus sandwichensis beldingi, 376t, 379t, 400 Passerculus spp., 400 Passerina amoena, 438f, 488t passerines, 199, 400 passive samplers, 112 Patagioenas fasciata, 514 Pataki, Diane E., 887–90 paternoster ponds, 622 Patiria, 319f Patiria miniata, 320 Patterson, S. M., 179 payments for ecosystem services (PES), 273, 826, 848, 855–56 pea clams, 698f peaches, 875 peanut worms, 343 pear cacti, prickly. See prickly pear cacti pearl shell mussel, western, 725 pears, 875 Pearse, J. S., 400 peas, 416 chaparral pea, 483, 484t silky beach pea, 414f peat formation, 675–76 peccaries, 157f, 162, 163 Pectinophora gossypiella, 242 pedalogic clock, 49 pediments, 61 pedogenic silica, 49 Peinado, M., 410 pelagic algae, 698f pelagic ecosystem, wind-forced upwelling and, 96 pelagic fishes coastal, 792–93 small, 290 pelagic invertebrates, 370, 371 pelagics, small, 291 Pelagophycus, 313f, 319f Pelagophycus porra, 312f, 314t Pelecanus erythrorhynchos, 680, 704 Pelecanus occidentalis, 200, 372, 704, 788 Pelecanus occidentalis californicus, 1–2f, 772 pelicans, 401. See also brown pelicans American white pelican, 680, 704 pellagic zone, 344 Pelobatidae, 727f Peninsular Ranges geology and geomorphology, 64–65 lowland soils, 66 mountain soils, 65–66 Penitella, 343 Pennisetum setaceum, 659 Penstemon heterodoxus, 625 People for Open Space, 913 peopling of California, 171–72

migrations, population trends, and regional specializations, 173–77 paleoenvironmental trends, 172–73 peppergrass, 458 pepperweed, perennial, 682 Peprilus simillimus, 291 Peracarida, 723f perches. See also surfperches kelp perch, 317f Sacramento perch, 671 Perdido Key beach mouse, 207 peregrine falcon, 400, 628, 772 perennial grasses, 220 perennial pepperweed, 682 perennial seaweed, 349 Peri, D. W., 179 peridotite, 56, 57 periglaciation, 621 periphyton, 703, 728–29, 732 Perisoreus canadensis, 200 Perissodactyla, 203 perissodactyls, 157f, 160 periwinkles, 343 permafrost, 621 Perminalia, 138 Perognathus alticolus, 206t Perognathus californicus, 487t Perognathus inornatus, 206t Peromyscus, 162, 164 Peromyscus sp., 487t Peromyscus boylii, 487t Peromyscus californicus, 438f, 480, 487t Peromyscus eremicus, 657 Peromyscus maniculatis, 543 Peromyscus maniculatus, 417, 419, 457, 487t, 595t, 627–28, 765 Peromyscus maniculatus anacapae, 240, 770 Peromyscus polionotus trissyllepsis, 207 Peromyscus spp., 514 Persea, 138 Persicaria amphibia, 679 persistent seed banks, 486 pest control, pollination and, 271–72 pesticides, 110, 733. See also organic agricultural production Peterson, C. D., 412 petrale sole, 790f, 802 petrels. See storm petrels petrocalcic horizons, 62f, 63, 64 Petrochelidon spp., 400 Petromyzontidae, 727f Petrophila spp., 719 Petrosaurus, 198 pH-dependent charge, 50 pH of ocean surface, 351. See also ocean acidification Phacelia insularis, 766 Phaethon aethereus, 202 phainopepla, 514 Phainopepla nitens, 514, 656 Phalacrocorax penicillatus, 323, 371 Phalaria rotundata, 398 Phalaris aquatic, 461 Phalaropus fulicaria, 294 Phalaropus lobatus, 701 Phalaropus tricolor, 701 phalecia, 765 Phaleria rotundata, 397f phanerophytes, 624 phantom midge, 697 Phenacomys intermedius, 595t phenology, 255, 316, 450, 716, 850 Pheres blue butterfly, 195 Pheucticus melanocephalus, 559 Phoca vitulina, 323 Phoebastria immutabilis, 294

Phoebastria nigripes, 292f, 294 phoebes, 400 pholad bivalves, 343 Pholididae, 322 Phoradendron spp., 514, 656 phosphate fixation, 50 phosphorus, 497, 729f soluble reactive, 700 total phosphorus (TP), 695, 728, 729f photic zone, 699. See also euphotic zone photochemical smog, 108, 110f photodegradation, 463 Photopsis, 650f photosynthesis, 490 photosynthetic capacity, 518 photosynthetic efficiency, 519 photosynthetic sulfur bacteria, 703 Phragmatopoma, 343 Phragmatopoma californica, 321 Phryganidia californica, 515 Phrymaceae, 189t, 190t, 486t Phrynosoma, 198 Phrynosoma blainvillii, 459 Phrynosoma coronatum, 239 Phyllaplysia taylori, 372 Phyllobaenus, 650f Phyllodoce breweri, 594 Phyllospadix, 343, 344 Phyllospadix torreyi, 344f Phyllostomidae, 657 phylogenetic diversity, 266 Phyrmaceae, 190t, 486t Phyrnosomatidae, 198 Physeter macrocephalus, 294, 792 physical dormancy, 489 physiological dormancy, 489 Phytocrene, 138 phytophagous insects, 194 Phytophthora infestans, 236 Phytophthora ramorum, 86, 233, 236, 237, 515, 542, 559, 829, 849 phytoplankton, 96, 360, 374f, 694, 696, 700– 705, 717, 756 primary productivity by, 698–99, 700f phytoplankton groups, 288 phytotoxic indices, 113 Pica nuttalli, 200, 514, 516 Picea, 138, 141 Picea breweriana, 139, 557 Picea engelmannii, 222, 537t, 557, 593 Picea sitchensis, 218, 536, 537t Pickart, A. J., 415 Pickeringia montana, 483, 484t pickleweed, 360, 361f, 373, 641, 678f, 679 Picocystis sp. strain ML, 701 Picoides arcticus, 562, 596 Picoides nuttallii, 514 pied-billed grebe, 680 piedmont, 61 piedmont soils, 63–64 pigeon, band-tailed, 514 pigeon guillemot, 294 pigs, 205, 229, 238, 456, 516, 759t, 768, 841 feral pigs, 400, 516, 759t, 763, 773 wild pig, 241, 514 pikas, 161, 163, 164, 628, 630 American pika, 207, 594, 596, 621, 626–28 western pika, 257 pikeminnow, 721, 726 pikes, 188, 240 northern pike, 241 Pillsbury, N. H., 522 Pinaceae, 134, 495 Pincetl, Stephanie S., 886 Pinchot, Gifford, 549–50 pincushion plant species, 679

pine beetles, 564 mountain pine beetle, 564, 569, 599 Pine Creek Canyon glacier, 141f pine grosbeak, 201, 596 pine plantation forest, 568f pine siskin, 628 “pineapple express.” See atmospheric rivers pinemat manzanita, 594 pines, 193, 483, 599. See also bristlecone pine; foxtail pine; limber pine; lodgepole pine; ponderosa pine; white pine; whitebark pine; specific topics bishop pine, 67, 200, 215 Bolander’s beach pine/Bolander pine, 67, 214, 536, 592 California closed-cone pines, 142, 145 Coulter pine, 56, 218, 222, 483, 559 eastside pine, 558f, 560–61 foothill pine, 510, 512, 517f, 517t, 849 gray pines, 56, 218, 220 Jeffrey pine, 36, 56, 109t, 113, 114, 117, 216f, 221, 224, 536, 554t, 557, 559t, 560–61, 565f, 566, 589, 592 knobcone pine, 220, 483, 536 Monterey pine, 187f, 215 sugar pine, 145, 146, 221, 522, 536, 554t, 557, 559, 560t, 565f, 567, 599 Torrey pine, 215, 483 yellow pine, 32, 35, 557, 560–61, 567 Pinicola enucleator, 201, 596 pink abalone, 320 pink bollworm, 242 pink-pigmented facultative methylotrophic (PPFM) bacteria, 437 pink sand verbena, 416 pinnipeds, 295, 299, 389, 394, 401, 756, 770, 792 pintail, northern, 371, 680 pinto abalone, 320 Pinus, 560 Pinus, 29, 134, 483, 484t Pinus spp., 36, 642, 768 Pinus albicaulis, 141f, 146, 222, 253, 580, 580f Pinus attenuata, 142, 484t, 536 Pinus balfouriana, 146, 222, 557, 580 Pinus cembroides, 139 Pinus contorta, 147, 222, 253, 580, 696 Pinus contorta murrayana, 222, 557 Pinus contorta subsp. bolanderi, 67, 214, 536, 592 Pinus contorta subsp. contorta, 592 Pinus contorta subsp. latifolia, 592 Pinus contorta subsp. murrayana, 592 Pinus coulteri, 56, 218, 222, 559 Pinus flexilis, 147, 224, 580, 599 Pinus jeffreyi, 36, 56, 108, 221, 536, 557, 817 Pinus lambertiana, 145, 148, 221, 522, 536, 537t, 557, 817 Pinus longaeva, 147, 224 Pinus macrocarpa, 218 Pinus monophylla, 141, 142f, 224, 636 Pinus monticola, 146, 222, 253, 536, 537t, 557, 580, 589, 696 Pinus muricata, 67, 142, 200, 215, 484t Pinus ponderosa, 29, 133f, 145, 220, 254, 513f, 522, 536, 537t, 554, 817 Pinus radiata, 142, 215, 435, 484t Pinus radiata var. binata, 193 Pinus sabiniana, 56, 218, 484t, 510, 849 Pinus torreyana, 215 pinyon, 36, 146, 147, 642, 647 pinyon chaparral, 202 pinyon jay, 656 pinyon-juniper, 28t, 145

pinyon-juniper woodlands, 117, 123, 145– 48, 214f, 215f, 224, 254, 625, 647, 656 pinyon pines, 141, 142f, 146, 147, 224, 596, 661 Mexican pinyon pine, 139 single-leaf pinyon pine, 141, 636 pinyon trees, 648 pipefishes, 322 Pipilo maculatus, 488t, 773 pipits American pipit, 201, 400 water pipit, 628 Piranga flava, 202 Piranga ludoviciana, 559 Piranga rubra, 202 Pisaster, 347, 348, 350 Pisaster spp., 344 Pisaster ochraceus, 268, 344, 347, 348f piscivorous birds, 371 Pisidium spp., 698f Pismo clam, 396–97, 399 pistachios, 867 Pitcher, Tony J., 798 Pitkin Marsh paintbrush, 189t Pituophis catenifer, 514 Pituophis melanoleucus, 459 Pityostrobus, 134 Plagiobothrys glaber, 190t Plagiobothrys lithocaryus, 190t Plagiobothrys spp., 458 Planariae, 311 plane-tree, 134 plankters (planktonic organisms), 96 planktivores, 291, 293, 317, 319f, 320–21, 323, 324, 329 planktivorous fishes, 320, 321, 377, 788 plankton, 396, 705, 787, 798 plankton dynamics, 694 in a low-diversity ecosystem, 698–702 planktonic foraminifera, 254–55 planktonic phase, 290 Planktothrix, 720 Planning and Conservation League (PCL), 915 Planning and Research, Office of. See California State Office of Planning and Research plant area index (PAI), 518 plant community, defined, 213 plant diversity. See also flora: biodiversity patterns of, 192 Plantae, 343 Plantaginaceae, 485t Plantago erecta, 443 Plantago maritima, 679 plantains, 679f California plantain, 465 seaside plantain, 679 plantanoid, 134 Platanaceae, 138f Platanus, 139 Platanus racemosa, 512 Platycarya, 138 Platygonus, 162 platyhelminthes, 235f Platyhelminthes, 343, 723f playa soils, 61–63 playas, 61, 704 wet and dry, 641 Pleamis platurus, 198 Plecoptera, 194, 723 Plegadis chihi, 201 Pleistocene late Pleistocene, 145–46 mammals in, 162–63 terminal Pleistocene archaeological sites and isolated finds, 172f

INDEX  971

Pleistocene Ice Ages, sculpting during climate belt shifts, 762 continuous dry land vs. submergence, 762 sea level fluctuations and coastlines, 762 Pleistocene paleontological sites, 158f Plenthodontidae, 197 Plestiodon spp., 514 Plethodon asupak, 198 Plethodon stormi, 198 Plethodontid, 197 Pleuraphis rigida, 647 Pleurophycus, 313f, 319f Pleurophycus gardneri, 314, 314t, 315f, 317 Pliocene, mammals in, 162 Pliohippus, 157f Pliohippus, 161, 162 Pliopotamomys, 162 Plithocyon, 161, 369 plovers. See also snowy plovers black-bellied plover, 398, 399f mountain plover, 201 plume worms, 343 plums, 875 pluvial lake deposition, 642 Pluvialis squatarola, 398, 399f pneumatocysts, 312 Poa, 626 Poa douglasii, 414f Poa secunda, 513 Poaceae, 145, 625, 626 poachers, 322 pocket gopher, 56, 160, 162, 164, 457f, 516, 851 pocket mice, 206t, 487t pocketed free-tailed bat, 657 Podiceps nigricollis, 701 Podilymbus podiceps, 680 Poecile atricapillus, 200 Poecile gambeli, 559 Pogonichthys ciscoides, 189t Pogonichthys macrolepidotus, 371, 674, 739 Point Conception, 14 Point Reyes bird’s beak, 679f Point Reyes paintbrush, 189t poison oak, 513 Polemoniaceae, 485t Polemonium eximium, 141f police power, 900, 908 policy across California’s ecosystems. See also environmental law future directions, 936–37 policy goals and legislative tools, mismatches between, 917–18 policy needs, perspectives on, 936 Polinices. See Euspira Polioptila californica, 37, 201, 202, 434, 488t Polioptila californica californica, 202 pollen, 141, 144f Pollicipes, 343 pollination, 269, 466–67, 654 defined, 269 and pest control, 271–72 pollinators, 109t, 191, 195, 196, 203, 252, 261, 270f, 277, 418, 449, 470, 648, 651, 656, 657. See also pollination pollock, 293 pollution. See also air pollution; water pollution from shipping and cruise liners, 298–99 from urban activities, 890–91 wetland management and, 685 Polyartha spp., 698f polybrominated diphenyl ethers (PBDEs), 377 Polycentropis spp., 698f

972  INDEX

polychaete worm, 320, 321, 328, 394, 395, 398, 399, 729 polychaetes, 312, 318, 371, 397 colonial, 321 polyclimax, 846 Polygonaceae, 485t, 486t, 625 Polynesian rat, 229 Polyplacophora, 345 polyploids of creosote bush, distribution of, 655f polyploidy, 416 Polypodium californicum, 546f Polypodium scouleri, 539 Polystichum, 542f Polystichum munitum, 218, 542, 546f Pomeroy, C. A., 889–90 pomfrets, 795 Pomoxis spp., 726 pond lily, yellow, 678f pond turtles Pacific pond turtle, 680 western pond turtle, 727 ponderosa pine, 29, 36, 40, 109t, 113, 114, 117, 133f, 145, 146, 216f, 220, 221, 254, 536, 537, 554, 555, 557, 559–61, 566, 830 ponderosa pine forests, fires in, 32 Ponderosae, 560 Pooecetes gramineus, 459 Pooecetes gramineus affinins, 200 Poole, D. K., 435 popcorn flowers, 190t, 458 poppies California poppy, 416, 433t tree poppy, 139 population, human, 75, 78, 90–91 approaches to population-environment relationships, 77–78 and land use, 90–91 land use and, 74–78, 90–91 population density, 81 population geography, 78 population pyramids, 82 in Mission and Rancho eras (1769-1848), 78–79 unusual things about California’s population history, 78 early statehood era (1849-1889), 79 Progressive era (1890-1920), 79–80 interwar era through World War II (19201950), 80 postwar era (1950-1990), 80–81 current era (1990-present), 81–82 population contractions and expansions, plants, 142, 144 population growth, human and agriculture, 872–73 of California’s largest cities, 78, 79t “natural,” 76 population growth rates, 76, 77f wetlands and, 686 Populus, 139 Populus fremontii, 512, 647, 678f, 679, 935 Populus tremuloides, 580 Populus trichocarpa, 593 porcelain sand crab, 399 porcupine, 162 Porifera, 343 Porinchu, R. F., 697 porpoises, 157f, 163, 311 Port Orford cedar, 536, 537t portfolio effect, 267 Portolá, Gaspar de, 170 Portunus xantusii xantusii, 399 Postelsia, 350 Postelsia palmaeformis, 342, 344

postfire annuals, 483, 485t, 487. See also pyro-endemics postfire plant regeneration, 486. See also resprouting postfire predators, 486, 488t Potamocorbula amurensis. See Corbula (Potamocorbula) amurensis Potamopyrgus antipodarum, 722 potatoes, 542 potential biological removal (PBR), 301 potential evapotranspiration (PET), 10, 11f, 214, 510, 638, 887 Potentilla, 626 Potentilla multijuga, 190t Potentilla uliginosa, 190t pots. See traps and pots Powell, J. A., 415 Prabhu, S., 886 prawns, 797 spot prawn, 790, 790f Preble’s shrew, 203 precipitation, 9, 10f, 11f, 16f, 17. See also climate(s) on coastal dunes, 410 interannual variability, 19–20 latitude and, 18–19, 739 spatial variability of, 18–19, 739 predaceous diving beetle, 681, 698f predaceous ground beetle, 399 predaceous insects, 650f predacious beetles, 196 predator removal, 788 predators, 459 apex, 323 in California Current, 292–95 in deserts, 658 keystone, 516 marine, 347 mesopredators in deserts, 658 omnivorous, 658 postfire, 486, 488t pressure gradient, 22 pricklebacks, 322 prickly pear cacti, 646, 648 beaver-tail prickly pear, 655 coastal prickly pear, 433t primary consumers, 719 primary producers, 288–89, 317–18, 343, 368–69, 372, 716–17, 719 primary production aboveground net, 456f net, 456f, 461–63 primary productivity, 698–99, 700f annual, 698 interannual fluctuations in, 702–3 primates, 156, 157f, 158–60 primrose. See evening primrose primrose-willow, Uruguayan, 682 Pringle’s monardella, 190t Prionace glauca, 292f, 293, 769, 795 proboscideans, 161, 162 Procambarus clarkii, 726 Procyon lotor, 400, 459, 487t, 596t, 770 Procyonidae, 596t Prodipodomys, 162 Prodoxidae, 195 production curve, 781 Progne subis, 202 Progressive era (1890-1920), 79–80 Progressive Era legacy, 908, 911 prokaryotes, 288, 369, 370, 681, 700 prokaryotic cyanobacteria, 717 pronghorn, 161, 225, 455, 658, 852. See also Antilocapra americana American pronghorn, 657 pronghorn antelope, 522

propagule dispersal, 324, 603, 629 propagule pressure, 234t, 235 propagule transport, 422f propagules, 3, 35, 344, 348, 436, 763, 934 property rights, 912 taking of private, 900–901 Prosopis spp., 642 Prosopium williamsoni, 725 protected area challenges, 935–36 Protected Area Database. See California Protected Area Database protected area status of coastal sage scrub, 429, 431f protected areas, 84–85. See also marine protected areas protozoa, 460, 644, 658 protozoans, 235f, 254, 651, 681 Prunus, 139, 483 Prunus sp., 484t Prunus emarginata, 222, 482f, 484t, 594 Prunus ilicifolia, 484t Prunus subcordata, 484t Psaltriparus minimus, 438f, 488t Pseudacris spp., 514 Pseudacris regilla, 438f, 459, 721 Pseudityophthorus pubipennis, 237 Pseudo-nitzschia multiseries, 288 Pseudochironomus richardsoni, 719 Pseudococcus mauritimus, 239 Pseudococcus viburni, 239 Pseudodiaptomus forbesi, 371 Pseudodiaptomus marinus, 371 Pseudoregnia spicata, 225 Pseudotsuga, 542f Pseudotsuga macrocarpa, 35, 222, 483, 484t, 559 Pseudotsuga menziesii, 145, 218, 221, 237, 536, 537t, 546f, 555, 817 Psorothamnus, 147 Psuty, N. P., 422 Psychoglypha spp., 698f psychological ecosystem services, 271 pteromalids, 650f Pterygophora, 313f, 319f Pterygophora californica, 314, 315f Ptychocheilus grandis, 721, 726 Ptychoramphus aleuticus, 293, 770 public participation, ecological monitoring and, 938 priorities for advancing, 938 publicly owned treatment works (POTWs), 350 Puccinea jaceae var. solstitialis, 233 Puffinus griseus, 294 Pugettia, 319f Pugettia producta, 318 Puma concolor, 203, 487t, 514, 516, 595t, 626, 627t, 852 pump-and-fertilize approach, 879 pumpkin, 272 pupfishes, 727f desert pupfish, 656, 704 Owens pupfish, 188 purple martin, 202 purple needlegrass, 220, 450, 457, 460, 838, 839f purple sage, 433t purple sea urchin, 318, 321, 769, 797 purple veldtgrass, 417, 419f, 420 purse seines, 785 Purshia, 225 Purshia tridentata, 145, 647 pycnocline, 98 Pycnopodia, 319f, 321 Pycnopodia helianthoides, 321, 322 pygmy cypress, 536

pygmy forests, 67 pygmy mammoth, 204, 768 pygmy rockfish, 788 pyrethroids, 733 pyro-endemics, 487. See also postfire annuals Pyrocephalus rubinus, 202 pyroclastic flows, 55 quagga mussel, 722 quails California quail, 465, 488t mountain quail, 628 quaking aspen, 580, 593, 594f, 605 quaking aspen stands, 594f quantitative trait loci, 416 Quaternary, 49, 140. See also Holocene; Pleistocene Quaternary alluvial fan deposits, 49 Quaternary environments, 139–42 Quercus, 144f, 515 Quercus sp., 484t Quercus spp., 220, 768 Quercus agrifolia, 220, 236, 483, 484t, 510, 512t, 513f, 518f, 541, 849 Quercus berberidifolia, 484t Quercus chrysolepis, 218, 220, 510, 512t, 513f, 559 Quercus douglasii, 220, 432f, 509, 512t, 513f, 849, 932 Quercus dumosa, 55 Quercus durata, 484t Quercus engelmannii, 510, 512t, 513f, 932 Quercus garryana, 510, 512t, 513f, 559 Quercus kelloggii, 221, 237, 510, 512t, 513f, 557 Quercus lobata, 510, 512t, 513f, 518f, 932 Quercus parvula var. shrevei, 515 Quercus tomentella, 764, 765f Quercus vaccinifolia, 482f, 484t, 598 Quercus wislizeni, 559 Quercus wislizeni var. frutescens, 484t Quercus wislizeni var. wislizeni, 510, 512t, 513f quillback rockfish, 323, 327f Quimby Act of 1965, 910t, 913 Quino checkerspot butterfly, 439 Rabb, L.M., 177 rabbitbrush, 141, 147, 647 rabbits, 157f, 160, 161, 163, 417, 420, 455, 457, 594, 764 brush rabbit, 487t, 514, 543 European rabbit, 759t, 763, 769 raccoons, 160, 344, 400, 459, 487t, 724, 770 radiation, infrared, 20 radiative forcings, 118 railroad, logging, 821f Railroad Flat, 175f rails, 201, 372. See also black rails; clapper rails rainbow trout, 237, 241, 699, 721, 726, 735 Rallus longirostris, 674 Rallus longirostris levipes, 376t, 379t Rallus obsoletus levipes, 201 Rallus obsoletus obsoletus, 201, 372, 376t, 379t Ralph, F. M., 14 Ramalina menziesii, 509f, 513 Rana, 199 Rana boylii, 726, 727 Rana cascadae, 199 Rana catesbeiana, 680 Rana draytonii, 199, 238, 726 Rana muscosa, 199, 238, 628, 680, 698f Rana sierra, 199 ranching. See livestock ranching Rancholabrean Land Mammal Age, 163 range, 836, 857. See also rangelands

defined, 836 range community dynamics, models for, 845 range condition model, 846 range ecology, 843, 845–46 range ecosystem models, 843 range ecosystems in California, 835–36, 857 equilibrium and nonequilibrium models of community dynamics in, 837 forage in, 843–45 characteristics, 836 range livestock production in Mediterranean climate zone, 845 range management in California, 845 adaptive, 845–46, 851 conceptual tools for, 845–46 future of, 854–57 history, 841–43 incremental, 851 native plants, livestock production, and implications for, 850–51 range ownership, patterns of, 840 range production management systems, land tenure, and other influences on, 840t range production system model, 843 range retrogression, 837, 846 range riparian zones, 849–50 Rangeland Conservation Coalition. See California Rangeland Conservation Coalition rangeland ecosystems. See range ecosystems rangeland improvement, 33 rangeland resolution, 855 rangeland vegetation types, 836 rangelands, 835, 836, 857, 928, 928f definition and terminology, 835, 836, 845 rapid increase in C4 ecosystems event (RICE), 161 raptors, 201, 400, 442, 459, 514, 851 Rare Bird Committee. See California Rare Bird Committee rat tail fescue, 433t rationalization, 803 rats, 205, 344. See also packrats/pack rats black rat, 240, 759t, 772 cotton rat, 162 Norway rat, 759t Polynesian rat, 229 rattlesnakes, 198, 653, 851 Pacific rattlesnake, 486 sidewinder rattlesnake, 653, 655 western rattlesnake, 459 Rattus exulans, 229 Rattus rattus, 240, 769 Raven, P. J., 191 ravens, 400, 654 common raven, 488t, 650, 655, 656, 659 rayless brittlebrush, 142 rays, 359, 371, 795 bat ray, 322, 796 Pacific electric ray, 323 razor clam, Pacific, 396 reballasting, 299 Reclamation, Bureau of. See United States Bureau of Reclamation recreation, 275, 767–68, 935 dunes impacted by, 418 land use for, 84 recurrence intervals, 719 Recurvirostra americana, 685 red abalone, 320, 350, 786, 790 red alder, 541, 678f, 679 red algae, 315, 317, 318, 319f, 346 red-backed vole, southern, 561

INDEX  973

red bat, western, 724 red-bellied newt, 726 red-billed tropicbird, 202 red brome, 433t red cedar, western, 146, 536, 537t red crossbill, 596 red diamond rattlesnake, 653 red-eared slider, 727 red fir forest, 221, 222, 223f red firs, 28t, 40, 49, 57, 145–47, 215f, 216f, 221, 553–55, 557, 559f, 560f, 564, 566, 580, 589, 591, 592 Shasta red fir, 537 red foxes, 205 Sierra Nevada red fox, 204, 627t, 628 red heather, 594 red huckleberry, 540f red Irish lord, 323 red-legged frogs, 854 California red-legged frog, 199, 238, 726, 848 red-necked phalarope, 701 red oaks, 514, 515 red pharalope, 294, 296 red polychaete, 398f red sand verbena, 418 red scale, California, 239 red sea cucumber, giant, 797 red sea urchin, 318, 321, 322, 769, 797 red-shouldered hawk, 514 red swamp crayfish, Louisiana, 722, 726 red-tailed hawk, 459, 488t, 514, 628, 724 red tree vole, California, 206t red turf, 344 red turpentine beetle, 599 red urchin, 790f red velvet mites, giant, 195 redband trout, 728 redbud, 139 redflower currant, 193 redhead, 202 redox gradients, 700 redshank, 479, 479–80f, 483 redside dace, 725 redstem filaree, 433t, 452, 845 redtail surfperch, 399 redwood burl, 548f redwood exploitation, history of, 549 redwood forests, 537, 538f, 539, 550–51, 822. See also coast redwood forests aquatic systems and land-water connection, 547 climate, 543–44, 545f distribution of co-occurring tree species, 539 ecosystem services, 547 energy and carbon balance, 543–44 forest dynamics, 542–43 fire frequency and response, 543 regeneration, 543 forest structure, 541–42 future of, 550 nutrient dynamics, 545, 547 old-growth, 219f, 538f, 541–44, 547–49, 547t organisms below the ground, 540–41 organisms up in the canopy, 539–40 plants, 539 vegetation, 217–18 wildlife, 541 redwood giant trees, 543 Redwood National Park, 549, 550 redwood sorrel, 539, 542 redwood stump, resprouting, 548f redwood tree. See also redwoods

974  INDEX

ecosystem high above the ground in canopy of a mature, 539, 540f regeneration of new root system by, 547, 548f redwood violet, 542 redwoods, 139, 216, 218, 219f, 226, 236. See also specific topics distribution, 537, 538f, 539 early conservation and the fight to save the, 549–50 giant redwood, 192 Reed, Daniel C., 316 Reed, S. E., 935 reef fish, 327 reefs bivalve, 367 shallow rocky. See kelp forests reflective beaches, 391 refugia, 142, 144, 261 Regional Advance Mitigation Planning (RAMP), 919 Regional Blueprint, 919 Regional Greenprints, 919, 920 regional specialization, 171–72, 175 Regulus satrapa, 559 Reisner, Mark, 881 Reithrodontomys sp., 487t Reithrodontomys megalotis, 457, 487t, 595t, 765 Reithrodontomys megalotis distichlis, 376t Reithrodontomys megalotis limicola, 376t Reithrodontomys raviventris, 205, 206t, 379t, 680 Reithrodontomys raviventris halicoetes, 376t Reithrodontomys raviventris raviventris, 376t relicts, 192, 765 Rena humilis, 198 replacement level (fertility), 82 reproductive isolation, 375 reptiles, 196, 674, 725f, 727f, 764, 765. See also specific topics biogeography, 197–98 conservation context, 198–99 in deserts, 653, 656 evolutionary diversification, 196–97 riverine food webs and, 726–27 research, interdisciplinary and applied, 937–38 research needs that emphasize climate change as an emerging driver, 937 reservoirs, 704–5 residential development, 523–24 residual dry matter (RDM), 456, 462, 847, 852 residuum, 56 resilience (of ecosystem functions), 267, 470, 755, 774 resource depression reactionary archaeology, tempered overkill, and, 178–79 resprouting, 35, 543, 559, 660, 774 of chaparral, 486–89, 490f, 491, 494, 495, 499 of coast redwoods, 540, 542, 543, 548f of coastal sage scrub, 433t, 434, 437 epicormic, 35, 36 of oak woodlands, 521, 522 of saltcedar, 660 restart, American, 200 restrictive horizons, 52 Reuter, J. E., 699 Reynolds, J. F., 639–40 Rhagidia gelida, 621 rhagidiid mite, 621 Rhamnus, 220. See also Frangula sp.

Rhamnus californica, 433t, 513 Rhamphocottus richardsonii, 322 Rhaphiomidas terminatus abdominalis, 66 Rhaphiomidas terminatus terminatus, 415 Rhinichthys osculus robustus, 725 Rhinichthys osculus subspp., 725 rhinoceros auklet, 294 rhinoceroses, 157f, 159–62 rhizomes, 35, 414 Rhode, P. W., 870–71 Rhododendron, 540, 542 Rhododendron sp., 236 Rhododendron columbianum, 594 Rhododendron macrophyllum, 829 Rhododendron occidentale, 541 rhododendrons, 236 Pacific rhododendron, 829 Rhopalodia, 719–20 Rhopalodiacea, 719 Rhus, 139, 220, 483 Rhus sp., 484t Rhus integrifolia, 433t, 484t Rhus ovata, 484t, 490f Rhyacophila amabilis, 189t Rhyacophilidae, 189t Rhyacotritonidae, 727f rhyolite, 55 Ribes, 225, 513 Ribes divaricatum, 190t Ribes sanguineum, 193 Riccia sp., 848 rice, 871, 874–76, 880–83 RICE (rapid increase in C4 ecosystems event), 161 rice fields, 881 as wetlands, 684–85, 687 ricegrass, Indian, 648, 649f Richardson, D. M., 230 Richardsonius egregius, 725 Ricketts, Ed, 338 ridges (dunes), 412 Ridgway’s rails, 376t, 379t California Ridgway’s rail, 201 light-footed Ridgway’s rail, 201 Yuma Ridgway’s rail, 202 Rim Fire burn scar, 731f ring-billed gull, 399f rip currents, 391 Riparia riparia, 724 riparian habitats, 234 riparian wetlands, 674–75 riparian zones, range, 849–50 ripgut brome, 420, 455, 458, 468, 770–71 river basins, river flow and water right allocations in, 912 river bulrush, 684 river flow allocations, 912 river mouth estuaries, 367–68 river networks and ecosystems, general tendencies in, 714 river otters, 204, 681 North American river otter, 728 river systems, characteristics of various, 718t riverine food webs in Eel River, 719–22 key taxa in, 713, 714, 725f, 727f. See also under Eel River amphibians and reptiles, 726–27 fish, 724–26 invertebrates, 717, 722–24 mammals, 727–28 primary producers, 716–17 vertebrates, 724 spatial energy sources to channel them down the drainage network, 720–22

riverine invertebrate groups, 723f. See also under riverine food webs riverine wetlands, 670t, 672–75 rivers, California, 713, 715f, 740 hydrologic regimes in, 714, 716 past and alternative futures of Californians and, 729–30, 732–34, 738–40 runoff and, 713, 714, 730–33, 736, 738 Riverside, 115 Riverside County, western, 443, 444, 523, 872 Riverside fairy shrimp, 681 Riversidean sage scrub, 430, 432f, 438, 440–43 roadrunner, 488t, 656 Roberts, D. M., 699 Robinia, 139 rock broken rock habitats, 621 types of, 56–57 rock crabs, 323, 371 rock dove, 488t rock glaciers, 614, 621 rockfishes, 292, 294, 295, 318, 321, 323, 326, 327f, 769, 786–88, 790, 791f, 794, 795, 804 brown rockfish, 327f, 373 rockweeds, 343, 346, 347f, 349, 350, 352 Rocky Mountain lodgepole pine, 592 Rodalia cardinalis, 242 rodent species diversity in deserts, precipitation and, 652f Rodentia, 203, 205f, 206t, 595t rodents, 157f, 160, 162, 164, 204, 455, 514, 594, 628, 651, 652f, 654, 656, 657 horned rodents, 159 scatter-hoarding rodents, 480, 486, 488 Roosevelt elk, 188, 414, 673, 674f root profiles of native shrubs and exotic annual grasses, 435, 436f root rot, 564 Rosaceae, 145, 190t, 625, 626 rose, 134, 145, 146 roses, 484, 593 rosid, 134 rosy boa, 198 rot fungi, canker, 515 rotifers, 377, 696, 698f rough-winged swallow, northern, 256t round-headed Chinese houses, 414 round worms, 717 rove beetles, 399 rubber boa, 198 Rubus, 542f Rubus parviflorus, 542 ruddy duck, 371 ruffed grouse, 200 rufous-crowned sparrow, 439, 488t rugosa rose, 421 Rumex, 458 Rundel, Philip W., 624 runnel, 392 runoff, 566 agricultural, 733 beaches and, 394 chaparral and, 496–98 climate change and, 380, 622 from estuaries, 363–67, 370, 374, 377, 378, 381 floodwater, 687 lakes and, 693, 696, 698, 702, 703, 705, 706 pollutant/toxic, 374, 671, 675, 730, 891 rainfall, 14, 19

rivers and, 713, 714, 730–33, 736, 738 from snowmelt, 19–20, 599, 622, 684, 686, 693, 696. See also snowmelt stormwater, 61, 64, 350, 671, 698, 891 subsurface, 116, 122 surface, 64, 116, 122, 349, 464–65, 496, 498, 510, 852 terrestrial, 349 timing and seasonality, 363, 364, 366, 686, 713 urban, 374, 377, 378, 381, 736 water balance and, 464–65 wetlands and, 671, 673, 675, 684, 686, 687 runoff ratio, 363–64 Ruppia spp., 369 rural vs. urban areas, percentage of population living in, 76 rush-rose, island, 772 rust, 233 ryegrass, Italian, 458, 466 ryes, wild, 225, 450 Rykaczewski, R. R., 96, 99–100 saber-toothed cats, 162, 163, 204 false saber-toothed cats, 157f, 161 non-felid saber-toothed cats, 157f, 163 sablefish, 787, 790f, 794 Sacramento Flood Control Project, 671 Sacramento National Wildlife Refuge (NWR) Complex, 201, 677f Sacramento perch, 671 Sacramento pikeminnow, 721, 726 Sacramento River, seasonal flows of, 717f Sacramento River Delta, 875 Sacramento-San Joaquin system. See also San Joaquin River environmental contaminants through a longitudinal continuum, 731–33 Sacramento splittail, 674 Sacramento Valley agriculture in, 875 birds in, 201 wetlands, 671, 673 Sacrobatus, 145 sage, 433t. See also sagebrush big sage, 647, 648 Great Basin sage, 636, 642 Mojave sage, 146, 148 sage grouse, 37, 225, 628 sage scrub, 480. See also coastal sage scrub; specific topics Diegan sage scrub, 432f, 434, 438, 441f, 443 Riversidean sage scrub, 430, 432f, 438, 440–43 Venturan sage scrub, 430 sage sparrow, 201, 488t sagebrush, 36, 37, 139, 144f, 145–48, 213, 224, 224f, 225, 591, 642, 647, 675 California sage/California sagebrush, 433t, 437f, 439, 764 giant sagebrush, 141, 145 silver sagebrush, 849 sagebrush scrub, 194. See also sage scrub sagebrush steppe, 224–25 sagewort, coastal, 416 saguaro, 636, 647 Sahara mustard, 659 Sala, O. E., 259, 640 salal, 829 salamanders, 196–98, 539, 540f, 541, 723, 726, 727f. See also slender salamanders; tiger salamanders Salamandridae, 197 salema, 321

Salicornia virginica, 641. See also Sarcocornia pacifica saline-sodic soils, 54 saline soils, 54 salinity, 376, 694 Salix, 139, 536 Salix sp., 678f Salix spp., 594, 848 Salix hookeriana, 418 Salix laevigata, 721 Salix lasiolepis Benth., 721 Salmo trutta, 699, 726 salmon shark, 293 salmon spawning grounds, loss of, 880, 881f salmon weir, hupa, 175, 176f Salmonella, 853 Salmonidae, 727f salmonids, 699, 717, 725, 727f, 729 salmons, 83, 169, 175, 176f, 177, 181, 273, 292, 293, 295, 302, 366, 373, 376, 381, 680, 699, 717, 723, 724, 729, 739, 787, 789, 791f, 792–93, 794f, 796, 798, 800, 881, 882 Chinook salmon, 100, 175, 292, 379t, 680, 724, 728, 739, 786, 790f, 792, 798, 829 chum salmon, 728 coho salmon, 175, 292, 379t, 724, 728, 792, 798, 829 karuk dip net for, 176f Salsola, 648 salt cedar. See saltcedar salt marsh dodder, 679 salt marsh harvest mouse, 205, 206t, 376t, 379t, 680–81 saltbushes, 225, 647, 648, 657 saltcedar, 238, 642, 660, 682 salticids, 650f saltmarsh common yellowthroat, 201 Salton Sea, 202, 703–4 management challenges, 704 Salvelinus fontinalis, 696, 726 Salvelinus namaycush, 699 Salvia, 429 Salvia apiana, 432f, 433, 439 Salvia columbariae, 430 Salvia leucophylla, 433t, 435, 439 Salvia mellifera, 432f, 433t, 435, 485t Salvia mohavensis, 146 Salvia spp., 480 Sampson, Arthur, 85 San Bruno elfin butterfly, 195 San Clemente Dam, 737, 739 San Clemente Island, 758t–760t, 763–67, 772, 773 San Clemente Island loggerhead shrike. See island loggerhead shrike San Diego, nongovernmental organizations (NGOs) in, 907f San Diego fairy shrimp, 681 San Francisco, 2 special districts in, 907f San Francisco Bay, introduction of invasive species into, 235 San Francisco Bay Area, 913. See also South San Francisco Bay Area San Francisco garter snake, 198 San Francisco waterfront in 1851, 80f San Gabriel Mountains, 219f San Joaquin antelope squirrel, 206t, 847 San Joaquin Basin, 671–72 San Joaquin Experimental Range, 517f San Joaquin kit fox, 465 San Joaquin pocket mouse, 206t

INDEX  975

San Joaquin River, 671–72. See also Sacramento-San Joaquin system San Joaquin Valley agriculture in, 875–76, 878–79 birds in, 201 San Nicolas Island, 758t–760t, 763–66, 768 San Nicolas Island desert-thorn, 189t San Onofre Nuclear Generating Station, 327 sand bass, barred, 795 sand bass, spotted, 371, 795 sand-castle worms, 343 sand crabs, 396, 399, 400 sand-loving wallflower, 414 sand verbenas pink sand verbena, 416 red sand verbena, 418 yellow sand verbena, 409–10f, 414f sandalwood, 134 sanddab, speckled, 373 sanderling, 389–90, 394f, 399f, 400 sandhill cranes, 201, 680 sandpipers, 202 least sandpiper, 202, 372, 373f sandworts marsh sandwort, 188 sea sandwort, 417 Santa Ana winds, 21, 31, 34, 121 Santa Barbara Island, 763, 764t, 766t, 767, 769, 773. See also California Islands; Channel Islands of California Santa Barbara oil spill (1969), 299, 349 Santa Catalina Island, 766t, 767, 773 Santa Catalina Island fox, 773 Santa Catalina Island monkey-flower, 190t Santa Clara River, seasonal flows of, 717f Santa Cruz Island, 758t–760t, 763f, 764t, 766t, 775f Santa Cruz Island buckwheat, 764 Santa Cruz Island monkey-flower, 189t Santa Cruz long-toed salamander, 198 Santa Cruz manzanita, 829–30 Santa Cruz Mountains, 548 Santa Lucia fir, 221–22 Santa Miguel Island, 758, 758t–760t, 763, 764t, 766t, 773. See also California Islands; Channel Islands of California Santa Rosa Island, 758t–760t, 763f, 764t, 766–68, 769f Santalales, 134 Santiago Canyon Fire of 1889, 39 Sants Cruz cypress, 215 sap feeder, 515 saprophagous, 398 saproxylic insects, 194 sapsuckers, 201, 239, 486 Williamson’s sapsucker, 596 sapwood, 492t, 493, 543 Sarcobatus vermiculatus, 646 Sarcocornia pacifica, 373, 678f, 679 Sarda chiliensis, 291 Sarda chiliensis lineolata, 795 sardine productivity, climate-dependent, 793 sardines, 96, 101, 293, 296, 359, 376, 790, 791f, 793, 799, 800, 804 Pacific sardine, 292, 784, 788, 790f, 793 Sardinops sagax, 291, 291f Sargassum, 317, 343 Sargassum filicinum. See Sargassum horneri Sargassum horneri, 328, 770 Sargassum muticum, 327, 328, 344 sargassum seaweed, 770 Sargent’s cypress, 483, 536 Sarnelle, O., 703 saturation point, 18 Satureja palmeri, 193

976  INDEX

Satyr butterfly, 195 satyrid butterfly, 415 Sauromalus ater, 653 saury, Pacific, 292 savanna, 29, 162, 220, 844. See also oak savanna savannah sparrow, 376t, 400, 459 Belding’s savannah sparrow, 379t, 400 Save the Redwoods League, 549, 550 Sawyer, John O., Jr., 415, 535, 555, 557, 624 Sawyer, Lorenzo, 86 Sawyer decision, 86 Saxidomus gigantea, 371 Sayornis spp., 400 Say’s phoebe, 400 scale insects California red scale, 239 cottony cushion scale, 242 scallops, 374 Scapanus spp., 457 Scaphiopus sp., 653 Scaphohippus, 161 Scarab, 193 scarab beetles, 193 Scarabaeoidea, 193 scarp, 392 scatter-hoarding birds, 488 scatter-hoarding rodents, 480, 486, 488 scavengers, 193, 200, 293, 294, 373, 398f, 400, 681, 724, 727, 788 scavenging, 114 scavenging crabs, 399 scavenging isopods, fast-swimming, 399 Sceloporus, 198 Sceloporus occidentalis, 438f, 459, 486, 514 scenario planning, 889 Schinus terebinthifolius, 886 Schismus barbatus, 659 Schoenherr, Allan, 5 Schoenoplectus acutus, 671, 678f Schoenoplectus californicus, 679 Schrepfer, Susan, 550 Schwarzenegger, Arnold, 917 Schwinning, S., 640 Sciuridae, 206t, 595t sciuromorph rodents, 157f Sciurus sp., 203 Sciurus carolinensis, 231 sclerophyll, 217, 497 sclerophyllous leaves, 217, 490, 491t evergreen, 482, 483, 490, 496t sclerophyllous shrubs, 479, 487 evergreen, 220, 479, 483 sclerophyllous species, evergreen, 483, 494 sclerophyllous trees, 518 sclerophyllous woodland, evergreen, 480 Scoliopus bigelovii, 542 Scomber japonicus, 291 Scorpaena guttata, 323 Scorpaenichthys, 319f Scorpaenichthys marmoratus, 322 scorpionfishes, 323 California scorpionfish, 327f scorpions, 195, 650f, 651, 652 Scotch broom, 452 Scott, T., 525 scree, 621 Scripps’s murrelet, 201–3, 769–71 Scrophularia, 458 Scrophulariaceae, 414 scrub-jays, 200. See also island scrub-jay western scrub-jay, 200, 488t, 514 scrub oak, 220, 484t scuds, 717, 723f sculpins, 317f, 322, 323, 337, 725, 727f Pacific staghorn sculpin, 371

Scutellospora, 484 Scythrididae, 415 sea anemones, 343, 344. See also anemones sea basses giant black sea bass, 795 giant sea bass, 323, 769, 795, 804 white seabass, 323, 790, 790f, 795 sea breeze, 21–22 sea cow, 157f sea cucumbers, 320, 321, 343, 790f, 791f, 797 warty sea cucumber, 320, 326, 797 sea hares, 343, 350 California sea hare, 318 sea lettuce, 352 sea lions, 178, 295, 376, 400 California sea lion, 295, 401, 767, 769, 791 Steller’s sea lion, 792 sea mammals, 179 sea mussel, California, 179 sea otters, 83, 204, 268, 318, 320, 323, 325, 330, 372, 376, 400, 769, 788, 792, 797 sea palm, 342, 344 sea palm, northern. See northern sea palm sea palm, southern. See southern sea palm sea sandwort, 417 sea slug, 343, 344, 372 sea snakes, 198 sea stars, 320–23, 337, 343, 344, 347, 348f, 350. See also starfish ochre sea star, 344, 347f, 352 sea surface temperature (SST), 14–16, 98, 99. See also climate change sea turtles, 294, 299, 786, 795 green sea turtle, 199 leatherback sea turtle, 293f, 294, 298 loggerhead sea turtle, 294, 300 sea urchins, 204, 268, 269, 317, 318, 320–23, 325, 343, 769, 787, 788, 791f, 797, 800. See also urchins seabirds, 100, 199–203, 240, 241, 288, 291– 96, 298–99, 302, 343, 372, 381, 396, 399, 401, 547, 680, 756, 767, 769–71, 773, 774, 784, 786–88, 792, 795 seagrasses, 268, 345, 349, 367–69, 372, 392, 395, 400 seals, 311. See also elephant seal; fur seals harbor seals, 323, 401 searocket flowers, 418 searockets American searocket, 416, 419 European searocket, 419 seaside daisy, 409–10f seaside plantain, 679 seasonal flows of rivers, 717f seasonal fluxes of streamflow and stream nitrogen load, 463, 464f seasonal patterns of soil moisture and photosynthesis for chaparral shrubs, 490 seasonal phenology of California Current System, 97–98 of Eel River biota, 719–20 seasonal variations in temperature and precipitation, 17–18 driving plant growth, 450, 452f seasonal wind flow patterns, 110, 111f seasonality. See also growing seasons; runoff of Mediterranean-type climate, 489 seaweed flies, 398 seaweeds, 338, 342–45, 347, 349, 350, 398, 769–71. See also algae; kelp perennial, 349 Sebastes, 321, 323 Sebastes spp., 769, 794 Sebastes carnatus, 319f, 323 Sebastes caurinus, 319f

Sebastes chrysomelas, 319f, 323 Sebastes entomelas, 794 Sebastes flavidus, 786 Sebastes goodie, 291f, 786 Sebastes guttata, 319f Sebastes levis, 786 Sebastes maliger, 319f, 323 Sebastes melanops, 319f, 323 Sebastes mystinus, 319f, 321, 790 Sebastes nebulosus, 319f, 323 Sebastes paucispinus, 786 Sebastes pinniger, 786 Sebastes ruberimus, 786 Sebastes serriceps, 319f, 322 Sebastes wilsoni, 788 Sebastolobus spp., 794 Secchi depth, 698, 700f secretive marsh birds, 680 sedges, 145, 179, 849 sediment budget, 391, 393 sediment transport, 360 sediments, anoxic, 682 seed banks, 483, 486 seed-dispersers, mutualist, 516 seedling production, seedling-to-parent ratio, and resprout survival, 494f segmented worms, 343 sei whale, 792 Selasphorus sasin sedentarius, 765 selectivity (grazing), 845 Semicossyphus, 319f Semicossyphus pulcher, 322, 795 semivolatile organic compounds (SOCs), 109t, 118 semivoltine taxa, 722 Senate Bill 375 (SB 375), 911t, 917–19, 921 senesce, 450 señorita, 321 sequential depletion, 786 Sequiadendron giganteum, 597 Sequoia, 134, 542f Sequoia sempervirens, 68, 192, 195, 214, 217, 237, 536, 537t, 538f, 546f, 675, 678f, 817 Sequoiadendron, 134 Sequoiadendron giganteum, 192, 222, 553 seral conditions, 561 serial depletion, 766 Seriola lalandi, 291, 795 serotiny, 483 Serpentes, 197f, 198 serpentine, 192, 216 “serpentine syndrome,” 56 serpentinite, 56 serpulid worm, 328 Serpulorbis, 343 serviceberry, Utah, 594 servicesheds, 270 Sespedectes, 157f, 160 Sespia, 157f, 160 sessile eggs, 726 sessile lifestyle, 714 sessile organisms, 337–38, 343, 344, 346, 371, 719 seston, 703 Setophaga coronata, 559 Setophaga uticilla, 200 shad, 381 American shad, 373 shadscale, 141, 145–47, 225, 641 sharks, 100, 292, 293, 322, 323, 796 blue shark, 292f, 293, 769, 795 soupfin shark, 794, 796 sharp-skinned hawk, 488t sharp-tailed grouse, 202 Shasta crayfish, 681, 724

Shasta ground sloth, 654 Shasta red fir, 537 Shasta River, 739 Shasta snow-wreath, 557 Shaw, M. Rebecca, 856 shear strength, soil, 458 shearwaters, 293 black-vented shearwater, 202 sooty shearwater, 293f, 294, 298 sheep, 85, 89, 205, 455, 456, 467, 567, 603, 672, 760t, 768, 769, 840–42, 848, 852, 853, 868, 870, 871f. See also bighorn sheep; European mouflon sheep domestic sheep, 756 sheephead, California, 322, 327f, 795 shell midden, 767f shellfish, 178, 179, 376, 378, 379, 756 Shepherd, J. G., 804 shieldback katydid, Antioch Dunes, 189t, 195 ship strikes, 298 shipping and cruise liners, 297–98 shipping lanes, modifying to avoid whale strikes, 300–302 shoaling, 100, 342 shore crabs, 343, 371 shorebirds, 201, 317–18, 343, 344, 372, 373f, 373t, 377, 381, 389, 390f, 394f, 396, 398–402, 674, 680, 685, 687, 881 short-beaked common dolphin, 292f short-billed shorebirds, 398 short-eared owl, 201, 459 short-faced bears, 157f, 163 short-podded mustard, 433t shortfin mako shark, 293, 795 shortnose sucker, 728 shovel-nosed snake, western, 653 shoveler, northern, 371, 680 shredders, 722, 723 Shreve’s oak, 515 shrews, 161, 203, 204, 239, 376t, 594, 680–81 desert shrew, 657 Mount Lyell shrew, 206t, 627t, 628 Preble’s shrew, 203 water shrews, 627t, 724 shrimp, 195, 320, 343, 371, 372, 374f, 379, 381, 399, 641, 681, 722, 797. See also mysids California freshwater shrimp, 681, 724 fairy shrimp, 195, 641, 675, 681 ocean shrimp, 784, 790f, 791f Pasadena freshwater shrimp, 189t shrink-swell processes, 56, 60, 66 shrub oak, 483, 484t, 488, 495 shrub-ox, 162 shrublands, native plants and livestock production in, 849 Shuford, W. David, 200, 202 Sialia currucoides, 596 Sialia mexicana, 514 Sialidae, 723f Sialis spp., 698f Sickman, J. O., 696–97 side-blotched lizards, 653 common side-blotched lizard, 438f sidewinder rattlesnake, 653, 655 Sierra Club, 550 Sierra foothills (and inner coast range) heavy metal pollution, 732–33 soils, 50 Sierra juniper, 222, 580, 589, 591, 592, 605 Sierra lodgepole pine, 222 Sierra Nevada. See also specific topics ammonia distribution in southern, 115, 116 birds in, 201

cirque, cliff, and valley slope environments of eastern, 620 climate, 18, 19 distribution of forest types in southern, 559f distribution of vegetation types, 215, 216f, 218 forest fires in conifer forests in, 29f geology and geomorphology, 48–49 lakes of, 696–702, 706 ecological services provided by, 705 nitric acid distribution in southern , 115, 116 sediment and nutrient inputs from logging, livestock grazing, and fire in mid- and high-elevation, 731 soil formation influences, 49 soils of granitic terrain, 49–50 soils of montane valleys, 50–51 soils of volcanic terrain, 50 vegetation shifts since the last glacial maximum in, 142, 143f Sierra Nevada bighorn sheep, 594, 596, 627t, 628, 630 Sierra Nevada marten, 627t Sierra Nevada red fox, 204, 627t, 628 Sierra Nevada yellow-legged frog, 199, 628 Sigmodon, 162 signal crayfish, 724 Siguenza, C., 418 Silene acaulis, 626 Siliqua patula, 396 silk tassel, 220, 483, 484t silky beach pea, 414f silky flycatcher, 656 Sillett, S. C., 544 silt, 640, 641 Silveira, T. M., 422 silver fir, 537t silver sagebrush, 849 silverspot butterfly, Callippe, 195 Silvetia, 343, 346 Simimys, 157f Simpson, H. J., 700 Simpson, J. J., 96–97, 101 Simuliidae, 723f single-flowered mariposa lily, 189t single-leaf pinyon pine, 141, 636 single-tree selection, 825 Siphateles bicolor obesa, 725 Siphateles bicolor subspp., 725 Sipuncula, 343 Sirenidae, 163 siskin, pine, 628 Sitka spruce, 218, 536, 537t, 541 Sitta spp., 563 skates, 795 skinks, 459 Skinner, John E., 376 skipjack tuna, 292, 795 skunks, 459, 658. See also spotted skunks skypilot, 141f slacks, 412 Slagel, M. J., 422 slender salamanders, 198 Channel Islands slender salamander, 764, 765 slope and climate, 17 slugs, 458. See also banana slugs sea slug, 343, 344, 372 small wrasses, California, 322 smartweeds, 684 water smartweed, 679 smectite clays, 52 smelts, 371, 379t, 727f, 799 delta smelt, 371, 372f, 377, 379t, 738 topsmelt, 321, 371

INDEX  977

Smeringurus mesaensis, 650f Smith, J. K., 37 Smith, Jedediah, 542 Smith, Michael, 2 smog, photochemical, 110f smooth cordgrass, 237 smooth turban snail, 318 snailfishes, 322 snails, 195, 318, 320–23, 342–44, 350, 380, 418, 458, 515, 675, 681, 717, 720, 723f, 797. See also mudsnails; turban snails Trinity bristle snail, 828 snake mackerel, 795 snakes, 162, 196, 198, 199, 442, 459, 653, 656, 721, 765. See also garter snakes; rattlesnakes water snakes, 727f snakeweed, 648 snow and growing seasons, 253, 554, 555, 583, 598, 624, 677 snow depth, 19, 20f, 583, 586f. See also precipitation snow-dominated ecosystems, 696–97 snow-wreath, Shasta, 557 snowbell, 139 snowfields, permanent, 622 snowmelt, 19, 599, 600, 622, 624, 693, 696, 705–6, 716, 718t. See also runoff runoff from, 19–20, 599, 622, 684, 686, 693, 696 snowpocket forests, 587 snowy egret, 323, 399f snowy plovers, 398, 399f, 401 western snowy plover, 201, 202, 379t, 399f, 400, 401, 414, 421 sodic soils, 54 soft brome, 433t soft chess, 452 soft-shell clam, 371 soft turfs, 343 soil biota, 436–37, 853 soil conditions, future of, 68 soil crusts, biological, 643–44 soil depth, 50 “soil engineers,” 516 soil formation, 641 soil hydrophobicity, 38 soil organic carbon (SOC), 49, 58, 59 soil organic matter, 38, 52, 55, 58–61, 463, 467 soil shear strength, 458 soil system and livestock production, 852–53 soil(s), 454. See also under Central Valley; Sierra Nevada bulk density, 456–57, 852–53 of Central Valley, 52–54 and desert ecosystem processes, 640–41 dune, 413 fire effects on, 38 grassland subtypes determined by unique, 453 hydric, 670, 674 lowland, 66 mountain, 63, 65–66, 555. See also Klamath Mountain soils piedmont, 63–64 playa, 61–63 saline-sodic, 54 valley, 50–51, 58, 60 volcanic, 49, 59–60, 617f wetland, 67–68, 682, 683f Solanaceae, 189t solar reflectance, 20 Solaster, 319f Solaster dawsoni, 323

978  INDEX

soles English sole, 371–73, 381 petrale sole, 790f, 802 Solifugae, 195 song sparrows, 375, 376t, 543, 680, 769 songbird species, spring arrival timing of, 255, 256t songbirds, 202, 255, 256t, 372, 400, 561 Sonoma chipmunk, 206t Sonoran Desert, 202, 840t. See also deserts Sonoran-Mojave creosote bush, 225 Sonoran mud turtle, 727f sooty crayfish, 189t sooty shearwater, 293f, 294, 298 Sorbus, 139 Sorbus californica, 594 Sorex lyelli, 206t, 595t, 627t, 628 Sorex monticolus, 595t, 627t Sorex ornatus, 680, 769 Sorex ornatus salaries, 376t Sorex ornatus salicornicus, 376t Sorex ornatus sinuosus, 376t Sorex pacificus, 680 Sorex palustris, 595t, 627t Sorex preblei, 203 Sorex tenullus, 627t Sorex vagrans, 627t Sorex vagrans halicoetes, 376t Soricidae, 206t, 595t Soricomorpha, 203, 205f Sork, V. L., 525 sorted circles, 621, 623f Soulé, M. D., 442 soupfin shark, 794, 796 Sousa, W. P., 346, 714 South Bay Salt Pond Restoration, 933f South Coast, 876 South Coast Air Basin, wind patterns in, 110, 111f South Fork Eel River, 719, 721 South San Francisco Bay Area, urban development in, 873 southern bull kelp, 312f southern California, geologic history of, 760–62 Southern California Bight (SCB), 100, 756, 760, 762, 795. See also Channel Islands of California southern grasshopper mouse, 657 southern red-backed vole, 561 southern sea otter, 322, 379t, 767, 769 southern sea palm, 314, 315, 317 soybean, 874, 875 spadefoot toads, 653, 656, 726, 727f Spanish-cedar, 138, 139 Spanish-Mexican period (1760-1848), agriculture during, 869 Sparks, J. P., 645 sparrows, 375, 459, 488t. See also savannah sparrow; vesper sparrow black-throated sparrow, 656–57 Lincoln’s sparrow, 201 rufous-crowned sparrow, 439, 488t sage sparrow, 201, 488t song sparrows, 375, 376t, 543, 680, 769 Spartina, 368, 377 Spartina spp., 269 Spartina alterniflora, 237, 269, 377, 674 Spartina foliosa, 237, 269, 377, 674, 678f, 679 Spartina hybrid, 380 spawning, broadcast, 397 Spea hammondii, 726 Spea intermontanus, 726 special districts, 906, 907f, 913, 916 specialized grazing systems, 845

species non-native, 229, 230. See also invasive species nonindigenous, 230. See also invasive species number of, 230 species ranges, 257–59 species richness, 187f, 188, 234t, 265. See also biodiversity specific conductance, 695 specific leaf area (SLA), 490, 518 speckled dace, 725 speckled sanddab, 373 sperm whale, 294, 792 Spermophilus, 164 Spermophilus sp., 487t Spermophilus beecheyi, 457, 487t, 851. See also Otospermophilus beecheyi Spermophilus lateralis, 487t Spermophilus mohavensis, 206t Spero, J. G., 523 Speyeria callippe callippe, 195 Sphaeriidae, 723f Sphyraena argentea, 323, 795 Sphyrapicus thyroideus, 596 spider crabs, 318 spiders, 272, 647, 650f, 651, 720, 723, 724, 731 lycosid spiders, 721 trapdoor spider, 415 Spilogale gracilis amphiala, 765 Spilogale putorius, 655 spinach, 272 spiny clotbur, 771 spiny hopsage, 147, 648, 649f spiny lobsters, 323, 326, 371, 787, 788, 790, 791f, 797 California spiny lobster, 321–22 spiny mendora, 648 spiny sand crab, 399 Spionidae, 397 Spiraea splendens, 482f Spirinchus thaleichthys, 379t Spirobranchus, 343 Spirogyra spp., 720 spits, crenulate, 390 Spizella atrogularis, 488t splittails, 371 Clear Lake splittail, 189t cyprinid splittail, 739 Sacramento splittail, 674 Spodosols, 58, 67 sponges, 320, 321, 343 sporophylls, 328 sporophyte, 315 spot prawn, 790 spotfin croaker, 399 spotted owls, 37, 829 California spotted owl, 201, 561, 563f northern spotted owl, 541, 561, 828 spotted sand bass, 371, 795 spotted skunks, 655 Channel Islands spotted skunk, 765 spotted towhee, 488t, 773 sprawl, 80, 81, 888, 903 spring tides, 395 spring transition, 97 springs, 642 springtails, 515 spruces, 138, 139, 145 Engelmann spruce, 222, 537t, 557, 593 fir-spruce, 539 Sitka spruce, 218, 536, 537t, 541 squamates, 197f squash, 855 Squatina californica, 795–96

squids, 292–94, 302, 772, 790, 791f, 799 Humboldt squid, 293, 295 jumbo squid, 798 market squid, 769, 784, 789–90, 790f, 796 squirrels, 157f, 158, 160–62, 457, 654, 724. See also ground squirrels Douglas’s squirrel, 596 eastern gray squirrel, 231 golden-mantled squirrel, 627 northern flying squirrel, 561, 563f San Joaquin antelope squirrel, 206t, 847 tree squirrels, 160, 203 St. John’s wort, 242 staghorn sculpin, Pacific, 371 Stallcup, Rich, 199 Standiford, R. B., 524–25 staphylinid beetles, 721 starfish, 268, 311. See also sea stars bat star, 320 morning sun star, 323 sunflower star, 321, 323 starry flounder, 373 starthistles maltese star-thistle, 433t Napa starthistle, 460, 461 yellow starthistle, 205, 231, 232–33, 454, 461, 466, 469, 513, 850 state-and-transition models, 845–46 State Highway Act of 1947, 909t, 911–12 State Water Project (SWP), 90, 730, 912 Steatoda, 650f Stebbins, G. Ledyard, 192 steelhead, 366, 373, 379t, 680, 724, 728, 792, 798, 829 as dwindling iconic species in coastal central and southern California, 735–37 Steinbeck, John, 337, 875 Stellaria media, 513 Stellar’s jay, 540f Steller’s sea cow, 163 Steller’s sea lion, 792 stem resistance to cavitation, 492, 493f stemflow, 498 Stenolepis gigas, 323 Stephanocystis, 317, 343 Stephen’s kangaroo rat, 206t, 207, 443 Stereolepis gigas, 769, 795 Sterna antillarum browni, 201, 379t, 401 Stevenson, K. M., 849 stewardship, conservation, and restoration, 925–27, 939 conservation, 928–29, 931 ecological restoration, 931–34 ecosystem stewardship framework, 927–28 habitat connectivity, 931 management status of protected lands, 929 percentage of surface area protected in various regions, 929 stewardship, ecosystem defined, 927 framework, 927–28 and policy across California’s ecosystems, future directions for, 934–37 principles, 928 Stewart, I. T., 19 Stewart, O. C., 180 Stewart, William, 523–24 Stichaeidae, 322 sticklebacks, 725, 727f stickyseeds, 192 stiletto flies, 415, 417 stilt, black-necked, 373f, 685 stilt sandpiper, 202

Stipa sp., 513 Stipa pulchra, 220, 450 stipe, 314 Stirling Management Area, 828 stock, fish. See fish stock(s) stock assessment, 782 stock-recruitment relationships, 781 stockers (livestock production), 841 Stoddard, J. L., 696, 697 stolons, 414. See also rhizomes stomata, 413 stomatal conductance, 518 Stone, Edward C., 416, 547 Stone and Timber Act of 1878, 87 stoneflies, 194, 722, 723 Stonemyia volutina, 189t Stoner, Reginald C., 155f stones, 57 storm petrels, 202 ashy storm-petrel, 202, 769, 770 fork-tailed storm-petrel, 200 Leach’s storm petrel, 202, 294, 296 stormwater runoff, 61, 64, 350, 671, 698, 891 Stralberg, D., 569 stratification (water), 368, 369, 488, 694, 696 stratus clouds, 12 strawberry, coast, 415, 416, 417f strawberry anemone, 321 Streptanthus, 192 Streptocephalus woottoni, 681 stress tolerance, 492 striped bass, 373 striped marlin, 292, 293 Strix nebulosa, 199, 201 Strix occidentalis, 37 Strix occidentalis caurina, 541, 561, 828 Strix occidentalis occidentalis, 201, 561 Strix varia, 200, 828 Stromberg, M. R., 851 strong interactions, 268 Strongylocentrotus, 343, 350 Strongylocentrotus sp., 268 Strongylocentrotus spp., 769 Strongylocentrotus franciscanus, 318, 319f, 769, 797 Strongylocentrotus purpuratus, 318, 319f, 769 Strongylocentrotus purpurescens, 797 sturgeons, 180, 371, 373, 381, 727f green sturgeon, 379t, 728 white sturgeon, 798 Sturnella neglecta, 459 Styela, 319f, 321 Styrax, 139 sub-replacement fertility. See replacement level subalpine adaptations to extreme physical conditions, 588 subalpine-alpine ecotone, 614f subalpine fir, 222, 557, 580, 593 subalpine forest ecosystem dynamics biogeochemical cycling and hydrology, 599–600 insects and pathogens, 599 wildlife, 598–99 subalpine forest ecosystem services, 605 agents of anthropogenic change climate change, 602 regional environmental changes, 602–4 biodiversity, 600–601 cool refugia under warming climates, 601 future climate scenarios, 604–5 management issues and climate adaptation, 604 recreation, scientific, and commodity services, 601–2

snowpack and water supply, 600 subalpine forest ecosystems, 579–80, 582 distribution, 580, 581f, 582 environmental controls current climate and climate variability, 583–88 geology, geomorphology, and soils, 582–83 mammals that use them as habitat, 594, 595–96f origins, 596–98 wildlife diversity, 594, 596 woodland structure, 579, 580f subalpine forest species, origins of, 596–98 subalpine forest zone. See subalpine zone subalpine meadows, 594 subalpine woodland, 222, 223f, 224 subalpine zone, 221 ecosystems, 588–96 subduction, 339 subduction zone, 760 submarine canyons, 393 subsidence inversion, 12, 110 subsoil. See B horizons subspecies, 375 subtidal habitats, 367 suburbanization, 80 succession in chaparral, 494–96 community, 648, 837 ecological, 346, 714 forest, 36, 542, 547, 561, 562f models of, 422, 837, 846 natural, 442, 443 postfire, 437, 441 vegetation, 415, 417, 421, 422, 480, 494– 96, 644, 648, 684, 837 successional pathways for historic mixedconifer forests, 562f successional processes, natural, 442 successional stages, 30, 417, 421, 461, 714, 928 suckers, 725, 727f, 728 sudden oak death (SOD), 86, 233, 236–37, 515, 542, 559, 829, 849 Suding, Katharine N., 438 suffrutescents, 437, 483, 485–86t, 487 sugar bush, 483, 484t, 490, 491 sugar pine, 145, 146, 221, 522, 536, 554t, 557, 559, 560t, 565f, 567, 599 Sula leucogaster, 202 sulfur dioxide (SO2), 118, 122t sulfur fungus, 515 Sulphur Bank Mercury Mine, 702 sumacs, 139 laurel sumac, 433t, 483, 491, 493 summer tanager, 202 sun star, morning, 323 sunfishes, 726, 727f green sunfish, 242 sunflower star, 321, 323 sunflowers, 146, 271, 272, 646, 768 common woolly sunflower, 193 sunlight zone. See euphotic zone supply-side ecology, 348 surf thistle, 417 surface fire regimes, 32 surface fires, 28, 29 surface horizons, 50, 60, 63, 67 surface runoff, 349 surfgrass, 343, 344, 346, 350, 352, 395f, 397, 787 surfperches, 322, 373, 399, 727f, 790 surplus production, 781 Sus scrofa, 229, 238, 241, 400, 514, 763 suspension feeders, 343, 395–97

INDEX  979

Sustainable Communities and Climate Protection Act of 2008, 911t, 917, 918, 936. See also Senate Bill 375 Sustainable Communities Strategies (SCSs), 919 Sustainable Fisheries Act of 1996 (SFA), 789 Swainson’s hawk, 459, 465 Swainson’s thrush, 256t Swallenia alexandrae, 193 swallows, 256t bank swallow, 724 barn swallow, 256t, 400 cliff swallow, 256t, 400, 724, 732 swamp crayfish, Louisiana red, 722, 726 swarming termites, 195 swash climates, 391 sweet cherry, 272 sweetgum, 138, 139 Sweetwater Mountains, 615, 617f swell shark, 322 swift, Vaux’s, 256t swimming crabs, 399 sword fern, 218, 542 swordfish, 292, 293, 769, 787 sycamores, 139, 724 western sycamore, 512 Sylvilagus, 162 Sylvilagus sp., 487t Sylvilagus spp., 457 Sylvilagus audobonii, 487t Sylvilagus bachmani, 487t, 514, 543 Symphoricarpos, 225 Syncaris nigrescens, 189t Syncaris pacifica, 681, 724 Synchirus gilli, 317f Syngnathidae, 322 Synodus, 319f Synodus lucioceps, 323 Synthliboramphus scrippsi, 201, 769 tabanid fly, Belkin’s dune, 415 Tabanidae, 189t, 415 tadpole shrimp, 641, 681 tadpoles, 675, 720, 726, 727, 853 Taeniatherum caput-medusae, 454 Tahoe, Lake. See Lake Tahoe Tahoe National Forest (TNF), 928 tailed frog, 197–98, 726 taking of private property rights, 900–901 Talitridae, 398 tall fescue, 461 Talluto, M. V., 438 Talpidae, 764 talus fields, 614 talus slopes, 621 Tamallia, 486 Tamarix spp., 238, 642, 660, 682 Tamias sp., 487t Tamias alpinus, 164, 206t, 595t, 627, 627t Tamias amoenus, 595t Tamias merriami, 487t Tamias minimus, 595t Tamias ochrogenys, 206t Tamias quadrimaculatus, 487t, 595t Tamias sonomae, 206t, 487t Tamias speciosus, 595t Tamias umbrinus, 595t Tamiasciurus douglasii, 595t, 596 tanagers, 202, 256t tanbark oak, 218, 480, 536. See also tanoak tanoak, 145, 146, 216, 236, 512t, 536, 539, 540, 542, 543, 559, 829. See also tanbark oak tansymustard, western, 193 Tapaja, 198

980  INDEX

taphonomy, 132 tapiroids, 157f, 159, 160 tapirs, 160 taproot, 646 tarantulas, 195 Taricha, 726 Taricha sp., 197 Taricha rivularis, 726 Taricha torosa, 726 tarns, 614, 622 tarweed, 192, 232, 450, 458 Tasmanian blue gum, 235 tassel, silk, 220, 483, 484t Tastiotenia, 650f Taxidea taxus, 459, 487t, 514, 516, 596t Taxodiaceae, 134 Taylor, A. H., 598 Tecate cypress, 40 tectonic processes, 693 Tegeticula antithetica, 654 Tegeticula synthetica, 654 Tegula, 318, 319f, 320 Tegula aureotincta, 318 Tegula eiseni, 318 Teiidae, 198 Teilhardina, 156, 157f Teleoceras, 162 temperature inversion, 12 tenebrionid beetles, 650, 651 Tenebrionidae, 415, 650f termites, 644, 650f, 651 swarming termites, 195 terns, 202 California least tern, 201, 379t, 401 terranes, 197 terrapins, 671 Tertiary, 32 tertiary consumers, 322–23 Tethya, 319f, 321 Tetradymia, 225 tetrigid grasshoppers, 721 Tettigoniidae, 189t Thaleichthys pacificus, 379t, 728 thalli, 345 Thamnophis sp., 459 Thamnophis atratus, 680, 721 Thamnophis elegans elegans, 697, 698f Thamnophis gigas, 198, 680 Thamnophis sirtalis, 721 Thamnophis sirtalis tetrataenia, 198 Thelazia californica, 828 Themiste, 343 Therevidae, 415 thermal inversions, 108 therophytes, 624 thicket, willow, 201 thicktail chub, 189t thimbleberry, 542 Thinopinus pictus, 398f, 399 thistles. See also starthistles artichoke thistle, 433t cobweb thistle, 417 lost thistle, 189t surf thistle, 417 Thomomys, 162 Thomomys bottae, 457, 514, 595t, 851 Thomomys mazama, 595t Thomomys monticola, 595t Thomomys spp., 164 Thoracophelia, 398–99 Thoracophelia dillonensis, 399 Thoracophelia mucronata, 398f, 399 Thoracophelia williamsi, 399 thornbush, Anderson’s, 648, 649f thornyheads, 794 thrashers

Bendire’s thrasher, 202 California thrasher, 201, 488t Le Conte’s thrasher, 199 threatened species. See endangered species thresher shark, common, 293, 796 Thryomanes bewickii, 488t Thuja, 139 Thuja plicata, 146, 537t Thunnus spp., 795 Thunnus alalunga, 292, 795 Thunnus albacares, 292, 795 Thunnus obesus, 795 Thunnus orientalis, 292 Thunnus thynnus, 795 Thysanoessa spinifera, 290, 296 thysanurans, 651 tidal creeks, dendritic, 673 tidal freshwater habitats, 367 tidal marsh bird races, 375 tidal marsh habitats, endangered/threatened species that require or use, 379t tidal marshes, California, 375 mammals and birds endemic to, 374–75, 376t tides and estuaries, 360, 363 Tidestrom’s lupine, 409–10f, 414f, 419 tidewater goby, 379t, 703 tiger beetles, 399, 417 hairy tiger beetle, 415 Ohlone tiger beetle, 195, 196f tiger salamanders, 459, 854 barred tiger salamander, 198, 232 California tiger salamander, 198, 232 465, 675, 680, 848, 851 tilapia, 704 Tilia, 139 Tiliaceae, 134 timber and nontimber forest products, 273 timber harvest. See also logging; lumber industry environmental stewardship and, 827t ownership class and, 821, 822f timberlands (of California), 830. See also forests habitat conservation plans on private, 828–29 harvested products from private and U.S. Forest Service, 822, 823t history, 820–21, 822f managing landscape-scale biodiversity on private, 827 Timema cristinae, 37 tiny artiodactyls, 161 tiphiid wasps, 650f Tipulidae, 723f titanotheres, 157f, 159, 160 titmouse, oak, 514 Tivela stultorum, 396, 396f toads, 198, 721, 726, 727f spadefoot toads, 653, 656, 726, 727f western toad, 438f, 721 Yosemite toad, 628 tobacco, coyote, 193 tocalote, 513 tomatoes, 271, 855 tongue, California fettid adder’s, 542 toothed whales, 294 top snails, 318, 343 topography and precipitation, 18 topsmelt, 321, 371 topsoil. See A horizons Torpedo californica, 323 torrent salamanders, 727f Torrey pine, 215, 483 Torreya, 139 Tortanus dextrilobatus, 371

tortoises, 198 total maximum daily load (TMDL), 699, 853–54 total nitrogen (TN), 728, 729f total phosphorus (TP), 695, 728, 729f tough rockweeds, 343 tourism, 84, 767–68 towhees, 488t, 543 California towhee, 201, 488t spotted towhee, 488t, 773 Toxicodendron diversilobum, 513 Toxostoma bendire, 202 Toxostoma lecontei, 199 Toxostoma redivivum, 201, 488t toyon, 433t, 483, 484t, 488, 491, 830 Trachemys sripta elegans, 727 Tracheophyta, 343, 716 Trachurus symmetricus, 291 tradewinds, 15 transhumance, 837 transient killer whale, 292f transient seed banks, 486 transit-oriented development, 891 transpiration, 498 transportation legislation (highways), 83, 909t, 911–12 transverse dunes, 412 Transverse Ranges, 110, 136, 199, 224, 480, 590 geology and geomorphology, 64–65 lowland soils, 66 mountain soils, 65–66 Transverse Ranges block, rotation of, 760, 762 trapdoor spider, 415 traps and pots (fishing), 785 trawling, 785 treaties. See international agreements tree cutting. See also logging; lumber for firewood and range management, 521–22 tree frog, Pacific, 438f, 459, 721 tree poppy, 139 tree squirrels, 160, 203 tree vole, California red, 206t treefish, 322 treeline, 579–80 treeline isotherm, 580 trees age, 823, 824f carbon stock in live, 824f defined, 579 treeshelters, 849 Tresus spp., 371 Triakis semifaciata, 796 triangleleaf bursage, 642, 647 Trichoptera, 194, 723, 723f tricolored blackbird, 680 Trifolium, 454 Trigonoscuta rossi, 189t Trigonoscuta yorbalindae, 189t trillium, western, 542 Trillium ovatum, 542 trim line, 620 Tringa semipalmata, 399f Trinity bristle snail, 828 trolling techniques (fishing), 785 trophic cascade, 268 trophic interactions, 288 trophic level. See tertiary consumers trophic pyramid, 648 trophic status, 694 trout, 188, 199, 237, 240, 260, 628, 696– 99, 717, 723, 728. See also cutthroat trout anadromous white trout, 792

food web of high-elevation Sierran lakes with introduced, 697, 698f rainbow trout, 237, 241, 699, 721, 726, 735 Trust for Public Land (TPL), 913 Tsuga, 139, 542f Tsuga heterophylla, 145, 218, 537t Tsuga mertensiana, 222, 536, 537, 580 tube snails, 343 tubeworms, 321 Tui chubs, Lahontan, 725 Tulare Basin, 671 Tulare Lake, 671 tule elk, 164, 188, 204, 455, 522, 671, 680, 681, 850, 852. See also Cervus canadensis nannodes tule fog, 22 tules, 359, 361f common tule, 671 tumbleweed, 648 tunas, 292, 293, 791f, 795 albacore tuna, 100, 292, 293, 790f, 791f, 794f, 795 bluefin tuna, 293 tundra, defined, 614 tungsten, 602 tunicates, 235f, 320, 321, 343, 371 colonial tunicate, 328 tunny fish, 376 Tuolumne River, 731f turban snails, 318, 321, 343, 352 wavy turban snail, 318, 797 Turbelliaria, 723f turbid estuaries, 369 turbot, diamond, 373 turfs, soft, 343 turkey vulture, 400, 658, 724 turkeys, 868t turpentine beetle, red, 599 turtles, 159, 162, 196, 198, 294, 727, 786. See also sea turtles Pacific pond turtle, 680 twig borers, 515 Tylos punctatus, 397f, 398 Tympanuchus phasianellus columbianus, 202 Tyrannus spp., 400 Tyto alba, 459 Uca crenulata, 371 uintatheres, 160 Ulex europeus, 452 Ulmaceae, 134 Ultisols, 50, 55, 57–59, 68, 555 ultramafic rock, 47, 56, 216 ultramafic soils, 56–57, 555 Ulva, 343, 722 Ulva spp., 368 Umbellularia, 139, 542f Umbellularia californica, 218, 237, 480, 512, 515, 539, 546f, 559, 829 Undaria pinnatifida, 328, 380 understory, 218 underwater acoustics, 297–98 Undulus parvipinnis, 371 ungulates, 158, 159, 594, 648, 657, 658, 659t, 767, 768, 851, 852 unicellular microzooplankton, 290 uniform subdivision regulations, 909t United States Bureau of Reclamation (USBOR), 684, 730 United States Department of Agriculture (USDA). See Natural Resource Conservation Service United States Forest Service (USFS), 549, 891 United States National Forests, 821 University of California, 2, 877

Blodgett Forest Research Station, 825 univoltine taxa, 722 upland alpine plateaus, 620 upland slopes and basins. See under alpine ecosystems upland soils, 60 Upogebia pugettensis, 371 Upper Arm (Clear Lake), 702, 703f upper montane zone, 221 upwelling, 12, 95, 96, 97f, 321, 322, 345, 396. See also intertidal ecology changes in, 99–100 urban activities, and their impact on surrounding areas, 890–91 urban areas of California, 75, 885, 887f, 892. See also cities, California’s future of, 891–92 urban-driver effects, 891 urban ecosystems of California, 885–86, 892. See also cities urban heat islands, 886 urban metabolism, ecological footprints, and life cycle assessment, 888–89 urban runoff, 374, 377, 378, 381, 736 urban sprawl, 80, 81, 888, 903 Urban Strategy, 917 urban vs. rural areas, percentage of population living in, 76 urban-wildlife interface (UWI), 886, 889 urbanism, new, 891 urbanization, 83, 872–73, 891. See also cities and agriculture, 439–40 urchin barrens, 317 urchins, 322, 330, 337, 343, 345, 350. See also sea urchins red urchin, 790f Uria aalge, 200, 293–94, 792 Urocitellus beldingi, 594, 627, 627t Urocitellus beldingii, 595t Urocyon cinereoargenteus, 442, 459, 514, 658 Urocyon littoralis, 188, 204, 206t, 400, 764 Urocyon littoralis catalinae, 773 Urocyron cineroargenteus, 487t Urophora sirunaseva, 233 Urosalpinx cinerea, 371 Urosaurus, 198 Ursidae, 161, 369, 596t Ursus americanus, 514, 596t, 627t Ursus arctos, 204, 541, 852 Ursus arctos californicus, 514, 926 Ursus arctos horribilis, 680 Urticina, 319f, 321 Uruguayan primrose-willow, 682 Urycyon cinereoargenteus, 400 usufructory right, 912 Uta, 198 Uta stansburiana, 438f, 653 Utah juniper, 141, 145, 146 Utah serviceberry, 594 Vaccinium, 540, 542f Vaccinium cespitosum, 594 Vaccinium ovatum, 539, 546f, 829 valley grassland, 220, 450, 836, 838–39, 843–44, 848. See also interior grassland valley oak, 214f, 510, 511f, 512, 514, 515, 517, 518, 521, 522, 525, 932 valley slope environments, 620 valley soils, 50–51, 58, 60 valley wild rye, 450 valleys agricultural, 89–90 montane, 50–51 Sierra Nevada, 50–51, 620 Van Pelt, R., 544 van Wagtendonk, J. W., 564

INDEX  981

vascular plants, 646–48, 649f, 673, 716 exotic, 231 origin of introduced, 229–30 Vasey, Richard B., 547 Vaux’s swift, 256t vedalia ladybeetle, Australian, 242 vegetation, 213–14, 226 climatic and environmental gradients, 214–16 defined, 213 local variation in, 216, 217 terminology, 213 vegetation classification, 216–17 vegetation distribution, 259, 260f vegetation models, dynamic global, 257–58 vegetation prehistory, 131–33, 149. See also specific topics lessons from the past for the future, 148–49 reconstructing the past, 133–34 vegetation types, 217, 225, 763f. See also National Vegetation Classification cool semidesert scrub and grassland, 224 cool temperate coastal vegetation, 217–18 cool temperate forests and woodlands of California mountains, 221–25 distribution of, 213–16 global warming and changes in total cover of, 260, 261f warm desert and semidesert scrub and grassland, 225 warm temperate forests and woodlands of Mediterranean California, 218–21 veligers, 345 velvet grass, 461, 850 velvet mites, giant red, 195 Veneroida, 723f Venerupsis philippinarum, 371 Venturan sage scrub, 430 verbenas. See sand verbenas vermilion flycatcher, 202 vernal pool fairy shrimp, 681 vernal pool tadpole shrimp, 681 vernal pools, 192, 675, 676f, 677, 677f, 704 vertebrates. See also specific topics in desert, 652–53, 656–58 numbers of endemic species and subspecies on Channel Islands, 765, 766t in oak woodlands, 514–15 in redwood forest, 541 riverine food webs and, 724 in wetlands, 680–81 Vertisols, 52, 55, 59f, 60, 65–67 vesicular horizons, 62–64, 640, 641 vesper sparrow, 459 Oregon vesper sparrow, 200 Vespertilionidae, 595t, 657 vetch, Borrego milk, 652 viburnum, 138, 139, 236 Viburnum, 138 Viburnum sp., 236 Vicars, W., 696 Vilà, M., 418 Viola sempervirens, 542 Viperidae, 198 Vireo bellii pusillus, 201, 202 Vireo vicinior, 202 vireos gray vireo, 202 least Bell’s vireo, 201, 202 warbling vireo, 256t Virginia opossum, 203, 231, 442 viruses, 460, 461, 773, 890 citrus virus, 886 West Nile virus, 200, 237, 684, 774

982  INDEX

viscous turbulent processes, 342 viticulture, 767 volatile organic compounds (VOCs), 108. See also semivolatile organic compounds volcanic rock, 58, 60 volcanic soil properties, 59–60 volcanic soils, 49, 617f volcanic terrain, soils of, 50 volcanism, 139–40, 582 voles, 162, 457–59, 487t California red tree vole, 206t California vole, 376t, 438f, 759t southern red-backed vole, 561 Voluntina Sonemyia fly, 189t Vulpes fulva, 459 Vulpes macrotis mutica, 847 Vulpes vulpes, 595t Vulpes vulpes necator, 204, 627t, 628 Vulpia microstachys, 513 Vulpia myuros, 433t waders, 685 Walker, Edwin, 177 wallflowers, 414 walnuts, 89, 134, 138, 139, 867 California walnut, 512 Walvoord, M. A., 645 wandering shrew, 376t warblers, 256t orange-crowned warbler, 256t, 765 Wilson’s warbler, 201, 256t warbling vireo, 256t “warm water pool,” 15–16 warty sea cucumber, 320, 326, 797 Washington clam, 371 Washingtonia filifera, 225 wasps, 515, 650f, 651 cynipid wasps, 516 parasitic wasps, 196 waste assimilation, 374 waste disposal from livestock confined animal feeding operations (CAFOs), 879, 880 water. See also estuaries; hydrology; runoff; snowmelt agriculture and, 872–74 in desert, provisioning services of, 658 montane forests and, 566, 569, 570f water beetles, 681, 698f water birch, 593 water boatman, 698f water cannons at Malakoff Diggins, 732f water deficit. See drought water distribution and allocation, 684. See also irrigation water districts, 876–77 water flea, 698f water hyacinth, 682 water loss, 342, 646, 650–53, 656–57 water mites, 681, 717 water pipit, 628 Water Plan. See California Water Plan water pollution, 349, 377–78. See also pollution pollutant/toxic runoff, 374, 671, 675, 730, 891 water potential, 413, 491 water problems, California’s, 738, 739. See also rivers water production and regulation, 273 water purification, 273–74 water quality, 350 and livestock production, 853–54 Water Resources, Department of. See Department of Water Resources water right allocations, 912

water scavenger beetles, 681 water shrews, 627t, 724 water smartweed, 679 water snakes, 727f water sources, 874. See also irrigation water storage and delivery system, components of, 732–33, 734f water stress, 491–93 tolerator/avoider strategies for coping with, 492 water table, 391 water table outcrop, 391 water turtles, 727f water use and supply, balance of, 876 water vapor, integrated, 15f waterbirds, 201, 670, 673, 681, 683–85 waterfowl, 201, 202, 372, 376, 381, 670, 671, 680, 683, 685–87, 704, 728, 733, 878 migratory, 201, 372, 675, 683, 685, 686, 728, 733 waterfowl habitat, 881 watergrass, 684 watermelon, 195, 271 watermilfoil, Eurasian, 682, 699 watersheds, 360 Watersipora, 328 waterweed, Brazilian, 682 wave exposure, 342 waves and estuaries, 360 wavy turban snail, 318, 797 weak stock management, 786 weasels, 162 long-tailed weasel, 627t weathering. See freeze-thaw cycles web-toed salamanders, 198 weevils, 189t, 233, 654 filbert weevil, 515 intertidal weevil, 398 weir, hupa salmon, 175, 176f West Nile virus, 200, 237, 684, 774 western azalea, 541 western blind snake, 198 western bluebird, 514 western Canada geese, 680 western fence lizard, 438f, 459, 486 western harvest mouse, 376t, 759t, 765 western hemlock, 145, 146, 218 western juniper, 147, 560t western kingbird, 256t western meadowlark, 459 western mosquitofish, 726 western pearl shell mussel, 725 western pika, 257 western pine beetle, 564 western pond turtle, 727 western rattlesnake, 459 western red bat, 724 western red cedar, 146, 536, 537t western sandpiper, 202 western scrub-jay, 200, 488t, 514 western shovel-nosed snake, 653 western snowy plover, 201, 202, 379t, 399f, 400, 401, 414, 421 western sycamore, 512 western tanager, 256t western tansymustard, 193 western toad, 438f, 721 western trillium, 542 western white pine, 146, 147, 222, 253, 536, 537t, 557, 559t, 564, 580, 589, 592, 596–99, 605 western white pine forests, 589 western wood-peewee, 256t western yellow bat, 654 Westman, W. E., 430 Westoby, M., 846

wet meadows, 56, 57, 147, 215f, 594, 600t, 601, 603, 622, 624, 625, 628, 675–77, 696 wet playas, 641 wetland ecosystem services, 683 carbon sequestration, 683 flood storage, 683 recreation, aesthetics, and valuation, 683–84 wetland management, 684–85 toxicity and pollution issues, 685 water distribution and allocation, 684 wetland restoration, 686–87 wetland vegetation, 678–79 vegetation types, 678f, 679 wetland vegetation management, 684 wetlands, 622, 623f. See also freshwater wetlands defined, 669–70 first uses of the term, 669 hydrology, 670, 673 runoff and, 671, 673, 675, 684, 686, 687 Wetlands Reserve Program (WRP), 686, 687 wetlands restoration, 931, 933 wetting front, 464 whale strikes, modifying shipping lanes to avoid, 300–302 whales, 157f, 163, 203, 204, 292f, 294–98, 300–302, 376, 792, 794 gray whale, 207, 297, 298, 300, 792 wheat, 89, 867, 870, 871f, 874 whelks, 345 Kellet’s whelk, 321, 797 whip-poor-will, Mexican, 202 whipple yucca, 146 Whipple’s monkey-flower, 190t whipsnake, Alameda, 198 whiptail lizards, 198 Whitaker, A. R., 178, 179 white abalone, 195, 196, 769, 796, 804 white alder, 512, 721 white anemone, large, 321 white bursage, 142, 147, 148, 642, 647, 649f white bursage desert scrub, 225 white-eared pocket mouse, 206t white-faced ibis, 201 white fir, 32, 221, 222, 224, 537, 554f, 557, 558f, 559, 560f, 561, 562f, 566, 580, 817, 821, 825 White Mountains, 224f, 626, 628 white oaks, 515 white pelican, American, 680, 704 white pine, western, 146, 147, 222, 253, 536, 537t, 557, 559t, 564, 580, 589, 592, 596–99, 605 white pine blister rust (WPBR), 564, 599 white pine forests, western, 589 white sage, 433t white sea urchin, 318 white seabass, 323, 790, 790f, 795 white shark, 293, 323, 796 white sturgeon, 798 white-tailed hare, 589 white-tailed jackrabbit, 627t white-tailed kite, 459 white-tailed ptarmigan, 628, 629 white trout, anadromous, 792 whitebark pine, 131, 141f, 146–48, 222, 253, 586, 588–93, 596, 599, 605, 615f whitebark pine cones, 589, 596, 597f whitebark pine ecosystems, 589 whitebark pine forests, 132, 587–90, 599, 603, 605 wind-sculpted and wind-thrown, 587 whitebark pine seeds and seedlings, 589, 596 whitebark pine woodlands, 598

whitefish, mountain, 725 Whittaker, Robert, 535 widow rockfish, 794 Wiese, A., 916 Wieslander, Albert, 902 wild bees, 271–72 wild buckwheat, 772 wild cucumber, 433t wild flax, dwarf, 192 wild ginger, 542 wild lettuce, 450, 458 wild oat, 433t, 452, 455, 457, 458, 460, 461 wild onion, 192 wild pig, 241, 514 wild ryes, 225, 450 Wild Scenic Rivers Act. See California Wild Scenic Rivers Act of 1972 wilderness areas, 84, 602 wildfire. See fire wildland-urban interface (WUI), 567 Wilkes, C., 177 willet, 373f, 399f Williams, J. A., 415, 420 Williams, W. T., 415, 420 Williamson Act of 1965, 910t, 913 Williamson’s sapsucker, 596 willow flycatcher, 200, 201 willow scrub, 678f willow thicket, 201 willows, 139, 145, 187f, 359, 361f, 418, 594, 600t, 624, 721, 724, 848 Wilson, Pete, 918 Wilsonia pusilla, 201 Wilson’s phalarope, 701 Wilson’s warbler, 201, 256t windward, 18 wind flow patterns, 15 seasonal, 110, 111f wind-forced upwelling, 96 wind scorpion, 195 wind-sculpted and wind-thrown whitebark pine forests, 587 wind(s), 565–66 Foehn, 21 northwesterly, 12 Santa Ana, 21, 31, 34 wingless walking-stick, 37 winterfat, 147, 648, 649f witch hazel, 134 Wohlgemuth, P. M., 38 wolf eel, 322 wolfberry, 147 Wolford, R. A., 706 Wolkovich, E. M., 435 wolverine, 204, 628 wolves, 164 wood borer (woodboring beetle), 114, 515 wood-peewee, western, 256t wood rat. See woodrats wood stork, 202 woodland salamander species, 198 woodlands, 224, 509–10. See also oak woodlands defined, 509 woodpeckers, 37, 201, 514, 516, 563 black-backed woodpecker, 562, 563f, 596 woodrats, 162, 204, 480, 486, 487t, 764. See also packrats/pack rats bushy-tailed woodrat, 627 desert woodrat, 654 dusky-footed woodrat, 514, 828–29 Woodruff v. North Bloomfield Mining and Gravel Company, 86 Wooley Creek Drainage, 219f woolly sunflower, common, 193 “working landscapes,” 855

World War I, 325–26 worms, 320, 337, 343, 459, 698f, 717 California oakworm, 515 earthworms, 459, 467, 515 filbertworm, 515 flatworms, 343, 717, 723f oligochaete worms, 681 pink bollworm, 242 plume worms, 343 polychaete worm, 320, 321, 328, 394, 395, 398, 399, 729 serpulid worm, 328 tubeworms, 321 wrack feeders, 397–98 wrasses, California small, 322 wrens Bewick’s wren, 486, 488t house wren, 256t marsh wren, 680 wrentit, 201, 438f, 486, 488t “X disease,” 108 Xanthium strumarium, 771 Xanthocephalus xanthocephalus, 202 Xantusia riversiana, 199, 765, 765f Xantusidae, 198 Xerces blue butterfly, 189t, 195 xeric crevice, 600t xeric moisture regime, 56 xero-riparian plant communities, 642 Xiphias gladius, 292, 769, 787 Xyleborinus saxeseni, 237 Xyleborus californicus, 237 xylem, 435, 483, 492, 493 yarrow, 416, 458 Yelenik, S. G., 436, 764 yellow bat, western, 654 yellow-bellied marmot, 594, 627, 630 yellow-billed magpie, 200, 201, 514 yellow bush lupine, 416, 417, 420, 421 yellow-cedar, Alaska, 222, 580, 593, 597 yellow-cheeked chipmunk, 206t, 540f yellow dwarf viruses, 461 yellow-footed gull, 202 yellow-green algae, 702 yellow-headed blackbird, 202 yellow-legged frogs, 697 foothill yellow-legged frog, 726, 727 mountain yellow-legged frog, 199, 238, 680, 697, 698f yellow pine, 32, 35, 557, 560–61, 567 yellow pond lily, 678f yellow rockfish, 323, 327f yellow sand verbena, 409–10f, 414f yellow starthistle, 205, 231–33, 454, 461, 466, 469, 513, 850 yellow warbler, 256t Yellowbanks, 775f yelloweye rockfish, 786 yellowfin croaker, 373, 399 yellowfin tuna, 292, 293, 795 yellowtail, 291, 323, 795 yellowtail rockfish, 321, 786, 795 yellowthroats common yellowthroat, 376t saltmarsh common yellowthroat, 201 Yolo Bypass, 739, 740f, 881 Yorba Linda weevil, 189t Yosemite National Park, 223f Yosemite toad, 628 Younger Dryas extraterrestrial impact hypothesis, 178 Yucca brevifolia, 145, 225, 636, 654 Yucca brevifolia jaegeriana, 654 yucca moths, 194–95

INDEX  983

Yucca spp., 195 Yucca whipplei, 146, 434 yuccas, 195, 654 whipple yucca, 146 Yuma clapper rail, 202 Zalophus californianus, 295, 767, 792 Zamia, 138 Zapodidae, 595t Zapus princeps, 595t Zapus trinotatus, 595t

984  INDEX

Zavaleta, Erika, 525 zebra mussel, 722 Zedler, P. H., 439 Zedler, R. P., 417 Zelkova, 139 Zenada macroura, 488t Zielinski, W. J., 541 Ziziphus parryi, 493 Zoarcidae, 322 zoea, 396 zonation (beaches), 395

zonation (in rocky intertidal), 346, 347f zoning, 916 zooplankton, 96, 368, 628, 694, 696, 697, 699, 701, 703, 756 crustacean, 289 zooplankton species in California Current, 290 photographs of dominant, 290f Zostera japonica, 380 Zostera marina, 368, 371, 372, 679 Zygnematales, 720