Ecology, Conservation, and Restoration of Tidal Marshes: The San Francisco Estuary 9780520954014

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The publisher gratefully acknowledges the generous contribution to this book provided by the Stephen Bechtel Fund.

Ecology, Conservation, and Restoration of Tidal Marshes

Ecology, Conservation, and Restoration of Tidal Marshes The San Francisco Estuary Edited by

Arnas Palaima

University of California Press Berkeley     Los Angeles     London

University of California Press, one of the most distinguished university presses in the United States, enriches lives around the world by advancing scholarship in the humanities, social sciences, and natural sciences. Its activities are supported by the UC Press Foundation and by philanthropic contributions from individuals and institutions. For more information, visit www.ucpress.edu. University of California Press Berkeley and Los Angeles, California University of California Press, Ltd. London, England © 2012 by The Regents of the University of California Library of Congress Cataloging-in-Publication Data Ecology, conservation, and restoration of tidal marshes : the San Francisco estuary / edited by Arnas Palaima. p.  cm. Includes bibliographical references and index. ISBN 978-0-520-27429-7 (cloth : alk. paper) 1. Wetlands—California—San Francisco Bay Area.  2. Estuaries— California—San Francisco Bay Area.  3. Ecology—California— San Francisco Bay Area.  4. San Francisco Bay Watershed (Calif.)  I. Palaima, Arnas. QH87.3.E34 2012 333.7309794'6—dc23   2012014652  

 

19 18 17 16 15 14 13 12 10 9 8 7 6 5 4  3 2 1 The paper used in this publication meets the minimum requirements of ANSI/NISO Z39.48-1992 (R 2002) (Permanence of Paper). Cover image: San Francisco Bay Area from International Space Station, Expedition 4. Photo from NASA.

to rita and vida To All that is One

Contents

Contributors  /  ix San Francisco Estuary Tidal Marshes (Map)  /  xii–xiii Global Salt Marshes (Map)  /  xiv

Foreword: Some Thoughts on San Francisco Bay and Its Wetlands  /  xv Paul Keddy Preface and Acknowledgments  /  xix Arnas Palaima

1 • Diverse Perspectives on Tidal Marshes: An Introduction  /  1

Joy B. Zedler

Part I  •  Ecology: Environment 2 • Historical Formation  /  21

5 • Pollution: Emerging Contaminants  /  67

Frances Malamud-Roam and Michelle F. Goman

Carol A. Vines and Gary N. Cherr

3 • Geomorphology, Hydrology, and Tidal Influences  /  35

Elizabeth Burke Watson

6 • Tidal Marshes in the Context of Climate Change  /  87

V. Thomas Parker, John C. Callaway, Lisa M. Schile, Michael C. Vasey, and Ellen R. Herbert

4 • Pollution: Persistent Organic Contaminants and Trace Metals  /  53

Hyun-Min Hwang, Peter G. Green, and Thomas M. Young

Part II  •  Ecology: Organisms 7 • Tidal Vegetation: Spatial and Temporal Dynamics  /  97

V. Thomas Parker, John C. Callaway, Lisa M. Schile, Michael C. Vasey, and Ellen R. Herbert

8 • Tidal Wetland Vegetation and Ecotone Profiles: The Rush Ranch Open Space Preserve  /  113

Christine R. Whitcraft, Brenda J. Grewell, and Peter Baye

9 • Invertebrates: Past and Current Invasions  /  135

Elizabeth D. Brusati 10 • Invertebrates: A Case Study of China Camp State Park, Marin County  /  147

April Robinson, Andrew N. Cohen, Brie Lindsey, and Letitia Grenier

12 • Bird Communities: Effects of Fragmentation, Disturbance, and Sea Level Rise on Population Viability  /  175

John Y. Takekawa, Isa Woo, Karen M. Thorne, Kevin J. Buffington, Nadav Nur, Michael L. Casazza, and Joshua T. Ackerman 13 • Small Mammals  /  195

Howard Shellhammer

11 • Fishes  /  161

Peter B. Moyle, James Hobbs, and Teejay O’Rear

Part III  •  Conservation and Restoration 14 • Ecosystem Services  /  207

Arnas Palaima 15 • Policy: Achievements and Challenges  /  215

17 • Natural and Restored Tidal Marsh Communities  /  233

Katharyn E. Boyer and Whitney J. Thornton

Marc Holmes 16 • Research Reserves as a Model for Conservation Science and Management of Tidal Marshes  /  225

Matthew C. Ferner

Index  /  263

18 • Current Issues in Tidal Marsh Restoration  /  253

John C. Callaway and V. Thomas Parker

Contributors

Joshua T. Ackerman

Gary N. Cherr

U.S. Geological Survey Western Ecological Research Center [email protected]

University of California–Davis [email protected]  

Andrew N. Cohen

Peter Baye

Center for Research on Aquatic Bioinvasions [email protected]

Annapolis Field Station [email protected]

Matthew C. Ferner

Katharyn E. Boyer San Francisco State University [email protected]

San Francisco Bay National Estuarine Research Reserve [email protected]

Elizabeth D. Brusati

Michelle F. Goman

California Invasive Plant Council [email protected]

Cornell University [email protected]

Kevin J. Buffington

Peter G. Green

U.S. Geological Survey Western Ecological Research Center [email protected]

University of California–Davis [email protected]  

Letitia Grenier

John C. Callaway

San Francisco Estuary Institute [email protected]

University of San Francisco [email protected]

Brenda J. Grewell

Michael L. Casazza

USDA-ARS Exotic & Invasive Weeds Research Unit / University of California–Davis [email protected]

U.S. Geological Survey Western Ecological Research Center [email protected]



 



ix

Ellen R. Herbert

April Robinson

San Francisco State University / Indiana University [email protected]

San Francisco Estuary Institute [email protected]

Lisa M. Schile

James Hobbs University of California–Davis [email protected]

San Francisco State University / University of California–Berkeley [email protected]

Marc Holmes

Howard Shellhammer

The Bay Institute [email protected]

H. T. Harvey & Associates / San Jose State University [email protected]

 

Hyun-Min Hwang Texas Southern University [email protected]

 

John Y. Takekawa

Paul Keddy

U.S. Geological Survey Western Ecological Research Center [email protected]

Lanark County, Ontario [email protected]

Karen M. Thorne

Brie Lindsey Oregon State University [email protected]

Frances Malamud-Roam University of California–Berkeley [email protected]  

Peter B. Moyle University of California–Davis [email protected]  

Nadav Nur

U.S. Geological Survey Western Ecological Research Center [email protected]

Whitney J. Thornton San Francisco State University [email protected]

Michael C. Vasey San Francisco State University [email protected]

Carol A. Vines University of California–Davis [email protected]  

PRBO Conservation Science [email protected]

Teejay O’Rear

Elizabeth Burke Watson

University of California–Davis [email protected]

U.S. Environmental Protection Agency, ORD-NHEERL, Atlantic Ecology Division [email protected]

Arnas Palaima

Christine R. Whitcraft

Ecological Economics Innovations Center [email protected]

San Francisco State University / San Francisco Bay NERR / California State University [email protected]

 

V. Thomas Parker San Francisco State University [email protected]

x

Isa Woo U.S. Geological Survey Western Ecological Research Center [email protected]

Contributors

Thomas M. Young

Joy Zedler

University of California–Davis [email protected]

University of Wisconsin [email protected]

 

Contributors

xi

Map 1.  Historical (1800, above) and modern (1998, next page) tidal marshes of the San Francisco Estuary. Courtesy of the San Francisco Estuary Institute.

Map 2.  Estimated salt marsh abundance around the world. The map identifies coastal regions that tend to have many salt marshes (darker areas) and those that have the fewest (lighter areas). Courtesy of the University of California Press and The Nature Conservancy.

Low or none

High

by Marine Ecoregion

Salt Marsh Abundance

Foreword: Some Thoughts on San Francisco Bay and Its Wetlands

crab, California halibut, and Pacific salmon all rely on the Bay as a nursery. Some 750 wildlife species, along with 120 species of fish, use the estuary (Nawi and Brandt 2008). San Francisco Bay may be exceptional on the West Coast, but it should not be treated as an exception. This bay exemplifies all the general principles that drive the formation of estuaries around the world. In all such tidal wetlands, a physical template controls the kinds of wetlands that arise and the particular plant and animal species that live in the wetlands. In general, the physical template is created by the flood regime, the nutrients, and natural agents of disturbance such as fire and grazing (Keddy 2010). In estuaries, salinity is also an important factor. In San Francisco Bay, this template is created and controlled by tides near the coast but also by inputs of freshwater and sediment from the rivers, the Sacramento and San Joaquin. Together, such rivers carry freshwater into the Bay at an average rate of about 24 million acre-feet per year (900 m3/s). The key ecological factors in the Bay are therefore (1) the area of silt and shoreline upon which wetland plants can potentially grow and (2) the salinity gradient, which ranges from fresh to saline. If you know how much of the Bay area is wet and how much wetland arises under each salinity regime, you can predict, rather well, the kinds of plants and animals that will occupy the Bay. The salinity gradient, of course, varies. Pulses of freshwater from the rivers will reduce salin-

Wetlands occur where land meets water. Wetlands, therefore, are not evenly spread over the surface of the Earth. Some areas, such as the Amazon and Siberia, have enormous wetlands, each of these being more than one million square kilometers. Most wetlands along the ocean coasts are much smaller, being limited on one side by uplands and on the other by ocean water that is too deep for rooted plants. The water levels in wetlands along coasts are influenced by tides, hence the general name tidal marsh. Tidal marshes can be further divided into two types, the narrow fringing marshes along steeper shorelines, and the somewhat larger areas of wetlands associated with estuaries and deltas. This book is about the latter type, wetlands in a large estuary, San Francisco Bay. The steep and rocky shorelines of California offer few estuaries and deltas. San Francisco Bay is therefore an ecologically important exception. Indeed, it is the biggest estuary on the west coast of the New World (Nawi and Brandt 2008). Here, two rivers that collectively drain nearly 40% of California meet and empty into the Pacific Ocean. These rivers deposit silt, and they dilute ocean salinity. Depending upon how you do the calculations, these rivers have created from one to four thousand square kilometers of estuary and tidal marsh. The rarity of such large wetlands on the West Coast magnifies their importance to wildlife. Millions of waterfowl using the Pacific flyway visit the wetlands and mudflats. The Dungeness



xv

ity, while periods of drought will lead to higher salinity. The range of variation and the extremes may be more important than the mean conditions. The idea that extreme conditions rather than mean conditions drive ecosystems has been around for a long time, but because extremes are often more difficult to observe and measure, we tend to overlook them. Tidal wetlands are a class of ecosystems where extremes must be considered. San Francisco Bay was formed over a long series of geological events, the most recent being the general rise in sea levels at the end of the ice age some twenty thousand years ago. During each ice age, sea levels fall as water is stored in vast continental glaciers. At the end of each ice age, the water from melting glaciers steadily raises sea levels. During the maximum of the last ice age, sea levels were about 100 m lower, and one large river drained through a canyon that is now called Golden Gate, into the Pacific Ocean. That would have been a sight to see. The current bay, then, the one with which we are familiar, is a flooded river valley. Prior ice ages led to similar alterations in ecological conditions. This book sensibly includes a chapter on these events. Here is another general principle that the Bay illustrates: global temperature and sea level are closely linked. Coastal wetlands are at great risk from changes in sea level, and whether you live in California, Louisiana, China, or Bangladesh, rising sea level from global warming is going to become an issue of overriding importance. Of course, at one time tidal marshes would move inland with rising sea levels, but now many are backed up against cities, and both the wetlands and the cities are in the path of tides and stormdriven surges. Chapter 6 by Tom Parker and coworkers has more to say explicitly about climate change, but if you pay attention, you will see that this theme emerges throughout the book. California, and Louisiana, could therefore lead the world in taking steps to control climate change. Or not. Speaking of human impacts, the first human visitors to San Francisco Bay were generally agreed to be Asians who had walked across the Bering Strait during the last ice age. Some of their migrations may also have been along the coast, since ocean currents naturally carry objects from north to south and since many of the tribal cultures included the use of boats. We can say xvi



with some confidence that the first human visitors likely saw the Bay around 10,000 years ago, when sea levels were still somewhat lower. The first recorded visits by Europeans were by Spanish seafarers in the late 1700s. At this time the Bay had some 2,200 km2 of wetlands. Now, after several centuries of human development, that figure has fallen to only 125 km2—a 95% rate of loss of original wetland area (U.S. Geological Survey 2003) (Map 1). The direct loss of wetlands over the last two centuries has been obvious. Large areas of wetland were filled with debris for development or cut off from the estuary by dikes to allow for agriculture. These direct impacts are fairly easy to map. Other factors, however, may be equally significant but less obvious. One of the greatest historical impacts of humans was hydraulic mining in the watershed in the 1860s. The debris was carried downstream, where it eventually filled in the eastern ends of the Bay. This not only changed the biology of the estuary but continues to this day to contribute to the expense of dredging, a significant federal subsidy to the shipping industry courtesy of the U.S. Army Corps of Engineers. Another significant impact was due to changes in the rivers themselves. Wetlands depend upon particular rhythms of high and low water levels, and every species, from fish to wading birds, times its life cycle to these highs and lows. Dams upstream disrupt the timing and duration of floods, thereby decreasing the area of wetlands and disrupting these cycles. Other effects from upstream disturbances such as deforestation and agriculture may be spread out more widely, but since the runoff eventually enters the rivers and concentrates in the Bay, these diffuse effects may be amplified in the Delta. The effects of changes in water quality, particularly with regard to nutrient levels, and toxic chemicals may therefore appear most severely in deltaic areas. Therefore, attempts to wisely manage the Bay may start with the most obvious local actions—infilling and diking—but eventually require that we consider human activities in the watershed as a whole. This requires us to expand our thinking—from considering first the remaining areas of wetlands themselves, through to thinking about the entire watershed that feeds water to these wetlands— that is, the entire Central Valley of California. The biological significance of the Bay is something I touched on above—think clouds of birds

Foreword

 

 

 

 

 

 

and schools of fish—and many of the later chapters in this book expand on this theme. We will let the authors speak for themselves, although this may be a good place to mention a recent book, San Francisco Bay: Portrait of an Estuary (words by John Hart and photographs by David Sanger), for a nontechnical treatment of these topics. And for a regional context to the Bay, two government scientists, Stephen Veirs and Paul Opler (1998), have written a succinct and nicely illustrated introduction to California, including both the original distribution of ecosystems and the many impacts of humans upon them. Apart from its biological significance, the Bay also has considerable cultural significance. The book says less about this, so let me give four examples. The writer Samuel Clemens, better known as Mark Twain, moved to San Francisco in 1864, where he began his career. The Golden Gate Bridge is a national engineering icon and forms a part of the human cultural landscape— familiar to people who have never seen the city except in films. San Francisco frequently appears in popular music—the 1967 hit “San Francisco (Be Sure to Wear Flowers in Your Hair)” became an iconic sound track to the cultural turmoil of the 1960s and to the war in Vietnam. The Bay also is an important location in films and television shows. For example, it is the site where a Klingon Bird of Prey crashes (in the future, when humans must return to the past to capture and release whales to repopulate Earth’s oceans). This episode of Star Trek IV manages to capture both the 1986 and the future San Francisco, with a riveting scene of an alien spaceship flying under the Golden Gate Bridge. And, on the topic of popular entertainment, if you want a view of San Francisco in the time of Mark Twain, there is a pair of episodes of Star Trek: The Next Generation, called “Time’s Arrow,” where some of the crew return to San Francisco and even meet Samuel Clemens (and a young Jack London too). Of course there are many more cultural examples that could be mentioned, but the point to emphasize is that the fate of the Bay is not just an issue of interest to the local humans. The Bay is part of our human heritage and has significance far beyond the seven million or so people who actually live in the immediate area. Returning to tidal wetlands, scientists have spent some decades trying to decide how to protect the existing wetlands and how to restore  

 

 

areas that have been degraded. There are many complicated technical issues that have to be considered, enough to keep some generations of biologists gainfully employed. At the same time, the issues are relatively simple. For those who live in the Bay Area, and those who watch from a distance, there are a few simple indicators we should be watching carefully. These are (1) the area of wetland, (2) the salinity levels in the Bay, (3) the amplitude and natural rhythm of high and low water levels as they change seasonally, and (4) the quality of the water in the rivers feeding the Bay. The good news is that if we get the water right, a lot of the other issues will fall into place naturally. That is where the topic of marsh restoration arises: the more we can restore the quantity and quality of water to natural conditions, the more we can protect the wetlands that still survive and, equally, restore the wetlands that have been degraded or lost. The last two chapters in this book tell us about recent progress in restoration. Here is where you will also have to master a new vocabulary—not the names of plants and animals, but the names of government agencies, research teams, and working groups. Most important among these is CALFED (California Bay-Delta Program and its successor, California Bay-Delta Authority), a joint federal-state agency that may, or may not, achieve the goal of sensibly managing the area. While “CALFED can provide a model for a functioning, collaborative ecosystem restoration program, it also demonstrates the difficulty of any effort to achieve success in resolving long-standing conflicts that involve both the protection and the utilization of a scarce and unique natural resource” (Nawi and Brandt 2008, 113). Of course, if CALFED does not survive, something will have to replace it, and the same problems will have to be solved: too many people vying for scarce resources, principally freshwater and food. This conveniently brings us back to general lessons that extend well beyond the Bay. We have already noted some general scientific principles: how the Bay illustrates the distribution of tidal wetlands around the world, how a few key environmental filters control tidal wetlands, how pulses of freshwater and periods of drought create natural extremes, how wetlands provide valuable ecological services, and how we can restore ecosystems by restoring the natural filters, including their normal patterns of variation. There is one more general principle that this

Foreword

 

xvii

book illustrates. We can view an area like San Francisco Bay as a test of our humanity. By humanity I mean a range of mental traits including compassion and rational behavior. Will we, like the shortsighted citizens of Easter Island, destroy ourselves by destroying our natural environment? Are greed, selfishness, and ignorance really so deeply rooted in our psyche that we cannot coexist with the other species with whom we share the Earth? That is the question. It is time to decide. Thus, somewhere on the shelf that holds this book, you may wish to also add Ronald Wright’s concise book, A Short History of Progress. How will we judge our progress? Although this book is about science conducted by humans, those in the jury are mostly not human. Ultimately the wetlands exist because of, and are used by, nonhumans, from the marsh grasses to the wading birds. If we do our part—by getting the water right—these other species will continue using the Bay. Scientists can study the history of the Bay, document the lives of the species that live there, and designate the ecological indicators for measuring progress in restoration. But in the end, the other species in the Bay will be the judge of whether we have done our jobs properly as scientists and citizens.  

 

Literature Cited Hart, J., and D. Sanger. 2003. San Francisco Bay: Portrait of an estuary. University of California Press, CA. www.sanfranciscobaybook.com. Keddy, P. A. 2010. Wetland ecology: Principles and conservation. 2nd ed. Cambridge University Press, Cambridge, UK. Nawi, D., and A. W. Brandt. 2008. The California BayDelta: The challenge of collaboration. In Largescale ecosystem restoration: Five case studies from the United States, edited by M. Doyle and C. A. Drew, 112–146. Island Press, Washington, DC. U.S. Geological Survey. 2003. Coastal wetlands and sediments of the San Francisco Bay system. USGS Fact Sheet. http://pubs.usgs.gov/fs/coastal-­ wet lands/index.html. Veirs, S. D., and P. A. Opler. 1998. California. In Status and trends of the nation’s biological resources, edited by M. J. Mac, P. A. Opler, C. E. Puckett Haecker, and P. D. Doran, 593–644. 2 vols. U.S. Department of the Interior, U.S. Geological Survey, Reston, VA. Wright, R. A. 2004. A short history of progress. Anansi Press, Toronto, ON.

Paul K eddy

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Foreword

 

 

Preface AND ACKNOWLEDGMENTS

failures and achievements, conservation and restoration efforts. As a result, the rich literature on the San Francisco Estuary tidal marshes can provide valuable lessons as well as solutions for tidal marshes (and other ecosystems as well) located in any part of the world. The purpose of this book is to review and integrate such knowledge and make it available to a broader audience, with a hope that old mistakes are not repeated and new awareness is spread and implemented. I am grateful to Jaime Kooser for planting the seed, Chuck Crumly for his encouragement and advice, and 31 reviewers for their comments that helped to improve this book. I was honored to work with 38 outstanding contributing authors, who did a great job in reviewing and synthesizing available information about tidal marshes of the San Francisco Estuary. I am also grateful to Danutė Januta, Jake Patrick Keenan, and Maria Adriaans for their financial contribution. This book would not have happened without the constant loving support of my wife, Rita Stanikūnaitė. A r na s Pa l a im a

Living Planet Report 2012, published by WWF with numbers from Global Footprint Network (www .footprintnetwork.org), revealed that humanity's Ecological Footprint has more than doubled since 1966. In 2008, the most recent year for which data are available, humanity used the equivalent of 1.5 planets to support its activities. Tidal marshes represent just one example of an ecosystem whose historical and present states accurately reflect the above-mentioned trend. Despite being instrumental, in providing ecosystem services to human society (in addition to their own intrinsic value), tidal marshes have been rapidly declining over the last century around the world, as land is given up to coastal development. The tidal marshes of the San Francisco Estuary fall into three broad overlapping categories: the freshwater marshes of the Delta, the brackish-water marshes of Suisun Bay, and the salt marshes of San Francisco Bay. Over time the San Francisco Estuary tidal marshes have experienced it all: undervaluation and destruction, research and monitoring, public education, protection policy

Berkeley, California 2012





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chapter One

Diverse Perspectives on Tidal Marshes An Introduction Joy B. Zedler

CONTENTS

regions. For the United States, current estimates of salt marsh area are for the Atlantic and Gulf of Mexico coasts, based on recent sampling of aerial photos by the National Wetland Inventory (Dahl 2006). As of 2004, salt marshes made up 5% (~1.6 million ha) of the total wetland area (~43.6 million ha), which in turn made up only 5.5% of the total land area (Dahl 2006). These estimates lead one to ask, Why is salt marsh area so restricted? Salt marshes develop in a dynamic environment that is shaped by tides, which influence sediment type, salinity, hydroperiods, and soil formation (Redfield 1965; Allen and Pye 1992). Fine sediments (clay and silt) accrete only in quiet waters along the shores of bays, estuaries, and lagoons. Salt marsh boundaries are dynamic, as indicated in a new model that illustrates how fringing marshes expand, contract, or drown in response to relative sea level rise, marsh aggradation processes, sedimentation rates, wave climate, and tidal range (Schwimmer and Pizzuto 2000). Where waves and currents are stronger, fine sediments are kept in suspension and only sands and cobbles accumulate, forming sand flats and beaches. Fine sediments are essential for plants to grow roots and further stabilize the substrate. Silt and clay can retain essential nutrients and moisture between high tides, thereby sustaining favorable growing conditions, whereas coarser substrates are easily drained during low tide and leached of nitrogen, thereby limiting plant growth (Langis et al. 1991). With the appropriate sediment as a

Global and National Perspectives A California Perspective Biological Perspectives Functional Perspectives Provisioning Services Regulating Services Supporting Services Cultural Services The Conservation Perspective Nutrient Pollution Elevation Shifts Contamination Species Invasions Global Change Earthquakes The Restoration Perspective The Research Perspective Synthesis and Future Directions

Global and National Perspectives Marine vegetated coasts make up less than 2% (90 million ha) of all ocean margins, and salt marshes comprise about 44% of that (~40 million ha), according to estimates by Duarte et al. (2005) (Map 2). Sea grasses and mangroves make up the rest, with the latter confined to frost-free



1

foundation, the daily rise and fall of seawater regularly inundates and exposes the shore. Below mean higher high water, the low marsh is inundated daily; at intermediate elevations, the marsh plain is inundated during spring tides every other week; and the upper extremes of the tidal range are inundated seasonally. While the gravitational pull of the moon and sun make “astronomic” tide levels highly predictable, the actual water levels vary in response to river inflows, atmospheric pressure, winds, waves, and sea swells. Seawater averages about 3.4% salt, and although daily tidal inundation ensures relatively constant salinity in the lower marsh, the higher intertidal elevations are subject to both the dilution and concentration of salts. When saline soils experience prolonged exposure during warm, dry weather, they become extremely hypersaline. When low tides coincide with rainfall or river flooding, intertidal soils can become brackish. Where major rivers flow into estuaries, fresh­ water can be pushed back up the river channel by incoming tides and released by outflowing tides, forming freshwater tidal marshes. If freshwater and saline tidal marshes within a region are compared, the effects of salt can be isolated from the effect of regular inundation. Salt reduces both plant diversity and productivity (Odum 1988; Barendregt et al. 2009). Salt marshes can extend inland beyond the highest predicted tides, in response to sea spray and extreme high-water events (e.g., storm swells), which carry salts inland. In San Francisco Bay, storms can elevate sea levels by 60 cm above predicted tides (Cayan et al. 2008). Where the transition from maximum predicted high tide to upland is relatively flat and undeveloped, an occasional high water can introduce sufficient salt to limit vegetation to halophytes of the upper salt marsh. Without saline soil, upland plants, and especially trees, have a competitive advantage over low-growing halophytes (Kozlowski 1997). Edith Purer (1942) was the first to hypothesize that biotic factors limit inland expansion of halophytes and that abiotic factors determine their lower-elevation limits within salt marshes. Just which biotic and abiotic factors are most responsible is not always clear, however. Support for Purer’s hypothesis comes from various transplantation experiments. As elevation increases and soil salinity decreases, glycophytes typically displace halophytes through competition, 2

although various studies also implicate predators, parasites, chemical inhibitors, and facilitators as causing more abrupt boundaries of salt marsh vegetation than would result from abiotic factors alone (see review by Ungar 1998). At the lowerelevation limit of salt marshes, abiotic factors such as inundation become increasingly stressful for intertidal halophytes, although there are few direct tests of hydroperiod effects, of either water depth or duration (Egan and Ungar 2000; Colmer and Flowers 2008). The water column decreases light exponentially, slowing photosynthesis and oxygen production at depth. Where aerobic microbiota deplete soil oxygen, soils become anoxic and anaerobic bacteria flourish. Some convert sulfur to sulfides, which blacken the soil and emit hydrogen sulfide—the rotten-egg smell that indicates sulfide toxicity. Unless sufficient oxygen can move from plant leaves through aerenchyma (air spaces) to roots, belowground tissues will suffocate. Upslope-downslope patterns of vegetation control are mirrored in river-to-estuary patterns (i.e., brackish and saline wetlands), with salt implicated as the principal abiotic stress for glycophytes (Latham et al. 1994). This hypothesis was tested in New England, using reciprocal transplants of 10 native species into plots with and without neighbors. As expected, glycophytes were stressed in salt marshes, and halophytes grew well in fresh marshes—but only where neighbors had been removed; otherwise, they were outcompeted by glycophytes (Crain et al. 2004). Salt is a strong abiotic stressor, and where salinity is low (holding inundation constant), competition adds a strong biotic stress. The rarity of appropriate abiotic conditions restricts tidal marshes to a fraction of continental and island shorelines where halophytes form seemingly uniform canopies of short-statured vegetation. But salt marshes that appear homogeneous from a distance are actually patchy mosaics of depressions and hummocks or ridges that vary at small vertical (20 years) enrichment experiment in a New England salt marsh involved a sewage-based fertilizer; at the highest dosage, plant species richness declined from 11 to 4 species, and dominance shifted from Spartina alterniflora to Distichlis spicata (Valiela et al. 1985). Elevation Shifts Whether there is a net gain or loss in sediment, salt marshes persist only where sedimentation rates are not too great, not too sparse, but just right (Schwimmer and Pizzuto 2000). The rates of subsidence and sea level rise determine how much is enough. On average, Louisiana alone loses some 4,300 hectares per year in response to subsidence, sediment starvation, canal dredging, levee construction, and other stressors (Craig et al. 1979).

Subsidence Subsidence causes coastal wetlands to sink. To compensate for the sinking soil surface, marshes need to grow root and rhizome mats that match or exceed the rate of subsidence. This does not happen where levees interrupt tidal flows, surface water becomes stagnant, soil anoxia persists, and plant growth is impaired.

Sediment Starvation Sediment starvation can reduce accretion to less than sea level rise. This problem arises where river water is diverted upstream or where modifications to waterways cause sediments to bypass coastal wetlands. Extensive levees along the lower Mississippi River protect adjacent properties from flooding; they also confine sediment flows to a few delivery points. To compensate, it is necessary to engineer structures to divert sediments into marshes so they can accommodate rising sea level (e.g., the 1991 Caernarvon diversion: Army Corps of Engineers 1998).

Diverse Perspectives on Tidal Marshes

Excess Sedimentation Watersheds with urban hardscapes discharge excess sedimentation in flashy episodes that increase erosion and carry more sediments toward salt marshes downstream. At Tijuana Estuary (San Diego County), river flooding has elevated the intertidal plain 25–35 cm since 1963 and increased dominance by Sarcocornia pacifica, which in turn has reduced vegetation diversity and evenness (Zedler and West 2008). Sediment has accumulated on the marsh plain at a rate >10 times that of sea level rise (0.02 cm/y).  

Contamination Urban salt marshes are particularly vulnerable to receiving runoff and pollutants, which can alter productivity, lower benthic diversity, impair fish behavior, and compromise food webs (Weis and Butler 2009).

Toxins Insecticides, herbicides, various hydrocarbons, and heavy metals negatively affect water, soils, and biota. Halogenated hydrocarbons can accumulate in animal tissues and cause immunosuppression, reproductive abnormalities, and cancer (Kennish 2002). Polycyclic aromatic hydrocarbons are both carcinogenic and mutagenic to aquatic organisms. Heavy metals can slow fish feeding, alter physiology and neurological function, impair reproduction, and cause genetic mutations. In anaerobic soils, mercury becomes methylated, a form that enables bioaccumulation in food chains. A challenge is to understand the interactions of multiple pollutants under changing environmental conditions.

Oil Spills In summer 2010, the largest oil spill in U.S. history contaminated the Gulf of Mexico. Every day, news reporters revealed devastating statistics about the number of days of spillage (>100), the volume of oil being spewed into the Gulf, the area of oil plumes, the length of shoreline with inflowing scum, bird and turtle mortality, and the inadequacy of modern technology to cap the leak, collect the oil, protect the beaches, and rescue the wildlife. But the far-reaching consequences

of major oil spills on the Gulf’s fragile marshes are not yet quantified. A worst-case scenario for a storm or hurricane that crosses the region’s vast salt marshes might read as follows: Oil spreads throughout the Gulf. Cloaks of oil render pelicans and other seabirds flightless and unable to feed. Winds carry the oil scum onto the salt marshes, where it smothers algae and invertebrates. As the waters recede, leaves are coated in oil, sunlight is blocked, stomata are clogged, and photosynthesis ceases. As the storm subsides, the dying plant canopy collapses and plant roots and animals suffocate under a toxic blanket. The survivors are anaerobic microbes that decompose the dead worms, arthropods, crustacea, mollusks, and plants. Because roots and rhizomes no longer build soil, the marsh surface sinks irreversibly. Open water replaces the huge areas of salt marsh that once fed the plankton that fed the shrimp and fish. The marsh cannot recover.

While the BP disaster has not produced this grim scenario, all it would take is the sequence of a major oil spill followed soon after by a major sea storm. Species Invasions Introductions into Salt Marshes Whether deliberate or accidental, the introduction of alien propagules sets in motion a chain of events that potentially leads to dominance by the invader and loss of native species. Nearby harbors experience continual “propagule pressure” (the number of propagules introduced) as ships discharge ballast water that contains spores, seeds, larvae, and adult organisms. Trains and cars on the many roads that cross salt marshes disperse alien propagules, adding to those brought in on the fur of mammals, on the feathers and feet of migratory birds, and in the guts of granivores (i.e., digestion-resistant seeds or resting stages of microinvertebrates). Humans are the most widely traveled species and the most common source of both deliberate and unintentional importation. In the 1970s, the U.S. Army Corps of Engineers deliberately imported Spartina alterniflora from the Atlantic coast to Puget Sound to stabilize dredge-spoil islands. Decades later, plantings and other introductions of S. alterniflora led to widespread

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9

invasions into Pacific coast salt marshes. While the introduction was intended, the invasion was unanticipated. In the 1980s, ships inadvertently introduced Corbula amurensis to San Francisco Bay following the resumption of trade with China (Nichols et al. 1990). In this case, both the introduction and the invasion were unintended. Just why some species become invasive while others do not remains a matter of debate. Some 29 causes have been hypothesized and recently sorted into three categories—propagule pressure and abiotic and biotic factors (Catford et al. 2009).  

Impacts to Salt Marshes

the native S. maritima to form S. townsendii (later forming S. anglica after chromosomal doubling) (Raybould et al. 1991). S. alterniflora also invaded San Francisco Bay (Callaway and Josselyn 1992) and hybridized with the native S. foliosa, yielding more aggressive offspring with broader distributions than either parent (Ayers et al. 2004). Global Change For salt marshes, there are three very predictable impacts of global change. The first is that sea level will rise more rapidly than without a warming climate. The average global sea level rise was 10–25 cm for the twentieth century. A further rise—up to 30 cm more—is predicted by 2040. As a result, tidal marsh species will need to move inland in order to sustain biodiversity. Two realities will constrain such migration: first, adjacent lands are already developed and unavailable for encroachment by wetlands; second, not all species will migrate rapidly enough. Effects on animals are also difficult to predict, although the British are breaching berms and restoring tidal flows to former salt marshes in a program of “managed retreat,” and avian ecologists are determining how well such efforts might sustain shorebird populations (Atkinson et al. 2004). Globally, efforts are under way to manage the impacts of rising sea level and to replace lost wetlands. A loss of biodiversity is virtually ensured, but which species, where, and how fast are less predicable. Another very predictable change involves warmer temperatures and elevated carbon dioxide levels. Warmer temperatures will increase rates of plant and animal respiration, evaporation, and evapotranspiration, at least in areas where increased precipitation and cloudiness do not reverse those effects. Higher average temperatures could favor C 4 plants, while more carbon dioxide should favor plant species with C3 photosynthesis, and salt marshes with a mixture of both C3 and C 4 species could gradually shift toward C3 species. The future of Spartina (a C 4 genus) is uncertain, as are the shorelines that it anchors. The third very predictable change is increased storminess, which could be manifested as both increased frequency and intensity (Cayan et al. 2008). More rainfall alone could shift the boundaries between saline and brackish conditions,  

 

Invasive plants spread rapidly, reduce diversity, alter habitat, shift trophic structure, and affect nutrient cycling. Alien strains of Phragmites australis became widespread in Spartina alterniflora marshes along New England coasts some 200 years after introduction. This invader releases an allelopathic substance (gallic acid) that reduces growth of native plants and persists in the soil even after the plant is removed (Rudrappa et al. 2007). Alien animals also spread rapidly, especially if they have planktonic larvae. The European green crab (Carcinus maenas) was found in San Francisco Bay in the 1980s; it is now established in the marshes of Oregon, Washington, and Vancouver Island (Behrens Yamada et al. 2005). In Bodega Bay, green crabs outcompete a native crab and two native mussels. Two invasive animals that can convert salt marshes to mudflats are nutria (Myocastor coypus), introduced from South America to the U.S. Southeast in the 1930s and which eat plant roots, and shipworms (an Australasian isopod, Sphaeroma quoyanum), whose burrowing in mud destabilizes tidal creek banks. Spartina alterniflora, which invaded 5,000 ha of mudflats in Yaquina Bay, Oregon, over the past 15 years, substantially reduced shorebird feeding habitat (Patten and O’Casey 2007).

Hybridization Some of the most aggressive invasions involve interbreeding between alien and native species. Because hybridization is difficult to detect, impacts can become widespread before they are recognized and after the opportunity for control has been lost. In Great Britain, S. alterniflora hybridized with 10

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but stormier weather will not only cause coastal watersheds to discharge more water, sediments, nutrients, and contaminants; it will also cause catastrophic erosion throughout the watershed and catastrophic deposition downstream. Storminess has already increased in one California salt marsh that experienced a decadal climate oscillation. The increased storminess followed urban development upstream, where unstable soils released tons of sediment that river floods carried to the estuary. There, sediment accreted and elevated the marsh plain. The historically diverse salt marsh vegetation shifted toward dominance by two productive species, Sarcocornia pacifica (=  Salicornia virginica) and Jaumea carnosa (Zedler and West 2008). Analysis of the loss of diversity during the 30-year stormy period indicated that the vegetation experienced direct, indirect, interactive, additive, unseasonal, and sequential effects of climate change (Zedler 2010). Of these, the greatest impact was a sequential effect, namely, tidal mouth closure followed by a long drought during the growing season. That sequence nearly extirpated two rare, shortlived plants (Zedler 2010). The sequential effect was not only the most severe; it was also the most difficult to predict. Because increased storminess will involve many events in unique combinations and sequences, specific impacts, times, and places are uncertain. For salt marshes that occur behind coastal dunes, increased storminess and storm surges that coincide with high tides could cause more extreme dune washover events that would move seawater and salt sprays inland, shifting glycophytic upland vegetation toward halophytes from high marshes and ecotones. These effects should be strongest and most persistent in regions with low rainfall, where salt deposits persist for decades in coastal upland soils. In general, a loss of biodiversity is predictable. Salt marshes have uncertain futures as climate changes. I predict that the impacts of frequent extreme events of greater magnitude will overshadow the effects of gradually rising sea level. Earthquakes One conservation concern that is not directly or indirectly caused by human activities is earthquakes, although there is new concern that offshore oil drilling could cause quakes. Tectonic

forces can result in either slumping or uplift of salt marshes, greatly altering environmental conditions and forcing major changes in the biota. In California’s Elkhorn Slough, the 1906 “Great Earthquake” (Richter scale: 8.25) lowered the Moss Landing Pier by 4 m and adjacent land by ~60 cm, a much greater impact than that of the 1989 Loma Prieta quake (Richter scale: 7.1) (National Estuarine Research Reserve 2010). The rapid change in elevation is what makes earthquakes so damaging to salt marsh biota; overnight, an entire community of organisms can be drowned or overexposed. In contrast, many species can adjust to gradual uplifting, as occurs in Scandinavia where the earth is still rebounding from the weight of historical glaciers. There, a fringe of mudflat is slowly exposed and colonized by halophytes, and upland vegetation shifts seaward as upper intertidal soils are leached of salts. Rapid, episodic shifts in elevation due to earthquakes are much more likely to reduce diversity.

The Restoration Perspective Tony Bradshaw (1987), who initiated “restoration ecology,” reasoned that to restore a modified or degraded ecosystem, we need to understand it. When we understand how tidal marshes function, we will be better able to restore them. Restoration, in turn, requires understanding of politics, community support, and funding, as well as dogged determination. What convinces the public and the decision makers to invest in the restoration of tidal marshes? It is easier to indicate how we should restore tidal marshes; the answer is “adaptively,” using field experiments. Salt marshes are amenable to experimental manipulation, and long-term field trials have proven extremely valuable in testing alternative ways to restore community structure and ecosystem functioning (Zedler 2001; Zedler and Callaway 2003; Larkin et al. 2008). Also, because methods of reintroducing lost species and ecosystem functions are still developing, restoration and research need to proceed synergistically. Some of the earliest restoration efforts, namely, planting Spartina alterniflora, involved scientific experimentation to stabilize the shores of North Carolina salt marshes (e.g., Broome et al. 1975; Seneca et al. 1976). Rapid methods were devel-

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oped using tobacco-planting equipment modified for wet soils. Much later, the experimental planting of S. alterniflora in Delaware using plants from local, northern, and southern sources showed that genotypic variation has implications for nearly every component of the food web (Seliskar et al. 2002), and in Louisiana, dredge-spoil addition accelerated S. alterniflora recovery where subsidence had lowered the marsh surface (Mendelssohn and Kuhn 2003). Large field experiments helped explain several problems that developed during restoration. For example, in San Diego Bay, S. foliosa plantings on sandy dredge spoil did not consider sediment type and were shown to be nitrogen limited (Langis et al. 1991). At Tijuana Estuary, the mass mortality of thousands of salt marsh seedlings was attributable to extreme hypersalinity, which developed because the marsh plain was not low enough and the planting time was delayed by problems during marsh excavation (Zedler et al. 2003). And replicate 1 ha marsh plains that were excavated without sufficient topographic heterogeneity were shown to benefit from incised tidal creek networks that reduced sediment accretion, enhanced halophyte establishment, and gave fishes greater access to resources on the marsh plain (Wallace et al. 2005; Larkin et al. 2008). The importance of topographic heterogeneity is supported by work in Oregon, where replicate channels dug within salt marshes enhanced salmon use (Cornu and Sadro 2002), and in Galveston Bay, Texas, where incising tidal channels increased fish support (Minello et al. 2011). In each case, field experimentation allowed researchers to identify causeeffect relationships. Restoration efforts have many beneficial outcomes, although the time required to achieve functional equivalency with natural wetlands can be decades. For example, at Tijuana Estuary, positive effects of planting diverse assemblages (Callaway et al. 2003) were short-lived (Doherty et al. 2011), indicating the need for longer-term assessment to understand restoration outcomes. That is, diversity-function relationships changed from positive to negative or nonsignificant after just 10 years. And soils in North Carolina salt marshes were not fully restored within 25 years (Craft et al. 1999). Experimentation and long-term study are both needed to develop restoration methods and understand why the best methods achieve the desired outcomes. 12

The Research Perspective Research in salt marshes has contributed a multitude of scientific advances, including the energysubsidy concept that S. alterniflora is more robust where tidal action is most vigorous (Steever et al. 1976); the outwelling concept that estuaries export detritus and support commercial marine fisheries (Odum 2009); the role of microorganisms in enriching detrital particles and enhancing food quality (Odum et al. 1979); the importance of algae to the food base (high quality through proteins and lipids) (Haines 1976); the influence on trophic dynamics by top-down, bottom-up, and other forces (Moon and Stiling 2002); differential limitation by phosphorus (inland) and nitrogen (salt marshes) (Howarth and Marino 2006); and connectivity among rivers, estuaries, and the ocean (Odum 1988). Variation within a single species, S. alterniflora, was related to latitude (Turner 1976), genotype (Seliskar et al. 2002), and temporal variability of tidal conditions (Morris et al. 2002), and determinants of community composition were shown to involve facilitation, not just competition (Bertness and Shumway 1993). Ecological methods were developed for assessing gross and net productivity, as were methods of quantifying nitrogen fixation, denitrification, and the biogeochemistry of sulfur and other elements (Reddy and DeLaune 2008). An approach that is now widely used to track the flow of organic matter into and through food webs—namely, stable isotope analysis—was initiated and developed in salt marshes (Haines 1976). Salt marsh ecologists continue to advance science by questioning paradigms (Weinstein and Kreeger 2000). Under scrutiny are that (1) coastal ecosystems are primarily nitrogen limited, but phosphorus can also be limiting, especially to microorganisms (Fong et al. 1993; Sundareshwar et al. 2003); (2) salt marsh herbivores consume less than 10% of primary productivity (not true where snow geese or snails congregate) (Kerbes et al. 1990; Silliman et al. 2005); and (3) grading a smooth intertidal plain will restore salt marsh functions. Where tested at Tijuana Estuary (Figure 1.1), tidal creeks were essential to ecosystem functioning (Larkin et al. 2008), and shallow depressions (5 cm deep) were needed to impound tidal water and stress the regional dominant (Sarcocornia pacifica) and prevent it from outcompeting an annual succulent, Salicornia bigelovii  

 

Diverse Perspectives on Tidal Marshes

(Varty and Zedler 2008). As indicated earlier, large field experiments that are established in an adaptive restoration framework are useful in testing alternative methods of conserving and restoring salt marshes (Zedler and Callaway 2003). Despite a strong history of scientific research and adaptive restoration, and despite extensive efforts to conserve and restore tidal wetlands, the United States continues to lose tidal marshes and their ecosystem services (Dahl 2006). San Francisco Bay offers many opportunities to sustain existing wetlands and restore former salt marshes. Monitoring and conducting research on attempts to restore San Francisco Bay salt ponds, minimize upstream water diversions, control invasive species, and adapt to climate change will contribute substantially to scientific knowledge and the practice of restoration while enhancing ecosystem services. This book helps point the way.

Synthesis and Future Directions A Global Perspective Tidal marshes are minor features, based on total area, but critical as assessed by the ecosystem services they perform. Salt marsh ecosystems contribute substantially to disturbance regulation, nutrient cycling, biological control, habitat, food production, recreation, and cultural services. Tides create a moving environment that is simultaneously dynamic and stabilizing; inflowing tidewaters import nutrients, plankton, and fish, and outflowing waters export wastes, such as salt that might otherwise accumulate to form salt flats. Intertidal conditions vary greatly within marshes (with elevation and inundation) and within estuaries (based on riverine influences, salinity, and sediment dynamics). Unlike Atlantic and Gulf Coast marshes, Pacific tidal marshes are mostly small and urbanized. In San Francisco Bay, several are still large, despite historical dredging and filling. A Biological Perspective Tidal marshes support salt-tolerant organisms with aquatic affinities (species stressed by exposure) and terrestrial affinities (stressed by inundation). Salt limits vascular plant diversity, but not micro- and macroalgae. Invertebrates abound in the wet sediment, and seawater adds diverse plankton and fish, while mammals and birds are

mostly limited to brief drainage periods between twice-daily high tides. The biota vary gradually (across horizontal and vertical gradients) and patchily (due to deposition-erosion dynamics, tidal creeks, and intertidal pools). Superimposed on the physical-chemical environment are patches of vascular plants that form discrete clones by reproducing vegetatively. While plants and animals differ in diversity, they share high productivity, which leads to highly valued ecosystem services. All salt marshes and their services are threatened by human impacts, namely, development, degradation, and rising sea level (a result of human-induced climate change). California’s many endangered, threatened, and sensitive species are strong evidence of past abuse. A Research Perspective Tidal marsh patterns and processes are not understood well enough to manage existing resources or improve future restoration efforts. Needed are ways to eradicate alien invasive species, restore native species and ecosystem services, and accommodate rising sea level. We do not yet know how to design a restoration site to support biodiversity and provide ecosystem services quickly and sustainably. And since societal need often creates opportunities for research, the future could be very bright. Attracted to compete for funding, teams of researchers could emerge and focus on improving restoration. Restoration projects can include experimentation that allows “learning while restoring.” Using large field experiments, researchers could identify basic design criteria, such as optimal sizes, configurations, and connections among habitats, and compare alternative approaches to reconstruct geomorphological and hydrological conditions. By comparing alternative ways to achieve desired outcomes, researchers could link physical-chemical conditions to biodiversity support and ecosystem services. Scientists can learn a great deal from mistakes, such as marsh inlets that won’t stay open and weed-control measures that unintentionally damage endangered birds. But tests of alternatives in an experimental, adaptive framework allow researchers to learn which approach works and why. It is the why part that is elusive. Once restoration pathways and outcomes no longer surprise us, we will have achieved an ideal for both science and society, namely, predictability.

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Acknowledgments I thank the National Estuarine Research Reserve System for setting aside Pacific coast salt marshes, Earth Island Institute for funding, coauthors of Zedler et al. 2008 for help reviewing literature, and Arnas Palaima for inviting this chapter.

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The importance of edge for natant macrofauna in a created salt marsh. Wetlands 14:184–198. Moon, D. C., and P. Stiling. 2002. Top-down, bottomup, or side to side? Within-trophic-level interactions modify trophic dynamics of a salt marsh herbivore. Oikos 98:480–490. Morris, J. T. 1991. Effects of nitrogen loading on wetland ecosystems with particular reference to atmospheric deposition. Annual Review of Ecology and Systematics 22:257–279. Morris, J .T., B. Kjerfve, and J. M. Dean. 1990. Dependence of estuarine productivity on anomalies in mean sea level. Limnology and Oceanography 35:926–930. Morris, J. T., P. V. Sundareshwar, C. T. Nietch, B. J. Kjerfve, and D. R. Cahoon. 2002. Responses of coastal wetlands to rising sea level. Ecology 83:2869–2877. National Estuarine Research Reserve. 2010. Elkhorn Slough NERR. http://www.elkhornslough.org/. National Oceanic and Atmospheric Administration (NOAA). 2010. San Francisco Bay Environment. http://mapping2.orr.noaa.gov/portal/­sanfrancisco bay/sf b_html/sf benv.html. Nichols, F. H., J. K. Thompson, and L. E. Schemel. 1990. Remarkable invasion of San Francisco Bay (California, USA) by the Asian clam Potamocorbula amurensis. II. Displacement of a former community. Marine Ecological Progress Series 66:95–101. Noe, G. B., and J. B. Zedler. 2001. Spatio-temporal variation of salt marsh seedling establishment in relation to the abiotic and biotic environment. Journal of Vegetation Science 12:61–74. Odum, E. 2009. Tidal marshes as outwelling/­pulsing systems. In Concepts and controversies in tidal marsh ecology, edited by M. P. Weinstein and D. A. Kreeger, 3–7. Kluwer Academic, Boston. Odum, W. E. 1988. The comparative ecology of tidal fresh-water and salt marshes. Annual Review of Ecology and Systematics 19:147–176. Odum, W. E., P. W. Kirk, and J. C. Zieman. 1979. Nonprotein nitrogen compounds associated with particles of vascular plant detritus. Oikos 32:363–367. Patrick, W. H., and R. D. DeLaune. 1972. Characterization of the oxidized and reduced zones in flooded soil. Soil Science Society of America Journal 36:573–576. Patten, K., and C. O’Casey. 2007. Use of Willapa Bay, Washington, by shorebirds and waterfowl after Spartina control efforts. Journal of Field Ornithology 78:395–400. Pomeroy, L. R., and R. G. Weigert. 1981. The ecology of a salt marsh. Springer, New York. Purer, E. 1942. Plant ecology of the salt marshlands of San Diego County. Ecological Monographs 12:82–111.  

 

 

 

 

 

 

 

 

 

 

 

 

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Raybould, A. F., A. J. Gray, M. J. Lawrence, and D. F. Marshall. 1991. The evolution of Spartina anglica C. E. Hubbard (Gramineae): Origin and genetic variability. Journal of the Linnaean Society 43:111–126. Reddy, K. R., and R. D. DeLaune. 2008. Biogeochemistry of wetlands. CRC, Boca Raton, Florida. Redfield, A. C. 1965. Ontogeny of a salt marsh estuary. Science 147:50–55. Rudrappa, T., J. Bonsall, J. L. Gallagher, D. Seliskar, and H. Bais. 2007. Root-secreted allelochemical in the noxious weed Phragmites australis deploys a reactive oxygen species response and microtubule assembly disruption to execute rhizotoxicity. Journal of Chemical Ecology 33:1898–1918. Schwimmer, R. A., and J. E. Pizzuto. 2000. A model for the evolution of marsh shorelines. Journal of Sedimentary Research 70:1026–1035. Seliskar, D. M., J. L. Gallagher, D. M. Burdick, and L. A. Mutz. 2002. The regulation of ecosystem functions by ecotypic variation in the dominant plant: A Spartina alterniflora salt-marsh case study. Journal of Ecology 90:1–11. Seneca, E. D., S. W. Broome, W. W. Woodhouse Jr, L. M. Cammen, and J. T. Lyon III..1976. Establishing Spartina alterniflora marsh in North Carolina. Environmental Conservation 3:185–188. Silliman, B. R., J. van de Koppel, M. D. Bertness, L. E. Stanton, and I. A. Mendelssohn. 2005. Drought, snails, and large-scale die-off of southern U.S. salt marshes. Science 310:1803–1806. Sousa, A. I., A. I. Lillebø, M. A. Pardal, and I. Caçador. 2010. Productivity and nutrient cycling in salt marshes: Contribution to ecosystem health. Estuarine, Coastal and Shelf Science 87:640–646. Stedman, S., and T. E. Dahl. 2008. Status and trends of wetlands in the coastal watersheds of the Eastern United States 1998 to 2004. NOAA National Marine Fisheries Service and U.S. Fish and Wildlife Service, Washington, DC. Steever, E. Z., R. S. Warren, and W. A. Niering. 1976. Tidal energy subsidy and standing crop production of Spartina alterniflora. Estuarine and Coastal Marine Science 4:473–478. Sullivan, J. M. 1982. Distribution of edaphic diatoms in a Mississippi saltmarsh: A canonical correlation analysis. Journal of Phycology 18:130–133. Sullivan, M. J. 1975. Diatom communities from a Delaware salt marsh. Journal of Phycology 11:384–390. Sundareshwar, P. V., J. T. Morris, E. K. Koepfler, and B. Fornwalt. 2003. Phosphorus limitation of coastal ecoystem processes. Science 299:563–565. Teal, J. M. 1962. Energy flow in the salt marsh ecosystem of Georgia. Ecology 43:614–624. Thurman, C. 2003. Osmoregulation in fiddler crabs  

 

 

 

 

 

 

 

 

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(Uca) from temperate Atlantic and Gulf of Mexico coasts of North America. Marine Biology 142:77–92. Traut, B. H. 2005. The role of coastal ecotones: A case study of the salt marsh / upland transition zone in California. Journal of Ecology 93:279–290. Turner, R. E. 1976. Geographic variations in salt marsh macrophyte production: A review. Contributions in Marine Science 20:47–68. Ungar, I. 1998. Are biotic factors significant in influencing the distribution of halophytes in saline habitats? Biological Reviews 64:176–199. Ungar, I. 1991. Ecophysiology of vascular halophytes. CRC, Boca Raton, Florida. Valiela, I., and J. M. Teal. 1974. Nutrient limitation in salt marsh vegetation. In Ecology of halophytes, edited by R. J. Reimold and W. H. Queen, 547–563. Academic Press, New York. Valiela, I., J. M. Teal, C. Cogswell, J. Hartman, S. Allen, R. Van Ellen, and D. Groehringer. 1985. Some long-term consequences of sewage contamination in salt marsh ecosystems. In Ecological considerations in wetlands treatment of municipal wastewaters, edited by P. J. Godfrey, E. R. Kaynor, S. Pelczarski, and J. Benforado, 301–316. Van Nostrand Reinhold, New York. Valiela, I., J. M. Teal, S. Volkmann, D. Shafer, and E. J. Carpenter. 1978. Nutrient and particulate fluxes in a salt marsh ecosystem: Tidal exchanges and inputs by precipitation and groundwater. Limnology and Oceanography 23:798–812. Varty, A., and J. B. Zedler. 2008. How waterlogged microsites help an annual plant persist among salt marsh perennials. Estuaries and Coasts 31:300–312. Wallace, K. J., J. C. Callaway, and J. B. Zedler. 2005. Evolution of tidal creek networks in a high sedimentation environment: A 5-year experiment at Tijuana Estuary, California. Estuaries 28:795–811.  

 

 

 

 

 

 

 

 

Weinstein, M. P., and D. A. Kreeger. 2000. Concepts and controversies in tidal marsh ecology. Kluwer Academic, Boston. Weis, J. S., and C. A. Butler. 2009. Salt marshes: A natural and unnatural history. Rutgers University Press, Newark, New Jersey. Williams, P., and P. Faber. 2001. Salt marsh restoration experience in San Francisco Bay. Journal of Coastal Research Special Issue 27:203–211. Zedler, J. B. 1993. Canopy architecture of natural and planted cordgrass marshes: Selecting habitat evaluation criteria. Ecological Applications 3:123–138. Zedler, J. B. 2010. How frequent storms affect wetland vegetation: A preview of climate change impacts. Frontiers in Ecology and the Environment 8:540–547. Zedler, J. B. 2001. Handbook for restoring tidal wetlands. CRC, New York. Zedler, J. B. 1982. Salt marsh algal mat composition: Spatial and temporal comparisons. Bulletin of the Southern California Academy of Science 81:41–50. Zedler, J. B., C. L. Bonin, D. J. Larkin, and A. Varty. 2008. Salt marshes. In Encyclopedia of ecology (1st ed.), edited by Sven Erik Jorgensen and Brian D. Fath, 3132–3141. Elsevier B. V., Oxford, England. Zedler, J. B., and J. C. Callaway. 2003. Adaptive restoration: A strategic approach for integrating research into restoration projects. In Managing for healthy ecosystems, edited by D. J. Rapport, W. L. Lasley, D. E. Rolston, N. O. Nielsen, C. O. Qualset, and A. B. Damania, 167–174. Lewis, Boca Raton, Florida. Zedler, J. B., H. N. Morzaria-Luna, and K. Ward. 2003. The challenge of restoring vegetation on tidal, hypersaline substrates. Plant and Soil 253:259–273. Zedler, J. B., and J. M. West. 2008. Declining diversity in natural and restored salt marshes: A 30-year study of Tijuana Estuary. Restoration Ecology 16:249–262.  

 

 

 

 

 

 

Diverse Perspectives on Tidal Marshes

17

Part one

Ecology Environment

chapter Two

Historical Formation Frances Malamud-Roam and Michelle F. Goman

contents

exists to the south, the South Bay (Atwater 1979), but experiences significantly lower freshwater input, so a similar salinity gradient is not present there. The northern reach experiences a mean tide level 0.2 m higher than the southern reach, while the latter has a larger tidal range, with 2.6 m in the South Bay, 1.7 m at the Golden Gate, and 1.3 m at Suisun Bay (Conomos 1979; Conomos et al. 1985). This chapter reviews the natural history of the Estuary and its tidal marshes, largely focusing on the period prior to European contact and subsequent environmental impacts. Other sections of this volume, however, address the modern setting, as well as some of the changes in the Estuary that have occurred as a result of human activities in the late nineteenth and twentieth centuries.

Geologic History Sea Level Rise and the San Francisco Bay since the Last Glacial Maximum Tidal Marsh Communities of the San Francisco Bay Freshwater Sources and Their Variability Holocene Freshwater Flows and Tidal Marsh History Human Impacts Synthesis and Future Directions

O

n the central coast of California, a series of bedrock basins and narrow structural constrictions (straits) have produced what we recognize today as the San Francisco Bay Estuary (Conomos et al. 1985; Goals Project 1999). Pacific Ocean water passes through the Golden Gate and enters the Central Bay of the Estuary; from here saline water flows upriver with the tides into San Pablo Bay, through the Carquinez Strait, into Suisun Bay, and finally into the Sacramento–San Joaquin Delta. This water mixes with the combined freshwater of the Sacramento and San Joaquin Rivers (Figure 2.1) that flow westward through the Delta. Thus, the northern reach of the Estuary experiences a gradient of generally decreasing salinity and tidal character with distance from the Golden Gate (Atwater et al. 1979; Malamud-Roam et al. 2006a). An additional basin

Geologic History The San Francisco Estuary is the recurring product of a specific set of conditions resulting from global climate and sea level dynamics and as such is an ephemeral feature of the California coast when considered on geologic timescales (­Malamud-Roam et al. 2006a). The Quaternary period, covering approximately the last 2 million years, has experienced large, regular oscillations in world climate, between cool glacial phases, when large areas of the world’s high-latitude continents were covered with thick ice sheets and global sea level was low, and warmer interglacial phases such as today, when much of the continental ice sheets melted, causing higher

 





21

122°W

Sacramento River

10 km

N

38°N

San Joaquin River

Bay core Tidal marsh core Tidal wetland circa 1850

Pacific Ocean

Figure 2.1.  Map of San Francisco Bay Estuary and its watershed. Research sites on the tidal marshes and in the estuary are indicated by circles.

0

San Francisco San Francisco Bay

Meters

–20

Alameda

A

–40

B

–60

–80

C

E

D 0

F G

0

1mile 2km

Bay mud Alluvium Dune sand Rockland ash Figure 2.2.  Sediment layers beneath the San Francisco Bay Estuary between San Francisco and Alameda, the site of a proposed “southern crossing.” Layer A contains the modern interglacial estuarine sediments (Bay mud). Layers B, C, and D are previous interglacial estuary sediments; E, F, and G may be associated with C and D. Alluvial sediments, comprising river-borne sediments deposited during glacial periods, alternate between the estuary sediments. Adapted, with permission, from Sloan 2006.

Change relative sea level (m) from modern

-5

-25

-45

-65 Tahiti -85 New Guinea -105

Barbados

-125 3

6

12 9 Time (1,000 years BP)

15

18

Figure 2.3.  Sea level curve since the last glacial maximum. Adapted from Quinn 2000, with source data from Fairbanks 1989, Chappell and Polach 1991, Edwards et al. 1993, and Bard et al. 1996.

sea levels (Hays et al. 1977). These cycles, driven by astronomical forcing, generally occur on a 100,000-year scale, with long glacial periods (~90,000 years) and short interglacial periods (~10,000 years) (Bassinot et al. 1994; Shackleton and Opdyke 1976). An estuary formed in the San Francisco Bay during at least the last four interglacial periods (Atwater et al. 1977; Sloan 2006). However, the stratigraphic evidence for earlier interglacial estuaries was largely removed during the following glacial periods, when a wide river valley replaced the estuary and erosion dominated the landscape (Sloan 1992). A stratigraphic profile was constructed from a series of boreholes drilled along a transect between Alameda and southern San Francisco during the planning of a proposed “southern crossing” bridge, and it portrays the depositional history of the last several hundred thousand years (Figure 2.2) (Sloan 2006). Layers of “Bay mud” (estuarine sediments) are interspersed with alluvium (river sediments deposited during glacial periods) representing several interglacialglacial cycles in the San Francisco Bay. During the penultimate interglacial period, about 135,000

years ago, eustatic (global) sea level was about 6 m higher than in the modern interglacial (Chen et al. 1991), resulting in a larger estuary (Sloan 1992). Ingram and Sloan (1992) reconstructed a salinity history of this early incarnation of the estuary and found periods when conditions were significantly fresher than in modern times, which they attribute to climate fluctuations.

Sea Level Rise and the San Francisco Bay since the Last Glacial Maximum Two conditions are critical for tidal marsh establishment and persistence: protection from severe storms (Mitsch and Gosselink 2000; Zedler 2001) and adequate sediment supply (Frey and Basan 1985; Pethick 1992; Trenhaile 1997). Both are influenced by sea level rise. The recent glacial history of San Francisco Bay, since the last glacial maximum, provides the context for a discussion of tidal salt marsh formation around the edges of the Estuary. The last glacial maximum was about 21,000 years ago (Kutzbach et al. 1998), when eustatic sea level was at least 110–120 m lower than present (Figure 2.3) (Fairbanks 1989; Ruddiman 2001),

Historical Formation

 

23

Figure 2.4.  San Francisco Bay Estuary formation. (Upper) During glacial times, the shoreline was out beyond the Farallon Islands and a river valley occupied the site of the San Francisco Bay Estuary. (Lower) As global sea level rose since the height of the last glacial period, the shoreline gradually moved inland, passing through the Golden Gate approximately 10,000 years ago. Adapted from Sloan 2006, using data from Atwater 1979 and Atwater et al. 1977.

though others argue even lower, 130–140 m (Clark et al. 2004; Issar 2003). Relative (local) sea levels differ because of crustal movement (e.g., Atwater and Hemphill-Haley 1997), subsidence (Patrick and DeLaune 1990; Watson 2004), or diversions  

24



in glacial meltwater (e.g., Broecker et al. 1989; Fairbanks 1989). Tidal marshes most likely existed in fragmented pockets along the Pacific coastline during glacial times, migrating with the changing shorelines as postglacial sea levels rose

Ecology: Environment

(Atwater 1979; Malamud-Roam et al. 2006b). Sea water entered the river valley that would eventually become the current San Francisco Bay about 10,000 years ago (Figure 2.4) (Atwater et al. 1977; Atwater 1979). Initially, the Bay filled rapidly (e.g., as measured in southern San Francisco Bay by Atwater et al. 1977). This rapid rise (6–8 mm/year or more) precluded the early development of extensive tidal marshlands along the edges of the Bay until the rate slowed to about 1–2 mm/year globally and locally, about 6,000 years ago (Atwater 1979; Fairbanks 1989; Goman and Wells 2000). The oldest extant tidal marshes in the San Francisco Estuary became established following this global decline in the rate of sea level rise; however, most tidal marshes formed in the Estuary within the last 5,000 years (Atwater et al. 1979; Byrne et al. 2001; Goman and Wells 2000; Malamud-Roam and Ingram 2004). The geologic constriction that forms the Golden Gate provided a sheltered habitat around the newly forming Bay by buffering the coastal interior from high-energy conditions along the California coastline (National Oceanic and Atmospheric Administration 2003). Tidal marsh formation, health, and stability are also affected by sediment supply. If sediment supply had been sufficient during the early Holocene (10,000–6,000 years ago), marshes could have become established even at the high rates of sea level rise. Indeed, tidal marshes in the southern San Francisco Bay were able to persist during a period of rapid relative sea level rise that occurred in the last century as groundwater extraction caused rapid subsidence in the area (Patrick and DeLaune 1990; Watson 2004). In particular, Alviso Marsh experienced on average 3.9 cm y-1 of accretion between 1955 and 1983.  

 

 

Tidal Marsh Communities of the San Francisco Bay We provide here a brief synopsis of the dominant tidal marsh plant communities, so as to provide a context for paleovegetation reconstructions. For a more complete discussion of the plant communities of the tidal marshes in the San Francisco Bay Estuary, see Chapter 7, this volume. An east-west salinity gradient exists in the Estuary, reflected in vegetation composition on the marshes and overall biomass, which increases eastward as salinity declines (Conomos 1979; Goman 2001). Along the

northern reach of the Estuary, tidal salt marshes are found as far east as the western edge of the Carquinez Strait and are dominated by cordgrass (Spartina foliosa Trin.) growing in near monospecific bands at the interface between Bay waters and the marsh and by pickleweed (Salicornia virginica L.) in the slightly higher elevations. Other species are interspersed in the higher marsh, including Distichlis spicata L., Jaumea carnosa (Less.) A. Gray, and Grindelia stricta DC (Atwater et al. 1979; Goman 1996; Malamud-Roam and Ingram 2001). Brackish marshes form up-estuary of the Carquinez Strait extending east to the Delta. These marshes have greater biomass and plant diversity (Atwater and Hedel 1976). Brackish marshes are typically dominated by a variety of species of bulrush (Schoenoplectus), with species composition depending upon the local salinity regime, which varies over multiple timescales. The tidal marshes experience fresher conditions nearer to the Delta and are dominated by Schoenoplectus acutus, Typha angustifolia, T.  latifolia L., and Phragmites communis Trin.

Freshwater Sources and Their Variability Thus far we have primarily discussed oceanic influences on the San Francisco Bay Estuary. However, freshwater inflows to the Estuary have a direct and immediate impact on estuarine salinity and thus on tidal marsh ecosystems (Atwater and Hedel 1976; Byrne et al. 2001; Goman 1996; Ingram et al. 1996a,b; Malamud-Roam and Ingram 2004; Peterson et al. 1989). The riverine inputs derive from a large watershed (153,000 km 2) encompassing the western slopes of the Sierra Nevada, the southern slopes of the southern Cascade Mountains, parts of the Klamath and Coast Ranges, and the Central Valley. In total this comprises about 40% of California (Conomos 1979; Conomos et al. 1985). The Sacramento River drains the northern part of the watershed and delivers over 85% of the total freshwater inflow to the Estuary while the San Joaquin River drains the southern watershed and delivers about 15% (Peterson et al. 1989). The two rivers merge to form the Sacramento–San Joaquin Delta, approximately 50 miles inland of the Golden Gate (Conomos et al.1985). Each year, bay-wide salinity fluctuates because of seasonal precipitation and snowmelt in Cali-

Historical Formation

 

25

fornia (Cayan and Peterson 1993; Dettinger and Cayan 2003; Knowles 2000; Peterson et al. 1989). Salinity is low during the winter-spring wet season and increases throughout the summer-fall dry season. Year-to-year variations in climate affect salinity. For example, global phenomena like El Niño–Southern Oscillation (ENSO) (Cayan and Webb 1992; Mann et al. 2000) and, on longer timescales, the multidecadal Pacific Decadal Oscillation (Benson et al. 2003; Mantua et al. 1997) directly impact precipitation patterns over the watershed (Malamud-Roam et al. 2007). These natural climate variations are further complicated by recent increases in winter-spring temperatures (Cayan et al. 2001) that result in earlier and faster snowmelt and fluctuating streamflow patterns (Dettinger and Cayan 1995; Mote 2003; Roos 1991; Stewart et al. 2005).  

Holocene Freshwater Flows and Tidal Marsh History Several reconstructions of paleosalinity in the San Francisco Bay show that while there has been a trend toward increasing salinity over the last 6,000 years due to sea level rise, there has also been considerable variability due to changes in freshwater inflows (Ingram and DePaolo 1993; Ingram et al. 1996a,b; Ingram and Sloan 1992; Schweikhardt et al. 2002). Long records of highresolution year-to-year variations in salinity are difficult to generate using sedimentary records; however, dendroclimatological analysis of blue oak (Quercus douglasii) tree-ring chronologies from the foothills surrounding the Central Valley has been used to reconstruct annual changes in salinity of the San Francisco Bay for the past 400 years (Stahle et al. 2001). Recent research has also demonstrated that shell fragments preserved in archaeological sites in the San Francisco Bay can produce seasonal records of salinity (Schweikhardt 2007). While variations in freshwater flows to the Estuary do not affect the volume of water in the San Francisco Bay, the altered salinity of the Bay water results in local changes in estuarine and adjacent wetland ecosystems (Atwater et al. 1979; Byrne et al. 2001; Goman 2001; Josselyn 1983; Malamud-Roam and Ingram 2004; May 1999; McGann et al. 2002). To appreciate these ecosystem effects, we focus on three key periods in the watershed’s climate history and the downstream tidal marsh responses. A discussion 26



of the methodologies for reconstructing paleosalinity and past marsh vegetation assemblages is beyond the scope of this chapter; we refer the reader to a recent review (Malamud-Roam et al. 2006a). The first key period, known as the Neoglacial (ca 4,000 to 3,500 years ago), occurred shortly after global sea level rise declined and tidal marshes formed along the shores of the Estuary. Several lines of evidence show that cooler, wetter conditions prevailed in California, including changes in mountain vegetation (e.g., Anderson and Smith 1994; Edlund and Byrne 1991), tree line (LaMarche 1973), fire frequency (Brunelle and Anderson 2003), and lake levels (e.g., Benson et al. 2003). These conditions led to higher Sacramento and San Joaquin River flows and lowered salinity throughout the Bay Estuary, as shown in reconstructed records from Bay sedimentary isotopic data (Ingram and DePaolo 1993; Ingram et al. 1996a,b) and marsh macrofossils and metal concentrations (Goman and Wells 2000). Reduced salinity resulted in more diverse plant assemblages with species that prefer fresher conditions, as seen, for instance, at China Camp (Goman et al. 2008). Pollen and stable carbon isotopes show the dominant taxa, Salicornia virginica and Spartina foliosa, were replaced with a mixed assemblage that included Typha spp. and Schoenoplectus spp. Further up-estuary, brackish tidal marshes (Figure 2.1) also saw a shift in plant species: at Peyton Hill, the freshwater bulrush, S.americanus, became important over other, more salt-tolerant species (Goman 2001; Goman and Wells 2000); at Browns Island, near the Delta, Phragmites communis and Typha spp. became prevalent (Goman and Wells 2000; Atwater 1980). The Neoglacial gave way to a drying trend punctuated by episodic, extreme droughts and floods (e.g., Malamud-Roam et al. 2007; Schimmelmann et al. 2003). The second key period in the history of the Estuary marshes that we discuss occurred between about AD 900 and AD 1300 (the medieval period); this was a period of unusual warmth throughout much of northern Europe (e.g., Hughes and Diaz 1994; National Research Council 2006; Osborn and Briffa 2006) and is marked by dry conditions in California and the San Francisco Bay watershed (e.g., Hughes and Graumlich 1996; Meko et al. 2001). Two distinct “mega-droughts,” each lasting a century or so (Stine 1990, 1994), occurred during

Ecology: Environment

this time. Salinity in the San Francisco Estuary was variable, with prolonged periods of higher than modern (i.e., prediversion) salinity, a result of the droughts (Ingram et al. 1996a,b; Starratt 2004). Plant assemblages on the adjacent tidal marshes shifted to a less diverse community of salt-tolerant species. For instance, at Rush Ranch, the relatively fresh-brackish marsh assemblage (including Schoenoplectus spp., Grindelia stricta, some Salicornia virginica, and Distichlis spicata), which characterized higher river flow, rapidly transitioned to a marsh dominated by Salicornia virginica (Byrne et al. 2001). Similar shifts were also experienced at Roe Island (May 1999) and down-estuary at Benicia State Park marsh located in the Carquinez Strait (Figure 2.2) (MalamudRoam and Ingram 2004). Most of the mineral sediments that maintain tidal marshes are transported into the Estuary via the Sacramento and San Joaquin Rivers. During the mega-droughts of the medieval period, flows in these rivers were so reduced (Meko et al. 2001) that the down-stream effect was not only to increase Estuary salinity but also to reduce the sediments delivered to the system. For instance, sediment accumulation rates calculated at China Camp declined during this period to 100

Valium

a

Huggett et al. 2002

0.8

Diazepam

b

Reference



where current methods to remove them are not very effective (Ternes et al. 2002, 2003). Although some pharmaceuticals undergo degradation or phototransformation (some analgesics), others appear to be more persistent (antiepileptics, lipidlowering drugs) (Tixier et al. 2003). In WWTPs, removal of pharmaceuticals depends on charge, hydrophobicity, and retention time in sludge, and it varies significantly (0%–99%) depending on the class of pharmaceuticals and level of treatment (Fent et al. 2006). In effluents, degradation

Ecology: Environment

 

Table 5.2b Chronic Toxicity of Selected Pharmaceuticals

Compound

Common name Use

Nontarget species

LOEC a (mg/L)

Reference

Fluoxetine

Prozac

SSRI b

Ceriodaphnia dubia

0.056

Brooks et al. 2003

Carbamazepine

Tegretol

Antiepileptic

Cyanobacteria

17

Diatom

10

Clofibric acid

–

Lipid-lowering agent

Rotifer

0.377

C. dubia

0.025

Zebrafish

25

Cyanobacteria

23.5

Diatom Rotifer Zebrafish Ethynyl estradiol

–

Birth control

Algae Invertebrate (reproduction) Medaka (ovotestes)

a

Ferrari et al. 2003 Ferrari et al. 2004

>100 0.246

Ferrari et al. 2003

70 0.84 (EC 50)

Halling-Sorensen et al. 1998

0.105 (EC 50) 3.0 × 10 −8

Metcalfe et al. 2001

NOEC =No Observed Effect Concentration (unless otherwise indicated)

b

SSRI = slow serotonin reuptake inhibitor

or photolysis becomes an important mechanism for removal of pharmaceuticals. Other significant sources of pharmaceuticals include agricultural discharges (hormones and antibiotics) and runoff from landfills (Bound and Voulvoulis 2004). Although the individual levels of pharmaceuticals that reach aquatic systems may be quite low, the potential for additive or synergistic effects from compounds with similar or dissimilar modes of action cannot be ignored (Cleuvers 2003; DeLorenzo and Fleming 2008). For example, Daphnia exposed to low microgram-per-liter concentrations of fluoxetine (Prozac) and clofibric acid (lipid-lowering drug) exhibited significant mortality and deformities only in response to the combination of the two drugs (Flaherty and Dodson 2005). Data on concentrations and mechanisms of action for pharmaceuticals is generally lacking, particularly in ecosystems. Ethinyl estradiol (EE2, the active ingredient of birth control pills) has

Ferrari et al. 2004

been found in many WWTPs as well as in biota. As with other estrogens, EE2 has been found to induce the production of choriogenins (egg-coat proteins) and vitellogenins (yolk proteins) in male or immature fish, potentially altering reproductive success (Arukwe and Goskyr 2003). Antidepressants such as selective serotonin reuptake inhibitors (SSRIs, including Prozac) have been detected in fish tissues at nanogram-per-gram levels (Brooks et al. 2005); however, levels that result in toxicity to fish or invertebrates have not been reported in effluents (Brooks et al. 2003). In the laboratory, SSRIs have been reported to induce spawning of zebra mussels at low concentrations (Fong 1998). Diazepam and morphine have been found to alter neurotransmitter levels and cause oxidative stress in a freshwater mussel (Gagne et al. 2010). Toxicity studies have been conducted in the laboratory for a number of pharmaceuticals (Table 5.2a,b); however, laboratory studies typically investigate the acute effects

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of pharmaceuticals and employ concentrations that are orders of magnitude higher than what is detected in the environment (Fent et al. 2006). The effect of chronic, low-level exposure to pharmaceuticals, as well as pharmaceutical interactions (mixtures), requires further investigation. Perfluorinated Compounds (PFCs) Perfluorinated compounds (PFCs) have been manufactured for over 50 years for use in stain repellents (e.g., Teflon, Scotchgard); as firefighting foam; in the manufacture of paints, adhesives, waxes, polishes, metals, electronics, and caulks; and as food-packaging coatings (Giesy and Kannan 2001; Richardson 2010). Their unusual chemistry makes them both slightly hydrophobic and lipophobic and very stable in the environment because of the carbon-fluorine bond. Perfluoronated sulfonic acids such as perfluorooctane sulfonate (PFOS) and carboxylic acids such as perfluorooctanoic acid (PFOA) are the two major PFCs. PFOS is no longer made in the United States because of widespread concern over its presence in humans and wildlife. PFOA is still being used in the United States, but its use is expected to be significantly reduced by 2015. PFCs can be found in a number of matrices, including air, water, soil, sediment, sludge, and ice caps (Higgins et al. 2005; Lau et al. 2007). Global distribution of these compounds to remote marine environments suggests that longrange atmospheric transport is occurring, similar to that observed for PCBs and PBDEs. PFCs have been found in many organisms, including invertebrates, marine mammals, fish, birds, and humans worldwide, and concentrations are generally highest in organisms living near industrialized areas (Giesy and Kannan 2001; Kannan et al. 2005; Higgins et al. 2007; Lau et al. 2007; Richardson 2010). Kannan et al. (2005) demonstrated PFC bioconcentration by invertebrates and biomagnification of PFCs by top predators in a Great Lakes food chain. PFCs strongly adsorb to solids and there is widespread occurrence of these compounds in San Francisco Bay sediments and WWTP sludge (Higgins et al. 2005). This, coupled with its persistence and ability to bioaccumulate, has prompted the RMP to recommend monitoring of PFCs in San Francisco Bay (Hoenicke et al. 2007). Information on the biological effects of PFCs 76



in the field is generally lacking. In laboratory studies, PFCs have been shown to cause developmental toxicity in rats and mice following maternal exposure to PFOS, including mortality and decreased growth (Lau et al. 2007). Other adverse effects in mammals include hepatotoxicity, carcinogenicity, and immunotoxicity. Laboratory studies of aquatic organisms have found PFCs to be moderately toxic in acute exposures and slightly toxic in chronic exposures, with effects ranging from decreased growth to behavioral alterations to mortality (Giesy et al. 2010).

Detection of Emerging Contaminants Methodologies for the detection of ECs, or their metabolites or degradation products, are required to adequately monitor ECs in water, sediment, and biota. Traditional methods for the detection of legacy organics, such as gas chromatography–mass spectrometry (GC-MS) were initially used for the detection of ECs, but several problems exist with this technology as it applies to ECs (Hoenicke et al. 2007). GC-MS is especially useful for volatile, semivolatile, and lipophilic compounds but is not as useful for detecting polar, anionic, or cationic compounds without derivatization of the sample (Giger 2009). GC-MS is often not sensitive enough to detect ECs that may be at very low levels. Although sensitivity may be improved by modifications in sample preparation, interference by sample contaminants may still be present. With the enormous numbers of chemicals being manufactured, mass spectra libraries may not contain spectra for ECs, or their metabolites and degradation products. Similarly, standards to confirm the identification of ECs, metabolites, and degradation products may not be available. Sample optimization may also be sample or matrix specific, limiting reproducibility or interlab comparisons. With the development of new or improved mass spectrometry technologies, many of these limitations have been reduced. The use of liquid chromatography–mass spectrometry (LC-MS), especially time-of-flight mass spectrometry (TOF-MS), has enabled the detection of target and nontarget analytes at high resolution (Richardson 2010). Atmospheric pressure photoionization (APPI) is also being used with LC-MS for detection of nonpolar compounds such as nanomaterials, hormones, and PBDEs. Other advanced techniques

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include the use of two-dimensional gas chromatography (GC/GC/MS) or two-dimensional liquid chromatography (LC/LC/MS), which improves separation of complex mixtures, and ultraperformance liquid chromatography (UPLC), which has improved separation and reduced sample time. Improvements in sampling and extraction include the use of extraction cartridges (such as SPE and SPME cartridges), molecularly imprinted polymers (MIPs), and polar organic chemical integrative samplers (POCIS). Detection of nanomaterials has involved a variety of microscopy techniques, including transmission electron microscopy (TEM), scanning electron microscopy (SEM), atomic force microscopy (AFM), and confocal microscopy (Farre et al. 2009). Another recent advance is the use of environmental scanning electron microscopy (ESEM), which can be used to study nanomaterials under natural conditions. Nonmicroscopy techniques for studying nanomaterials include size-exclusion chromatography (SEC), often in combination with inductively coupled plasma mass spectrometry (ICP-MS), and capillary electrophoresis. Enzyme-linked immunosorbent assays (ELISAs) have also been developed for the measurement of some hormones and pharmaceuticals (antibiotics, nonsteroidal anti-inf lammatory drugs) (Deng et al. 2003; Huet et al. 2006; Shelver et al. 2008; Barber et al. 2009). ELISAs offer a rapid, low-cost, high-throughput alternative to costlier analytical methods and are being investigated for use in detection of other pharmaceuticals.

Biomarker Responses Detection of contaminants in abiotic matrices does not necessarily imply that they will be bioavailable or cause harm to organisms. Examining biomarker responses in caged or resident biota in conjunction with analytical testing may provide useful screening tools for regulatory agencies to determine whether organisms are actually affected by contaminants and whether these impacts translate to the population or ecosystem level (Blasco and Pico 2009). A thorough discussion of biomarkers is beyond the scope of this chapter. Commonly used approaches that are not discussed here include measures of oxidative stress (antioxidant enzymes, lipid peroxidation,

protein carbonyls), stress proteins (heat shock proteins, or HSP), metallothioneins, neuromuscular function (acetylcholinesterase), and genotoxic endpoints (DNA damage, DNA adducts). The reader is referred to an excellent review by van der Oost et al. (2003) for more information on these biomarker responses as well as those described below.

Reproductive Biomarkers Reproductive biomarkers have been widely used in investigating the effects of endocrine-disrupting compounds (EDCs) on many aquatic species (Arukwe and Goskyr 2003). Vitellogenins (yolk proteins) and choriogenins (egg-coat proteins) are normally produced by the liver in reproductive females in response to endogenous estrogen, and their presence in male or immature fish indicates exposure to environmental estrogens. Ovotestis (in which gonads have both ovarian and testis architecture) and ovarian tumors are biomarkers of effect in fish exposed to xenoestrogens. The Pacific Estuarine Ecosystem Indicator Research consortium (PEEIR, part of the Estuarine and Great Lakes, or EaGLe, Environmental Indicators Program funded by EPA’s STAR program) investigated reproductive biomarkers in fish from San Francisco Bay in conjunction with tissue and sediment analysis (Anderson et al. 2006; PEEIR; Hwang et al. 2006a,b; Hwang et al. 2008). Choriogenins were detected in male fish from a contaminated marsh (Stege Marsh, San Francisco Bay), and fish from this site also had evidence of ovotestes (PEEIR). Since many ECs have been shown to have estrogenic potential, using this approach with resident or caged organisms could provide useful information on bioavailability of contaminants and whether adverse effects to organisms are occurring.

Cytochrome P450 The cytochrome P450 monooxygenase family comprises more than 200 enzymes that are divided into several subfamilies grouped according to their function in xenobiotic metabolism or synthesis and degradation of endogenous substrates (van der Oost et al. 2003). Cytochrome P450 (CYP)1A has long been used as a biomarker of exposure to planar aromatic compounds such as polycyclic aromatic hydrocarbons and PCBs.

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During Phase I metabolism of contaminants, this highly conserved enzyme biotransforms lipophilic, nonpolar contaminants into more polar compounds that can be transformed into more water-soluble compounds by Phase II enzymes. These water-soluble compounds can then be eliminated from the organism. Although many CYP1A transformations involve detoxifying contaminants, some metabolic products may be formed that are more toxic than the parent compound, resulting in DNA or protein adducts. CYP1A is quantified either by analyzing enzyme activity using the ethoxyresorufin O-deethylase (EROD) assay or by quantifying protein or gene transcription using protein electrophoresis and Western blotting, ELISA, or PCR. CYP1A expression is highly variable at the individual, gender (expression of CYP1A is down-regulated in reproductive females), and species level, so selection of appropriate organisms is critical. In addition, CYP1A expression can be reduced in organisms chronically exposed to contaminants. Other CYP enzymes, such as CYP3A, are responsible for the metabolism of most drugs (Ekins et al. 2003), so up-regulation of CYP3A can be an indication of exposure to pharmaceuticals (Christen et al. 2009).

Toxicogenomics Toxicogenomics is actively being pursued as one approach to investigating the effects of contaminants. This technique utilizes DNA microarray technology to measure gene expression in a specific organism or tissue and has been developed for a growing number of aquatic species (Venier et al. 2006; Hook 2010). One drawback to this approach is the lack of genome sequence data available for most species. Recent advances in toxicogenomics have utilized multispecies microarrays to characterize gene expression profiles, in which highly conserved gene sequences for selected genes from a number of species are used to construct the microarray (Kassahn 2008; Baker et al. 2009). Application of genomics to field studies will require that responses to contaminants be measured in specific tissues in response to single contaminant concentrations in order to evaluate responses in feral organisms. If appropriately validated, changes in gene expression may provide an early warning of specific contaminant exposure before actual damage to organisms occurs, 78



and they may also help elucidate the mechanisms by which contaminants exert their effects. However, since there is inherent variability in gene expression levels due to age, gender, or physiological state, a major challenge will be to define this variability for individuals and species (Robbens et al. 2007). A further challenge will be to interpret changes in gene expression in response to multiple stressors.

MXR Transporter Activity Multixenobiotic resistance (MXR) transporters, also known as MDR (multidrug resistance) transporters, are a first line of defense in protecting cells and organisms from the effects of contaminants (Epel et al. 2008). These transporters are capable of transporting a wide range of substrates in addition to contaminants, including endogenous compounds, naturally occurring exogenous compounds (e.g., algal toxins), and products of metabolism. Typical MXR transporter assays involve the use of a MXR transporter substrate, such as calcein AM. Transporters prevent calcein AM (which is nonfluorescent) from entering the cell cytoplasm, but if transporter activity is inhibited (using commercially available inhibitors), calcein AM is hydrolyzed within the cell to release free calcein, which is fluorescent. High fluorescence within cells indicates inhibition of transporter activity. Inhibition of transporter activity, or chemosensitization, can lead to enhanced toxicity by other contaminants, since cells are unable to effectively eliminate the contaminant. Musk compounds, perfluorinated compounds, and several pesticides have been found to act as chemosensitizers or inhibitors in this assay (Smital et al. 2004; Epel et al. 2008). MXR transporters are believed to have evolved to deal with exogenous natural toxins, and the broad substrate specificity has enabled them to act as a defense against manmade contaminants as well. Interestingly, Smital et al. (2004) showed that substances secreted by an invasive algal species in the Mediterranean also acted as MXR inhibitors. Thus this assay could serve as a useful tool for investigating invasive species as well as effects of contaminants.

Synthesis and Future Directions Stewardship of the San Francisco Estuary requires evaluation and monitoring of histori-

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cal and potential new threats to Estuary health. The San Francisco Estuary is no exception to the reality of interactions between biotic and abiotic processes, as well as the presence of complex mixtures of contaminants that are characteristic of most aquatic ecosystems. With increasing population and introduction of ever more manufactured chemicals, the San Francisco Estuary faces a significant potential for adverse effects due to chemical contamination, and the evaluation of environmental effects of emerging contaminants remains a concern of regulatory agencies, resource managers, and the public. Ideally, toxicity data should be evaluated before chemicals are put into use; however, this is rarely the case, resulting in the potential for significant adverse effects and costly remediation efforts. Currently, the investigation of potentially harmful emerging contaminants is hampered by both a lack of knowledge about mechanisms of action and adequate technologies for detecting presence, extent of distribution, and chemical fate such as degradation pathways and behavior in environmental matrices. While methodologies for detection are improving, it is unlikely that detection or quantification of the vast majority of chemicals can ever be achieved. Prioritizing which chemicals or chemical classes should be monitored based on specific criteria is the approach that has been taken by the RMP in San Francisco Bay. Generally, those criteria are similar to programs adopted by other federal and state agencies, as well as internationally, and include parameters such as persistence in the environment, bioaccumulation, toxicity to humans or wildlife, and volume of production. The ECs discussed in this chapter all fulfill these criteria to some degree, although significant gaps in knowledge are evident. PBDEs and pesticides are known to be persistent, bioaccumulate, and cause toxicity in some organisms. For these reasons, monitoring will likely continue for these classes of compounds. PFCs are persistent, bioaccumulate, and have been found in humans and wildlife. Despite a lack of knowledge regarding toxicity, it is recommended that monitoring of PFCs be a priority. Pharmaceuticals are generally believed to be at low levels in the environment; however, since pharmaceuticals are designed to target biological processes at low concentrations, an understanding of their fate in the environment is needed. The sheer number of pharmaceuticals

in use prohibits any comprehensive monitoring efforts for individual compounds. A more rational approach would be to monitor classes of compounds or representative compounds with similar effects or properties to model the extent of contamination. The widespread use and persistence of triclosan and triclocarban is of increasing concern and will likely result in monitoring efforts. Although musks are also persistent and bioaccumulate, toxicity appears to be low, so monitoring for these compounds is not a priority at this time. Finally, the projected exponential increase in the use of nanomaterials will likely engender intense interest in monitoring, not only from an ecosystem perspective but also in regard to human health. Finally, the role of local, state, and federal stakeholders in promoting wise use of resources is crucial. To this end, fostering public education on alternatives to harmful products, appropriate use of products, and disposal methods for products should be a priority. Acknowledgments The authors wish to acknowledge the Bodega Marine Laboratory, University of California–Davis, for research opportunities and support, and funding support from EPA through its Science to Achieve Results (STAR) for the Estuarine and Great Lakes (EaGLe) research program, including the Pacific Estuarine Ecosystem Indicator Research program (PEEIR) (grant no. R82867601). This chapter is a contribution of the Bodega Marine Laboratory, University of California at Davis.  

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chapter Six

Tidal Marshes in the Context of Climate Change V. Thomas Parker, John C. Callaway, Lisa M. Schile, Michael C. Vasey, and Ellen R. Herbert

Global Changes

CONTENTS Global Changes Regional Changes Estuarine and Local Changes Impacts on Individual Wetlands Synthesis and Future Directions

The principal basis for climate change is the substantial increases in concentrations of carbon dioxide and other greenhouse gasses over the last few decades (Meehl et al. 2007; Richardson et al. 2009). The sources are understood to be anthropogenic and are on trajectories to continue dramatic increases; in fact, greenhouse gas concentrations are increasing at rates faster than predicted in the last Intergovernmental Panel on Climate Change (IPCC) models (Richardson et al. 2009). The extent to which these gasses may increase in the future depends on what actions are taken to curb their production; in models, these are called assumptions, and models suggest that by 2100 carbon dioxide will increase to 400 ppm or up to near 1,000 ppm (Meehl et al. 2007; Cayan et al. 2008b; Richardson et al. 2009). Given that some scientists indicate that global climates have already been affected by current levels of greenhouse gasses (Meehl et al. 2007), clearly these increases will have substantial impacts. One prediction of all models is a global increase in average temperatures. Because temperature influences processes on a variety of scales, temperature increases will be observed as global, regional, and local impacts (Figure 6.1). Temperature increases will influence melting of large ice sheets in the Arctic and Antarctic (Rahmstorf 2007; Vermeer and Rahmstorf 2009) at global and landscape scales, earlier melting of smaller

C

limate change and sea level rise (SLR) have created the history of tidal wetlands since the last glacial maximum over 21,000 years before the present (Malamud-Roam et al. 2006; Chapter 2, this volume). In the San Francisco Bay-Delta Estuary, the result was variable but rapid SLR that lasted until about 6,000–5,000 years before present, at which time vegetation and accretion processes could keep up with the reduced rate of SLR (Atwater et al. 1979; MalamudRoam et al. 2006). Climatic variations during the last 8,000–10,000 years have resulted in distributional shifts back and forth in what we consider salt, brackish, or freshwater tidal marshes. Current predictions, however, suggest an accelerated rate of SLR, along with additional resource and climatic changes that will challenge the ability of tidal wetlands to maintain themselves in the next century. Here we consider a brief framework for how environmental conditions are modeled to shift under climate change, and the implications for tidal wetlands in the San Francisco Bay-Delta Estuary.  

 





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snowpacks at the watershed scale, and individual plant physiology at the local scale. Of critical concern for tidal wetlands is the connection between increasing temperatures and accelerating SLR. Global average SLR, which has been close to 2–3 mm/y over the last few decades (e.g., Stevenson et al. 2002; Meehl et al. 2007; Church et al. 2008), is considered a result of both thermal expansion of the oceans and melting of formerly permanent ice (Howat et al. 2007). Before 1930, rates of SLR were generally less than 1 mm/y, afterwards rising to 2 mm/y between the 1930s and the 1950s, and declining slightly during the 1960s and 1970s because of increases in global volcanic activities (Church et al. 2005). Subsequently, increases in rates of SLR to 2–3 mm/y have been reported in most parts of the world (Cazenave and Nerem 2004; Holgate and Woodworth 2004; Hughes 2004; Church and White 2006; Beckley et al. 2007; Church et al. 2008). Forecasting future rates of SLR is a difficult process. The IPCC estimated SLR of 10–59 cm by 2100 but later raised estimates to 18–79 cm (Meehl et al. 2007). The conservative IPCC rates are more than a doubling of current rates. These rates are considered to be significant underestimates because the processes involved with melting ice sheets are not yet well understood (Richardson et al. 2009; Vermeer and Rahmstorf 2009). Since their predictions, evidence suggests increased melting rates of large ice sheets (Rignot and Kanagraratnam 2006; Hanna et al. 2008; Rignot et al. 2011). Taken together, the evidence suggests that rates of SLR actually may increase much more than the IPCC predictions (Rahmstorf 2007; Vermeer and Rahmstorf 2009). Most of the analyses of global data for the Copenhagen climate meetings suggest even faster current rates of SLR (Richardson et al. 2009). The most recent model estimates suggest well more than a meter increase by 2100 is possible (Richardson et al. 2009; Vermeer and Rahmstorf 2009; Rignot et al. 2011). Analyses of Arctic and Antarctic ice sheets indicate accelerated melting, which will dominate SLR within 50–100 years (Rignot et al. 2011).  

 

 

 

 

Regional Changes Increased average temperatures will also modify regional processes, and Northern California is 88



predicted to experience significant increases in temperature. Climate change models modified to Northern California indicate that temperatures may rise significantly in the next 100 years, the extent of which, 1.5oC–7oC, depends upon the magnitude of increase in greenhouse gasses (Dettinger 2005, 2006; Cayan et al. 2008b). Greenhouse gasses are actually increasing at a rate faster than the rates assumed in these models, suggesting a disconnect between these predictions and what may actually happen (Richardson et al. 2009). These regional climate models also indicate an increase in the proportion of precipitation arriving as rain rather than snow in the Sierran watershed (Lettenamier and Gan 1990; Knowles and Cayan 2002; Miller et al. 2003; Hayhoe et al. 2004; Knowles et al. 2006; Cayan et al. 2008b). Such a precipitation shift will reduce the magnitude and duration of mountain snowpack for the watershed of the Bay-Delta Estuary (Hayhoe et al. 2004; Cayan et al. 2008b). The decline in snowpack duration with warmer spring temperatures means that the seasonality of freshwater flows into the Estuary will be modified, increasing flows in winter, with potential flooding, and reducing summer river flows entering the Estuary (Knowles and Cayan 2002; Miller et al. 2003; Hayhoe et al. 2004).  

Estuarine and Local Changes Global and regional impacts appear to be principally increased average temperatures, accelerated rates of SLR, and reduced freshwater flow into the Estuary during summers. Another global impact is that carbon dioxide is a resource for plants and may differentially influence plant growth and interactions among plants and other organisms. These regional and global influences interact with each other and with local wetland conditions to produce an array of effects. Rates of SLR in the Bay-Delta, for example, have paralleled global averages, with the exception of localized regions or time periods of high rates of subsidence, and future increases are also likely to parallel global predictions (Cayan et al. 2008a). Other large-scale processes not directly connected to climate change will complicate the process of predicting changes in rates and patterns of SLR for the San Francisco Bay-Delta. The Pacific Decadal Oscillation, for example, exhibits

Ecology: Environment

positive and negative phases; during negative phases, rates of SLR will be slowed regardless of climate change influences, while a shift to the positive phase may result in a rapidly accelerated rise (Ramp et al. 2009; Largier et al. 2010). Other large-scale processes (e.g., El Niño–Southern Oscillation) also will modify tidal elevations. Global SLR is now estimated at over 3 mm/y (Church et al. 2008) and has already affected tidal wetlands (Donnelly and Bertness 2001). The principal issue for tidal wetlands is maintaining their relative elevation with SLR, and that is accomplished with accretion of both mineral and organic matter. Accretion rates over the past several thousand years have been roughly equivalent to SLR in Bay-Delta wetlands (Patrick and DeLaune 1990; Callaway et al. unpublished data). Most of the northern and southern San Francisco Bay waters contain sufficient sediment to keep pace with SLR of up to 6 mm/y (Patrick and DeLaune 1990; Orr et al. 2003), although recent data has shown a reduction in suspended sediment concentrations within the Bay (Wright and Schoellhamer 2004). Accelerated rates of SLR may pass 6 mm/y in the next several decades, and the Estuary may lack sufficient inorganic sediment to keep pace with increased SLR estimates by 2100 (Meehl et al. 2007; Meier et al. 2007; Church et al. 2008). Organic production would have to accommodate additional increases (e.g., Langley et al. 2009), or large-scale management actions would need to be implemented to remove structures blocking sediment flow into the Sacramento and San Joaquin Rivers. Lower summer freshwater flows from the watershed will interact with accelerating SLR to shift the freshwater-marine mixing interface up the Estuary. Estuarine waters will increase in salinity during the summer growing period (Goman and Wells 2000; Byrne et al. 2001; Stahle et al. 2001; Malamud-Roam et al. 2007). Ongoing water management actions have failed to maintain historic levels of Estuary salinity and will likely have little influence on even greater future impacts due to climate change (Enright and Culberson 2010). The San Francisco region also exists in a mediterranean climate, which includes summer droughts during the growing season. Warm temperatures during the rainless summers increase salinity in pore water of high marsh areas in salt and brackish tidal wetlands. The future interac 

tion of SLR, reduced freshwater input during the summer, and increased average temperatures will increase these already high soil salinities (Meehl et al. 2007; Cayan et al. 2008b; Richardson et al. 2009). Wetland soil salinities across the Estuary already increase during the growing season to two to three times the salt concentration in adjacent estuarine waters only a few meters away from channels (see Chapter 7, this volume). Given these potential changes in soil salinity, we can expect that brackish wetlands will convert to salt marshes and that brackish water will be introduced to areas that are currently fresh.

Impacts on Individual Wetlands Climate change will result in a number of alterations to tidal wetlands. On the one hand, the physical dynamics of these wetlands will change, potentially reducing the size of wetlands or changing their physical morphology. Increases in frequency of inundation and shifts in salinity regimes likely will be reflected in shifts in plant dominance and a reduction in species richness. Because carbon dioxide is a plant resource, and carbon dioxide–induced warming will affect plant and animal metabolism and survival, we can expect differential changes in composition and dominance and alterations in ecosystem carbon and mineral dynamics from these changes alone. Among tidal wetlands, the largest shifts in plant composition and productivity in response to elevated salinity levels will occur in freshwater and brackish tidal systems. The San Francisco Bay-Delta system currently experiences seasonal and annual variation in salinity (Fox et al. 1991; Peterson et al. 1995), and historically, the brackish zones have shifted across parts of the Estuary (Atwater et al. 1979; Goman and Wells 2000; Byrne et al. 2001; Malamud-Roam et al. 2007). More recently, the brackish zones have experienced increased salinity due to water diversions upstream, with subsequent shifts in plant composition (Stahle et al. 2001; Malamud-Roam et al. 2007). Such a historical perspective leads to confidence that wetlands have some resilience in the face of global changes. But the changes that will be experienced in the future will be accompanied not only by increased salinity in the system but also by shifts in carbon dioxide concentrations, higher temperatures, SLR, and increasing water  

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Global Climate Change Drivers

CO2 and greenhouse gasses Temperature Sea level rise

Regional & watershed impacts Estuarine impacts Increased salinity Lower productivity Increased tidal inundation Patterns of freshwater flow

Shifts from snow to rain Smaller snowpack Earlier melt and spring floods Lower summer river flows

Figure 6.1.  Principal processes driving climate change–based impacts on San Francisco Bay Estuary tidal wetlands.

diversions, all within the context of a highly urbanized estuary. Freshwater tidal wetlands contain large numbers of species sensitive to low levels of salinity and will quickly lose those species as salinity encroaches into the Delta. These predictable, plus unanticipated, changes will likely cascade into adjacent terrestrial and pelagic food webs. The high marsh areas of individual salt and brackish tidal marshes will exhibit species changes first, because of escalating soil salinities in areas not flushed daily by tides in our drysummer mediterranean climate. Plant metabolic processes will be stressed by conflicting demands of increased marsh plain salinity, significant temperature increases, and accelerated rates of evapotranspiration. Salt marshes will potentially lose species in the high marsh areas that cannot tolerate these shifts in conditions, but overall there will be little difference in the upper marsh communities. Brackish wetland plant communities will shift toward plants with greater salinity tolerance, for example, Sarcocornia pacifica, Spartina foliosa, and Distichlis spicata (Mall 1969). Spartina and Distichlis are both C4 plants and may be at a competitive disequilibrium with C3 plants like Sarcocornia because increases in carbon diox90



ide favor C3 metabolism (Rasse et al. 2005). The increased carbon dioxide concentration affects other processes as well, such as carbon storage and cycling, soil nitrogen fixation, and nitrogen dynamics (Drake et al. 2003; Johnson et al. 2003; Rasse et al. 2003; Pendall et al. 2004; Hungate et al. 2005, 2006; Marsh et al. 2005; Rasse et al. 2005). Increased carbon dioxide concentration also affects the carbon composition of plants, potentially reducing herbivore attacks or their effects (Stiling et al. 2003; Cornelissen et al. 2004). If accretion rates fail to maintain elevation with SLR, individual wetlands will experience increased frequency and duration of tidal inundation (Figure 6.1). While tidal flushing may initially reduce accumulating salinity, reducing some stress, S. pacifica produces biomass at very different rates throughout the northern Bay wetland system depending on frequency and duration of inundation along a salinity gradient (Schile et al. 2011). In well-drained areas, S. pacifica seems indifferent to changes in soil salinity and produces relatively high amounts of biomass. In poorly drained areas, plants are sensitive to salinity and decrease in productivity with increases in salinity, indicating the potential

Ecology: Environment

for shifts in productivity for this dominant species. The unknowns include the role of increased carbon dioxide concentration and how this may affect above- and belowground production and how the increased carbon density of plant tissues may affect the organic component of sediment accretion, either directly or by reductions in rates of decomposition.

Synthesis and Future Directions Global climate change will affect the composition of tidal wetlands considerably. The Bay-Delta presents relatively high levels of local species endemism, particularly in the brackish marshes of San Pablo Bay and Suisun Bay and in the freshwater reaches of the lower Delta (Grewell et al. 2007; Parker et al. 2011; Callaway et al. 2012). Localized or restricted endemic species are highly susceptible to the forces of rapid climate change (Loarie et al. 2008), particularly in cases where their dispersal is likely to be constrained by impermeable barriers. Many Bay-Delta endemic flowering plant species (e.g., Cirsium hydrophilum) are undoubtedly susceptible to increases in salinity, inundation frequency, or shifts in interactions among plant species. Historically, tidal wetlands have maintained themselves against postglacial SLR by both accretion and upland migration, but the range of climate change predictions suggests considerable challenges. Tidal wetlands will begin to evolve, shifting in composition and extent, but there are many unknowns associated with this evolution. Initially, as carbon dioxide increases and summer temperatures increase, C3 plants may increase in dominance; as salinity encroaches up the Estuary, halophytes like S. pacifica will begin replacing species like Bolboschoenus maritimus and Schoenoplectus americanus. Higher temperatures, however, will also favor the C 4 plants D. spicata and S. foliosa. Simultaneously, SLR will affect these systems, initially causing slight increases in inundation regimes. Depending on conditions along the wetland-upland border, wetlands could migrate inland; but in many cases around the Bay-Delta, dikes and development restrict wetland migration, a process often referred to as coastal squeeze (Titus 1991; French 2001). Human development, ditches, and diking have resulted in many wetlands that have lost connection to mainland areas and are now small islands,

subject to increased wave erosion (Crooks 2004). Increased rates of SLR also risk the viability of restoration projects in the Bay-Delta region by reducing the window of opportunity for wetland restoration (Callaway et al. 2007). Increasing inundation frequency and duration eventually will lead to reductions in wetland productivity and rates of organic matter accumulation. It is likely that tidal wetlands could withstand moderate rates of SLR for many decades, although the predicted rates of SLR at the high end are well above rates for which sedimentation could maintain wetlands (Patrick and DeLaune 1990; Orr et al. 2003; Meier et al. 2007; Church et al. 2008; Richardson et al. 2009; Vermeer and Rahmstorf 2009). Because of the higher mineral content in salt marsh soils, they probably will be more vulnerable to increases in SLR, and as salinity moves up the Estuary, SLR impacts will also shift in that direction. If the Bay-Delta loses substantial amounts of tidal wetlands, overall estuarine productivity could decline considerably, resulting in shifts in pelagic and bird populations. As tidal wetlands evolve, responding to climate changes, ecosystem processes like carbon cycling will also be affected. While wetlands are thought to be net carbon and heavy metal sinks, they may eventually lose spatial extent, limiting sequestration. Among additional future unknowns are the responses of belowground processes that may mitigate or accelerate wetland loss. Belowground productivity increases along the estuarine gradient and may allow rates of accretion in mesohaline to freshwater tidal wetlands to parallel SLR; increases in salinity in wetlands may initially inhibit decomposition as well. Similarly, increased carbon dioxide concentrations appear to modify plant tissue in ways that reduce decomposition, and modification of plant tissues by increased carbon dioxide may rapidly increase organic accretion rates and allow tidal wetlands to accrete at rates equivalent to SLR (Langley et al. 2009). Above- and belowground production, factors influencing decomposition, and how shifts in composition will affect interactions influencing these ecosystem processes are all areas that need more focused study to improve predictive models. The unknowns in these climate change scenarios are critical to the survival of tidal wetlands. We lack full understanding of how all these processes

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will interact within tidal wetlands. Our political structures may finally take significant action in reducing the net dependence of global economies on carbon-based energy sources, allowing mitigation and long-term reduction of atmospheric greenhouse gasses. On the other hand, we may have already reached a tipping point, and wetland ecologists may be working with an ecosystem that has already accumulated an extinction debt. Acknowledgments

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The authors have received funding supporting this work from a variety of sources, but especially from the CALFED program of the Bay-Delta Authority, State of California, and from the National Institute of Climate Change Research, Coastal Center, U.S. Department of Energy.

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Estuary and Watershed Science 1(1):article 5. http:// escholarship.org/uc/item/8hj3d20t. Accessed May 20, 2007. Parker, V. T., J. C. Callaway, L. M. Schile, M. C. Vasey, and E. Herbert. 2011. Climate change and San Francisco Bay-Delta tidal wetlands. San Francisco Estuary and Watershed Science 9(3). http://escholarship.org/uc/item/8j20685w. Patrick, W. H., Jr, and R. D. Delaune. 1990. Subsidence, accretion, and sea level rise in South San Francisco Bay marshes. Limnology and Oceanography 35:1389–1395. Pendall, E., S. Bridgham, P. J. Hanson, B. Hungate, D. W. Kicklighter, D. W. Johnson, B. E. Law, Y. Luo, J. P. Megonigal, M. Olsrud, M. G. Ryan, and S. Wan. 2004. Below-ground process responses to elevated CO2 and temperature: A discussion of observations, measurement methods, and models. New Phytologist 162(2):311–322. Peterson, D. H., D. R. Cayan, J. DiLeo, M. Noble, and M. Dettinger. 1995. The role of climate in estuarine variability. American Scientist 83:58–67. Rahmstorf, S. 2007. A semi-empirical approach to projecting sea-level rise. Science 315:368–370. Ramp, S., F. Chavez, and L. Breaker. 2009. Sea level off California: Rising or falling? Central and Northern California Coastal Ocean Observing System (CENCOOS), Integrated Ocean Observing System (IOOS). http://www.cencoos.org/sections/news/ sea_level.shtml. Rasse, D. P., G. Peresta, C. J. Saunders, and B. G. Drake. 2005. Seventeen years of elevated CO2 exposure in a Chesapeake Bay wetland: Sustained but contrasting responses of plant growth and CO2 uptake. Global Change Biology 11:369–377. Rasse, D. P., J. H. Li, and B. G. Drake. 2003. Wetland sedge community has high CO2 fixation capacity under ambient and elevated CO2: Measurements and model analysis. Functional Ecology 17:222–230. Richardson, K., W. Steffen, H. J. Schellnhuber, J. Alcamo, T. Barker, D. M. Kammen, R. Leemans, D. Liverman, M. Munasinghe, B. Osman-Elasha,  

 

 

 

 

 

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N. Stern, and O. WÊver. 2009. Synthesis report. In Climate change: Global risks, challenges and decisions, Copenhagen 2009. http://climatecongress.ku .dk/. Accessed November 2010. Rignot, E., and P. Kanagraratnam. 2006. Changes in the velocity structure of the Greenland ice sheet. Science 311:986–990. Rignot, E., I. Velicogna, M. R. van den Broeke, A. Monaghan, and J. Lenaerts. 2011. Acceleration of the contribution of the Greenland and Antarctic ice sheets to sea level rise. Geophysical Research Letters 38:L05503. doi:10.1029/2011GL046583. Schile, L. M., J. C. Callaway, V. T. Parker, and M. C. Vasey. 2011. Salinity and inundation influence productivity of the halophytic plant Sarcocornia pacifica. Wetlands 31:1165–1174. Stahle, D. W., M. D. Therrell, M. K. Cleaveland, D. R. Cayan, M. D. Dettinger, and N. Knowles. 2001. Ancient blue oaks reveal human impact on San Francisco Bay salinity. Eos 82(12):141–145. Stevenson, J. C., M. S. Kearney, and E. W. Koch. 2002. Impacts of sea-level rise on tidal wetlands and shallow water habitat: A case study from Chesapeake Bay. In Fisheries in a changing climate, edited by N. A. McGinn, 23–36. Symposium 32. American Fisheries Society, Bethesda, Maryland. Stiling, P., D. C. Moon, M. D. Hunter, A. M. Rossi, G. J. Hymus, and B. G. Drake. 2003. Elevated CO2 lowers relative and absolute herbivore density across all species of a scrub oak forest. Oecologia 134:82–87. Titus, J. G. 1991. Greenhouse effect and coastal wetland policy: How Americans could abandon an area the size of Massachusetts at minimum cost. Environmental Management 15:39–58. Vermeer, M., and S. Rahmstorf. 2009. Global sea level linked to global temperature. Proceedings of the National Academy of Sciences USA 106:21527–21532. Wright, S. A., and D. H. Schoellhamer. 2004. Trends in the sediment yield of the Sacramento River, California, 1957–2001. San Francisco Estuary and Watershed Science 2:article 2. http://repositories .cdlib.org/jmie/sfews/vol2/iss2/art2.

Ecology: Environment

 

 

 

 

 

 

 

 

Part II

Ecology Organisms

chapter Seven

Tidal Vegetation Spatial and Temporal Dynamics V. Thomas Parker, John C. Callaway, Lisa M. Schile, Michael C. Vasey, and Ellen R. Herbert

contents

(Chapman 1974; Greenberg et al. 2006; Keddy 2010). Tidal wetlands are highly productive, and in addition to supporting the local food web, they provide substantial inputs to the estuarine and shallow marine detrital food web (Odum 1968; Polis and Herdt 1996; Harding 2002; Mitsch and Gosselink 2007). Tidal wetlands also provide ecosystem services for human communities, such as water quality amelioration, flood and storm abatement, protection for infrastructure, and carbon sequestration (Costanza et al. 1997; Keddy 2010). Globally, wetlands have experienced dramatic historical declines in area and hydrologic integrity (Pennings and Bertness 2001). The San Francisco Bay-Delta Estuary extends nearly 100 km inland from the Golden Gate at the Pacific Ocean to the Sacramento–San Joaquin River Delta and is the largest estuarine system on the Pacific coast of North and South America (Callaway et al. 2012). This estuary historically supported the largest area of Pacific coast salt marsh, brackish, and freshwater tidal wetlands (Macdonald and Barbour 1974) and currently has the greatest extent of these wetlands, despite those losses (Josselyn 1983). Historic tidal wetlands, restored tidal wetlands, and diked baylands with seasonal wetlands create a patchwork of critical ecosystem types, all embedded in one of the country’s largest urban areas. The floristic composition of Bay-Delta tidal wetlands is relatively poorly documented (Macdonald and Barbour

History of Tidal Wetland Origins Estuarine Tidal Wetlands in a Conceptual Context Physical Processes Affecting the San Francisco Bay-Delta Tidal Wetlands San Francisco Bay-Delta Wetland Plant Communities Low Marsh Vegetation Marsh Plain Vegetation Interannual Variability Productivity Synthesis and Future Directions

T

 

idal wetlands occupy the transitional zone between intertidal mudflats and uplands. They are dynamic systems, maintaining relative elevation with sea level rise through inputs of mineral sediment and organic matter accumulation (Day et al. 1989; Reed 1990, 2002; Kirwan and Temmerman 2009). Tidal marshes are ecologically significant; they provide habitat for resident and migratory birds and mammals, estuarine fish, and numerous invertebrates (Madon et al. 2001; Mitsch and Gosselink 2007; Keddy 2010). They often support a large number of highly specialized and endemic species, including a number of endangered species





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1974), but studies of tidal marsh vegetation have been completed for selected locations throughout the Bay-Delta (e.g., Atwater et al. 1979; Watson and Byrne 2009; Vasey et al. in press).

Regional processes

Salinity gradient Inundation patterns Watershed flows Suspended sediment supply

History of Tidal Wetland Origins As little as 10,000–12,000 years before present, at the end of the last glacial period, San Francisco Bay was a river valley that drained into the Pacific Ocean. With the melting and retreat of glaciers, combined with tectonic activity, it is estimated that tidal wetlands formed as early as 6,000 years before present within the boundaries of the current estuary (Atwater et al. 1977, 1979; MalamudRoam et al. 2006). Pollen and macrofossils in soil cores from modern wetlands confirm that tidal wetlands formed along the fringes of the bay by about 5,000 years before present (Fairbanks 1989; Malamud-Roam et al. 2006, 2007; Goman et al. 2008). Wetlands continued to expand into the 1800s, when tidal wetland area exceeded that of the bay’s open water (van Geen and Luoma 1999). Large-scale anthropogenic impacts on the estuary began in the mid-1800s. Hydraulic mining in the Sierra Nevada changed the nature of many of the wetlands by increasing sediment availability and transport by an order of magnitude (van Geen and Luoma 1999). These gold-mining deposits increased the sizes of many wetlands (van Geen and Luoma 1999; Goman et al. 2008), although the increase was temporary. Most of the freshwater and brackish tidal wetlands have been diked, and many salt marshes have been diked and filled since the 1850s (Mount 1995). Water management activities of the last century, such as dams and levee construction, channelization, and water allocation shifts, reduced suspended sediment concentrations and modified water flow patterns from the extensive Bay-Delta watershed (Mount 1995; Wright and Schoellhamer 2004; Mount and Twiss 2005; Jaffe et al. 2007). Tidal wetland area declined with continual filling and levee building through the last half of the twentieth century, when it was estimated that only 4%–8% of the original area remained (Atwater et al. 1979). Enforcement of wetland mitigation associated with the Clean Water Act and the recognition of the critical importance of wetlands reduced loss rates in the last decades of the twentieth century. Restoration of tidal wetlands within the San Francisco Bay has increased in the last  

 

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Landscape-level processes Landscape context Wetland connectivity

Site-level processes Elevation Drainage Distance from channel Inundation Salinity Competition

Figure 7.1.  Critical influential processes affect tidal wetlands from many different levels.

decade, although many large-scale restoration projects are still in the early phases of planning and implementation (Williams and Faber 2001; Callaway et al. 2007).

Estuarine Tidal Wetlands in a Conceptual Context Estuarine tidal wetlands represent a productive but highly variable environment. The processes underlying patterns of species composition within wetlands and how they shift among years arise at

Ecology: Organisms

Table 7.1 Tidal Patterns in the San Francisco Bay-Delta Estuary

Mean range (m)

Spring range (m)

Mean tide level (m)

Sacramento River

0.7–0.98

0.88–1.31

0.43–0.67

San Joaquin River

0.7–0.95

0.97–1.21

0.49–0.62

Suisun Bay

0.92–1.27

1.25–1.65

0.64–0.86

San Pablo Bay

1.17–1.63

1.57–2.09

0.83–1.13

San Francisco Bay (central-south)

1.19–2.32

1.71–2.83

0.91–1.52

Location

source: Based on NOAA tidal predictions for 2010. note: Ranges shown are minimum and maximum ranges for the area of the estuary described.

different temporal and spatial scales (Figure 7.1) (Peterson and Parker 1998). For Pacific coast tidal wetlands, variation in the phase of the Pacific Decadal Oscillation influences the rate of sea level rise, slowing it in some Pacific Decadal Oscillation phases and accelerating it in others (Ramp et al. 2009; Largier et al. 2010). Differences in precipitation and snowpack vary freshwater flow into the Estuary, altering the salinity of estuarine water as the mixing zone shifts. At smaller scales, soil salinity across an individual salt marsh varies with rainfall patterns and summer temperatures, shifting competitive dominance among species. As a consequence, tidal wetlands are a dynamic mosaic, structured by the resilient dominant perennial plants and fluctuating abiotic influences.

Physical Processes Affecting the San Francisco Bay-Delta Tidal Wetlands The hydrologic regime defines the conditions in every wetland through its control on physical and chemical soil properties, habitat access and availability, and exchange of materials with waters outside the tidal marshes (Mitsch and Gosselink 2007; Keddy 2010). As in other estuaries, salinity influences the distribution of Bay-Delta vegetation on the broadest scale, with an estuarine gradient of salt, brackish, and freshwater tidal wetlands (Figures 7.1 and 7.2); the primary drivers for patterns of vegetation distribution within par-

ticular tidal wetlands are variations in inundation regimes and their effects on both soil anoxia and salinity. Regional hydrology varies significantly across the Estuary and combines two distinct forces operating in opposite spatial directions— the highly predictable tides entering the Golden Gate and unpredictable freshwater inputs from the Sacramento and San Joaquin Rivers, as well as from local rivers and streams. The Pacific coast is characterized by mixed semidiurnal tides, compared with even semidiurnal tides or diurnal tides on North American Atlantic and Gulf coasts. Tidal range is approximately 2 meters at the Golden Gate and increases gradually across the shallow South San Francisco Bay; through San Pablo Bay, the Suisun Bay, and the upper limits of the Delta, tidal range diminishes (Table 7.1). Thus, in the upper reaches of rivers and streams, the influence of tidal patterns declines and tidal ecosystems transition to riverine wetlands. The tidal range of the Bay-Delta contrasts with other areas with a similar climate, such as southwestern or southern Australia or the Mediterranean, in which the tidal ranges are micro- or mesotidal (e.g., Nelson 1970; Adam 2009). Tidal marshes are typically found above mean tide level (MTL), with the elevation of the marsh plain variable but generally close to mean higher high water (MHHW) (Figure 7.3). The low marsh zone is flooded on every high tide cycle, keeping the salinity close to that of the adjacent water (Figure 7.4). As elevation increases, frequency and duration of tidal inundation decreases. At higher

Tidal Vegetation: Spatial and Temporal Dynamics

 

99

San Pablo Bay

Suisun Bay

Oligohaline marsh

Brackish marsh

Salt marsh

Wetland type

Pacific Ocean

San Francisco Bay

San Francisco Bay

San Pablo Bay

Figure 7.2.  Distribution of marsh types according to the salinity regime across the San Francisco Bay and western delta.

Pacific Ocean

Suisun Bay

Meters

0

5,000 10,000

Freshwater marsh

Oligohaline marsh

Brackish marsh

Salt marsh

Wetland type

Freshwater marsh

Oligohaline marsh

Brackish marsh

Maximum tide Maximum tide

Figure 7.3.  General schematic of tidal elevations and salt marsh vegetation patterns. MHHW = mean higher high water, MHW = mean high water, and MTL = mean tide level.

elevations on the marsh plain, for example, tidal flooding is at most once per day because of the mixed, semidiurnal tides, and only with relatively high tidal ranges (e.g., around spring tides). At mean high water (MHW), over a 12-month period, the marsh plain is flooded during a tide cycle on only 137 days; 10 cm lower in elevation, flooding increases to 201 days, while at 30 cm below MHW, flooding occurs on 333 days (based on National Oceanic and Atmospheric Administration tide table data for China Camp). Generally, the marsh plain is infrequently flooded, ranging from 4–12 days per month during the growing season. Concurrent with the estuarine tidal gradient is a salinity gradient (Conomos et al. 1985). Because the ocean salinity is relatively constant, the salinity gradient is strongly influenced by the flow of freshwater into the Bay. Tidal water salinity in the Estuary is usually well below ocean concentrations but varies significantly both seasonally and annually. Dilution of the Bay salinity occurs in areas where freshwater inflows are large and marine water is restricted, for example, inland of the major constriction at the Carquinez Strait (Conomos et al. 1985). Salinity thus reflects the hydrologic controls, interannual variability, and, currently, the role of human management of the system (Peterson et al. 1995; Knowles and Cayan 2002). In contrast to tidal flow, river flows are strongly seasonal and vary greatly among years. The timing and magnitude of flows are governed by rainfall, runoff, and spring snowmelt, which have strong intra- and interannual climate-driven variability. Major storms can cause temporary flooding and disturbances, and the freshwater pulses in wetlands created by storms may affect vegetation dynamics by diluting salinity during seedling establishment and opening up habitat via wrack disturbance, sediment deposits, or erosional processes. The influence of high flooding  

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and disturbances in extremely wet years is greatest on the freshwater and brackish tidal systems, with diminishing impacts toward the mouth of the Estuary. However, due to salinity dilution, these impacts can also be significant in stimulating the establishment of certain species in salt marshes (Wayne 1995). The Estuary receives considerable, but variable, summertime freshwater flow from the Sacramento–San Joaquin River watershed (Conomos 1979). Variation in rainfall amounts and summer freshwater flow is a principal characteristic of this estuarine ecosystem and has been prior to anthropogenic modifications (Ingram and DePaolo 1993; Ingram et al. 1996; Goman and Wells 2000); however, the vegetation of the Suisun Bay region has become much more saline since large-scale water diversions began (Stahle et al. 2001; Malamud-Roam et al. 2007). These flows are now highly regulated by the Central Valley Project and State Water Project (Conomos 1979; Cayan and Peterson 1993). Compared with the Sacramento and San Joaquin River flows, tributary rivers and streams generally are less regulated; therefore, their flow regimes are far more climate driven. One of these tributary rivers, the Napa River, for example, averages a major flood event every 15 years. While these broad spatial patterns of upstream salinity variation are common to estuaries regardless of location, strong seasonal salinity variation is also common in mediterranean climates, with salinities lowest in winter to early spring and highest in late summer to fall (Figure 7.4). Surface salinities of Bay water fluctuate, for example, between 0 and 18 parts per thousand (ppt) among seasons at the Carquinez Straits where the majority of mixing between marine and freshwater flows occurs (Conomos et al. 1985). Plants, however, are influenced by the pore water salinity that results from an interaction among water salinity, inundation frequency, plant uptake, and evapotranspiration. The seasonal variation in Bay waters translates to a critical influence on wetland ecology and restoration in this region because exposed sites or higher wetland elevations can become hypersaline during periods without rain, especially during the warmer summers, strongly affecting vegetation. Thus the warm temperatures and lack of summer rainfall lead to high rates of evaporation, resulting in soil salinities that are significantly higher than in estuarine waters (Figure 7.4). Soil

Ecology: Organisms

 

Average water salinity (ppt)

China Camp Petaluma River Coon Island Rush Ranch Brown Island Sand Mound Slough

c

Distance from channel (m) Figure 7.4.  Variation in salinity (in parts per thousand) across the estuary, and variation from tidal channel edges to the interior of a wetland. Values were collected during fall 2008. China Camp and the Petaluma River marsh are salt marshes, Coon Island and Rush Ranch are brackish tidal wetlands, Browns Island is an oligohaline tidal wetland, and Sand Mound Slough contains freshwater tidal wetlands.

salinity in Napa River marshes, for example, can be a few parts per thousand in early spring, rising to 20–30 ppt in summer and fall, while marshes in the western Delta can have soil salinities of 0 ppt near channels and around 5 ppt 100 m away from channel edges in the early spring, and 1–6 ppt and 8–17 ppt, respectively, in fall (Figure 7.4). In the Suisun Bay, sites in the high marsh have been recorded with seasonal shifts from 18.5 to 81 ppt (Mall 1969). Both location along the salinity gradient and climatic history control salinity at large scales. Paralleling landscape-level influences, inundation and salinity regimes within a wetland are the most important processes affecting marsh ecology, and a different collection of factors influence or modify these processes at the local scale. Elevation, tidal regime, river and storm flows, sedimentation, channel proximity, drainage isolation, and vegetation type collectively control how water moves from any point within a tidal marsh and thus define the inundation regime (frequency,  

 

 

depth, and duration of inundation). Additionally, biological processes modify these patterns and processes through feedback mechanisms. The presence of a plant canopy will strongly influence whether evaporation concentrates salts at the soil surface, facilitating or constraining what set of species can germinate and grow (Bertness and Hacker 1994). Sites that lose vegetative cover in mediterranean salt marshes may form extremely salty surface soils that limit germination and inhibit future plant establishment. This general pattern is complicated by the channel networks, with which most of the plant species in Bay-Delta salt marshes are associated (Hopkins and Parker 1983; Sanderson et al. 2000). Channel networks are critical at the local scale because they are principal controls on patterns of inundation and salinity; plant composition is strongly modified with distance from within-marsh tidal channels (Parker et al. 2011). Tidal flushing and drainage are maximized near channels, reducing soil anoxia, sulfides, and soil

Tidal Vegetation: Spatial and Temporal Dynamics

103

salinity, conditions that greatly influence vegetation. For example, the general vegetation pattern in salt marshes in the Bay-Delta is characterized by a low marsh dominated by Spartina foliosa and a high marsh dominated by Sarcocornia pacifica (Figure 7.3). Adjacent to tidal channels are a number of other species, including Grindelia stricta var. angustifolia, Jaumea carnosa, Distichlis spicata, Frankenia salina, and Limonium californicum, among others. This general intrawetland pattern of species associated differentially with channels occurs in the brackish and freshwater tidal marshes but to a lesser extent, because the density and characteristics of tidal channels change as tidal ranges decrease. The resulting vegetation heterogeneity defines the three-dimensional marsh architecture that provides habitats for birds, fish, small mammals, and terrestrial invertebrates (Weinstein and Kreeger 2000; Mitsch and Gosselink 2007). Additionally, channel networks act as the circulatory system for exchanging materials, including plant propagules, within a marsh and between adjacent marshes and the Estuary (Allen 2000; French and Reed 2001).

San Francisco Bay-Delta Wetland Plant Communities Wetland plant communities are typically characterized by patterns that reflect gradients in elevation, salinity, inundation, changes in substrate chemistry, and competitive and positive interactions (Chapman 1974; Bertness 1992; Pennings and Bertness 2001; Grewell et al. 2007). Across the Estuary, large-scale distribution of wetland types indicates their approximate position along the mixing gradient of ocean water with river water (Krone 1979). Tidal marshes nearest the Golden Gate are saline; those within the Delta are freshwater (except during extreme drought conditions that can affect the western Delta); and those in between, such as in Suisun Bay, the Napa River, and the Petaluma River, span the brackish range, from polyhaline (18–30 ppt) and mesohaline (5–18 ppt) sites to oligohaline wetlands (0.5–5 ppt) (Figure 7.2). Tidal wetlands in the salt and brackish wetland regions are characterized by a mixture of species with varying salt tolerances. In brackish tidal wetlands, for example, plant assemblages on the upper marsh plain are similar to those in salt marshes, while channel vegetation reflects  

 

104



 

oligohaline communities farther upriver. The dominance of more salt-tolerant species in the high marsh reflects the lack of significant summer rainfall and high evapotranspiration rates, processes that concentrate soil salts through the summer and fall on marsh uplands experiencing tidal inundation infrequently. Although similar in pattern to most temperate zone tidal salt and brackish marshes, the Bay high marsh is higher in salinity and dominated by species with high salt tolerance. In the low marsh, the daily tidal flushing prevents solute concentration, resulting in relatively lower salinities. Plant diversity in Bay-Delta tidal wetlands varies greatly along the salinity gradient. Plant diversity ranges from fewer than 20 species in Bay-Delta tidal salt marshes to between 60 and 100 species in tidal freshwater marshes (Atwater et al. 1979; Parker et al. unpublished data) (see Table 7.2 for common species). The tidal plant communities have many species in common with tidal wetlands in other regions (e.g., Sarcocornia pacifica, Distichlis spicata, Schoenoplectus americanus, Bolboschoenus maritimus, Jaumea carnosa, Triglochin maritima). What distinguishes these plant communities from other temperate North American tidal wetlands is the increased dominance of succulent species (Josselyn 1983; Zedler et al. 1992; Callaway et al. 2012) and significant local endemism, especially in the brackish wetlands. This most likely arises from California’s mediterranean climate, with warm, dry summers and mild, rainy winters that result in the distinctive patterns of salinity discussed above (Josselyn 1983).

Low Marsh Vegetation Below mean high water, the lowest-elevation areas of San Francisco Bay tidal marshes are inhabited by only a few species that are tolerant of prolonged inundation depth and duration. These species may be found along creek banks and gradual sloping areas adjacent to mudflats, although some marshes have steep interfaces and lack low marsh vegetation at the Bay margin. Spartina foliosa dominates low marsh areas in most of San Francisco and San Pablo Bay salt marshes, often in monocultures. S. foliosa is endemic to the California Floristic Province and is found from northern Baja California to just north of the entrance to the San Francisco

Ecology: Organisms

Table 7.2 Common Species in the San Francisco Bay-Delta Tidal Wetlands, Indicated by General Wetland Salinity and Elevation within the Wetland

Salt Species

Brackish

High

Low

High

Sarcocornia pacificaa

+++

+

++

Distichlis spicata

b

++

Jaumaea carnosaa

++

+

++

++

+

+

Cuscuta salina (+C. subinclusa)

a

Grindelia stricta var. angustifola Spartina foliosa

a

Bolboschoenus maritimus a Schoenoplectus americanus a

++ +

+

+ ++

++

++

+

++

++

++

+

+++

+

+++

S. californicus 

++

S. acutus 

a

++

Typha speciesa Phragmites australis 

a

Low

+

a

Calystegia sepiuma

High

+

+++

b

Low

Fresh

+++ ++

+

+++

+

+

++

Plants with C3 photosynthesis

a 

Plants with C4 photosynthesis

b 

Bay-Delta (Vasey 2010). An introduced species from the Atlantic and Gulf coasts, Spartina alter­ niflora, has become established in the San Francisco Bay-Delta Estuary (Callaway and Josselyn 1992). Because S. foliosa is phylogenetically a sister taxon to S. alterniflora (Baumel et al. 2002; Aı ̈nouche et al. 2003), hybrids and recombinants formed from crosses between S. foliosa and the introduced S. alterniflora within San Francisco Bay. These recombinants have rapidly spread and are modifying South San Francisco Bay salt marsh habitats (Daehler and Strong 1994, 1996, 1997; Ayres et al. 2004). A large-scale effort to remove invasive Spartina species and hydrids through herbicide spraying is ongoing (see the San Francisco Estuary Invasive Spartina Project, http://spartina.org/). As salinity declines, other species become prominent in the low marsh, eventually displacing S. foliosa. Schoenoplectus acutus and S. californicus become the dominant species in brackish and freshwater marshes. Bolboschoenus maritimus and Typha angustifolia may coexist with S. acutus and S. californicus at some sites, depending on

salinity and site history. At the freshwater end of the estuarine gradient, other species such as Typha spp., Phragmites australis, Sparganium eurycarpum, and Persicaria maculata may also be found in this tidal zone.

Marsh Plain Vegetation The large marsh plains of California salt marshes are dominated by Sarcocornia pacifica, the most common plant in California tidal wetlands (Zedler et al. 1999). Salt marsh dodder, Cuscuta salina, co-occurs as a parasite on S. pacifica and infrequently on other species. Diversity in the marsh plain is found along the land interface or adjacent to tidal channels (Hopkins and Parker 1983; Zedler et al. 1999; Sanderson et al. 2000). Species include Distichlis spicata, Jaumea carnosa, Frankenia salina, Limonium californicum, and Atriplex triangularis. A San Francisco Bay endemic, the subshrub Grindelia stricta var. angustifolia, is also present near channels in San Francisco Bay and provides habitat for many endemic bird species. Disturbed areas and early

Tidal Vegetation: Spatial and Temporal Dynamics

105

successional stages of restored marshes may contain a number of annual species as well, including Salicornia bigelovii and S. depressa, as well as A. triangularis. Baye et al. (2000) note that other rare native species (e.g., Chloropyron maritimum) once were more common in transition areas at the upland interface of salt marshes, particularly where freshwater sources encountered these marshes. Bolboschoenus maritimus and Typha angustifolia are dominant on wetland plains above mean high water in meso-brackish regions, especially up the Napa River and in the western portion of the Suisun Bay. At lower salinity levels in brackish marshes in San Pablo Bay, Schoenoplectus americanus is abundant on the marsh plain and displaces B. maritimus. Diversity increases in these tidal wetlands wherever estuarine freshwater can prevent hypersaline salinities from developing during the rainless summers. Additional species found in the marsh plains in less saline conditions include Typha spp., Chloropyron maritimum, C. molle, Triglochin maritima, T. concinnum, Juncus leseurii, J. balticus, A. triangularis, and Potentilla anserina var. pacifica. Once in the freshwater tidal wetlands, the marsh plains are dominated by Schoenoplectus acutus, with S. americanus, Phragmites australis, Typha spp., Persicaria spp., and the fern Athyrium felix-femina as codominants in some areas. Large numbers of additional species are found, varying in abundance depending on local hydrology and plant density, and include Stachys albens, Mentha arvensa, Lycopus asper, L. americanus, Oenanthe sarmentosa, Glaux maritima, several species of Juncus, Eleocharis, and Carex, and large numbers of annuals (Leck et al. 2009). Lepidium latifolium is an invasive species that has a broad salinity range and has invaded most of the Estuary, from brackish to freshwater wetlands. Currently, removal and control operations for L. latifolium are being conducted on a limited basis. San Francisco Bay-Delta wetlands can also have a significant woody component. In salt and brackish marshes, this is usually limited to Grindelia stricta var. angustifolia, which is found on high ground, either associated with natural levees adjacent to tidal creeks or at the wetlandupland transition. As soil salinity declines, Baccharis pilularis, more often a dominant of coastal scrub, can be found at the wetland-upland interface. Freshwater tidal wetlands also have several 106



woody plants such as Salix spp., Cornus sericea ssp. occidentalis, and Cephalanthus occidentalis var. californicus.

Interannual Variability Seasonal and interannual variability in salinity (Figures 7.1 and 7.4) appears to strongly influence the composition of plant communities. Within brackish tidal wetlands, seasonal patterns in salinity have a prominent effect on wetland vegetation structure. Soils in the higher marsh plains accumulate salts through the summer and fall, resulting in vegetation composition that overlaps with more saline locations; similarly, the vegetation associated with channels appears to overlap compositionally with sites in fresher locations upestuary, reflecting the lower salinity of the tidal channel waters, especially in the spring. The effects of interannual variability can be seen in larger-scale vegetation shifts associated with droughts. For example, droughts in the late 1970s shifted vegetation up-estuary in the oligohaline western edge of the Delta or resulted in intramarsh distribution shifts (Atwater et al. 1979). Atwater et al. (1979) documented a decrease in both the abundance and the height of Schoenoplectus americanus, S. californicus, S. acutus, and Bolboschoenus maritimus near the Carquinez Strait during the drought and subsequent invasion of these areas by Sarcocornia pacifica. Salinities within the Bay near the eastern end of Carquinez Strait increased from below 10 ppt to 15–20 ppt during this period. Collins and Foin (1992) noted the spread of Spartina foliosa upstream in the North Bay during drought periods. Similar patterns of vegetation shifts have been found in cores taken from the oligohaline Browns Island, where wetland vegetation has shifted toward more saline communities and back to freshwater wetlands on a scale of hundreds to thousands of years (Malamud-Roam et al. 2006). Extreme wet years, such as those associated with El Niño, can also shift vegetation distribution and patterns. Low marshes in salt marsh areas are typically dominated by only Spartina foliosa in the Bay region. However, another species that might be found in the low marsh zone is Bolboschoenus maritimus. This species is more commonly found in brackish wetlands but can get established at the transition from S. foliosa low marsh to high marsh in extreme wet years. Once

Ecology: Organisms

 

2000 Species diversity ANPP

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4 1000

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Average spring water salinity (psu) Figure 7.5.  Aboveground net primary productivity (ANPP), in grams per square meter, and average species diversity in Bay-delta tidal wetlands from 2003 and 2004. Average species diversity was calculated from 136–448 circular plots of 7 m2 randomly located throughout each site. Productivity data are based on values from the dominant species at each site. Salinity is in practical salinity units (psu).

established, these stands tend to persist, varying in cover and extent depending on patterns of rainfall and salinity. Other shifts observed in wet years are often more temporary and result from species with high seed production. Atriplex triangularis and Grindelia stricta var. angustifolia expand their populations dramatically in wet years and can be found in locations where longterm survival is tentative. Following the El Niño year of the late 1990s, for example, Grindelia populations doubled at the China Camp salt marsh and even colonized edges of pannes. Individuals that established away from channels died out during the next few growing seasons.

Productivity Large-scale spatial patterns of salinity are also correlated with patterns of tidal wetland plant productivity that varies significantly across the Bay-Delta (Figure 7.5). Productivity is lowest in

tidal salt marshes, with two- to fivefold increases in brackish and freshwater tidal marshes. Within San Francisco Bay, Mahall and Park (1976) measured salt marsh annual productivity of 270–690 g m-2 y-1 for Spartina foliosa and 550–960 g m-2 y-1 for Sarcocornia pacifica, while Atwater (1979) measured peak biomass in salt marshes of 300– 1,700 g m-2 for S. foliosa and 500–1,200 g m-2 for S. pacifica. Schile et al. (2011) found that the combined effects of salinity and poor drainage negatively affect the productivity of S. pacifica; however, productivity was not affected by salinity alone. In fresh and brackish marshes peak biomass for Schoenoplectus californicus was approximately 2,500 g m-2 (Atwater et al. 1979). We found a similar trend of reduced biomass with increasing salinity in a survey of six tidal marshes from the Petaluma River to the western Delta (unpublished data); end-of-year biomass was 300–600 g m-2 in salt marshes and 800–2,400 g m-2 in brackish and freshwater marshes.  

 

 

 

 

 

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Synthesis and Future Directions

Acknowledgments

Although tidal wetlands within the San Francisco Bay-Delta formed during the last 5,000– 6,000 years, over 90% have been filled, diked, or severely degraded over the past century. Despite these massive losses, representative tidal wetland plant communities still persist within the San Francisco Bay-Delta, including a mix of salt, brackish, and freshwater tidal wetlands. The tidal wetland vegetation within the Bay-Delta Estuary is diverse, and this is especially true for the brackish and freshwater tidal wetlands. As pointed out in Vasey et al. (in press), there has been a tendency to not include Delta plant assemblages in past assessments of the San Francisco Bay-Delta vegetation (Josselyn 1983; Baye et al. 2000; Grewell et al. 2007). Our studies, consistent with those of Atwater et al. 1979, have confirmed that freshwater Delta tidal wetlands are some of the most diverse and imperiled of all the vegetation in the Bay-Delta Estuary. Recent and proposed efforts to restore wetlands within the Bay-Delta could substantially increase the extent of these important ecosystems (see Chapter 18, this volume). San Francisco Bay-Delta tidal wetlands represent the most intact mediterranean-climate wetlands in the United States (Fetscher et al. 2010). The degree of diversity and species endemism represented in these systems is largely due to the unique climate regime under which they evolved. The greatest threat to tidal wetlands over the next century will involve the consequences of global climate change, with predictions for higher temperatures, changing precipitation patterns, increasing rates of sea level rise, and altered salinity distributions across the Estuary that will have major effects on vegetation distributions and long-term survival of tidal marsh communities (see Chapter 6, this volume). Increased human population densities and development intensity in the San Francisco Bay region will compound these climate change impacts, as will ongoing alterations of freshwater flows for agriculture and other human use. As we face these challenges, it is critical that we take a holistic approach to sustaining the dynamic processes that have shaped the San Francisco Bay-Delta Estuary over the millennia and dedicate large landscapes representative of its diverse regions for restoration and recovery so that it will remain resilient to these changes over time.  

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The authors have received funding supporting this work from a variety of sources, but especially from the CALFED program of the Bay-Delta Authority, State of California, and from the National Institute of Climate Change Research, Coastal Center, U.S. Department of Energy.

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chapter Eight

Tidal Wetland Vegetation and Ecotone Profiles The Rush Ranch Open Space Preserve Christine R. Whitcraft, Brenda J. Grewell, and Peter Baye

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contents

he Rush Ranch Open Space Preserve (Rush Ranch) is located at the northwestern edge of the Potrero Hills and includes the largest remaining undiked tidal wetland within the Suisun Marsh region of the San Francisco Estuary. The brackish tidal wetlands grade into transitional vegetation and undeveloped grasslands of the Potrero Hills, and we describe diverse vegetation that reflects the estuarine position, land use history, and hydrogeomorphic complexity of the site. A useful framework for future study of vegetation at this San Francisco Bay National Estuarine Research Reserve site is presented. Rush Ranch includes four major estuarine geomorphic units that are widely distributed in the region and support vegetation: subtidal channel beds, fringing tidal marsh, tidal marsh plain, and tidal marshterrestrial ecotone. These are distinguished by small variations in hydrology and elevation, as noted and described through field observations and historic vegetation-mapping data. We discuss vegetation within each of these landforms, considering each vegetation community as a function of changing physical environment and biological iterations. Past land use and exotic plant species invasions have substantially altered Rush Ranch tidal marsh vegetation patterns. Our results indicate 27% of the current estuarineassociated flora at Rush Ranch are exotic species (Baye and Grewell 2011), and several are highly invasive. Despite these influences, Rush Ranch’s position in the landscape provides important and

Site and Setting Geomorphic and Historic Development of Vegetation Paleoecology and Historical Ecology Early Anthropogenic Influences on Estuarine Vegetation Exotic Plant Introductions Importance of Vegetation Presence and Type Contemporary Vegetation Patterns Relative to Hydrogeomorphic Settings Tidal Channels Fringing Tidal Marsh Tidal Marsh Plains Tidal Creek Banks and Natural Levees Artificial Channels and Levees Tidally Drained Marsh Plain Poorly Drained Marsh Plain Marsh Plain Ponds High Marsh Pans (Turf Pans) Freshwater Seepage Sites Seasonal Wetlands Diked Marsh Tidal Wetland: Terrestrial Ecotones Lowland Grassland (Sedge Rush Meadow) Riparian Bluff Active Alluvial Flats Modern Transformations: Climate Change Synthesis and Future Directions



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increasingly rare habitat linkages between the tidal marsh and upland grasslands, providing great potential for restoration and enhancement. We present a detailed vegetation analysis by hydrogeomorphic setting to provide an ecological framework for future monitoring, research, and adaptive conservation management at Rush Ranch.

Site and Setting The 425 hectares of estuarine wetlands at Rush Ranch, a component site of the San Francisco Bay National Estuarine Research Reserve, are part of the largest extant tidal marsh within the brackish Suisun Marsh reach of the San Francisco Estuary. The tidal wetland at Rush Ranch is unique because of its areal extent, largely intact prehistoric marsh platform, hydrogeomorphic complexity, continuity between tidal marsh ecotones and undeveloped grasslands, and habitat provision for endangered and endemic plant populations. Hydrology and geomorphology are fundamental determinants of the structure, dynamics, and productivity of wetland plant communities. The estuarine vegetation at Rush Ranch reflects hydrological influences on different spatial and temporal scales: (1) regional scale—location in the estuary, (2) temporal scale—historic land use, and (3) local scale—modern patterns of site-specific hydrogeomorphology. Rush Ranch is approximately 80 km up-estuary from the Golden Gate tidal inlet in the northern region of Suisun Marsh. Suisun Marsh is situated between the extensive Sacramento–San Joaquin Delta and the North and South Bay reaches of the San Francisco Estuary. In this region, the hydrology and tidal mixing of fresh and salt water have been spatially and temporally dynamic, and historic variability in physical processes was a key driver of historic biological diversity (Moyle et al. 2010). At the regional scale, the diversity of vegetation within the entire Suisun Marsh and particularly at Rush Ranch results from a combination of, and small variations in, physical and geological factors, such as distance from the ocean; the magnitude of freshwater input from direct precipitation, watershed runoff, and Delta outflow; salinity pulses; storms; and the duration of tidal submergence (Atwater et al. 1979; Josselyn 1983). In contrast to wetlands in the Delta, vegetation at Rush Ranch is influenced by large annual and  

 

 

 

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interannual ranges in salinity (Moyle et al. 2010). In recent times, key physical and chemical processes have been anthropogenically mediated by active management of Delta outflow and Suisun Marsh salinity regimes by state and federal water projects (Enright and Culberson 2009) that effectively reduce biologically important environmental variation (Moyle et al. 2010). The site includes a rich estuarine flora that corresponds to unique hydrogeomorphic features within the marsh and supports estuarydependent wildlife and complex food webs. This vegetation is quite different from graminoiddominated tidal marshes of the North American Atlantic coast, and the brackish plant community composition and structure is also floristically distinct from, and more diverse than, tidal marsh vegetation in San Francisco Bay or along the outer coast of California (Mason 1972; Baye et al. 2000; Grewell et al. 2007; Watson and Byrne 2009). While the tidal wetlands and terrestrialecotone vegetation at Rush Ranch are unique and largely intact as compared with most of the San Francisco Estuary, they have not been immune to change and do not represent a static, predevelopment condition. They have a legacy of agricultural and ranching use, hydrologic modifications, and alteration to vegetation (Whitcraft et al. 2011). As a result of this anthropogenic activity, the flora of terrestrial ecotones between estuarine marshes and uplands has been significantly degraded, and native flora from these areas is now regionally rare or extirpated (Baye et al. 2000). Despite historic alterations, the estuarine plant communities at Rush Ranch are a significant natural resource that merit conservation attention. Rush Ranch has one of the only remaining gently sloping undeveloped lowlands (alluvial fan topography and soils lacking intensive agriculture or urban/ industrial development) bordering undiked tidal marsh. This setting provides rare geomorphic accommodation space for estuarine transgression as sea level rises and a rare opportunity to conserve the high tidal marsh and its terrestrial ecotone. Aspects of the tidal wetland vegetation of Suisun Marsh previously have been reviewed by Mason (1972), Atwater and Hedel (1976), Atwater et al. (1979), Josselyn (1983), Wells and Goman (1995), Baye et al. (2000), Byrne et al. (2001), Hickson et al. (2001), Grewell et al. (2007), Watson and Byrne (2009), Vasey et al. (2011), and

Ecology: Organisms

Whitcraft et al. (2011). Floristic surveys and studies of plant ecology specific to Rush Ranch and contiguous tidal wetlands off-site have also contributed to our knowledge of the site (Wetland Research Associates 1990; Siegel 1993; Ruygt 1994; Grewell 1996; Fiedler and Keever 2003; Grewell et al. 2003; Fiedler et al. 2007; Grewell 2008a; Watson and Byrne 2009; Reynolds and Boyer 2010). These surveys and studies suggest the modern vegetation at Rush Ranch is typical of relict tidal wetlands elsewhere in the Suisun Marsh, and while rare species and plant assemblages occur, the Rush Ranch flora shares many elements of wetland flora with marshes elsewhere in the San Francisco Estuary and northern coastal California. This profile presents historical context, descriptions and baseline data on the floristic composition and ecology of estuarine vegetation at Rush Ranch, and the relationship of vegetation to hydrogeomorphic settings and associated hydrologic and other physical-chemical processes, as well as modern transformations of vegetation pattern. Baye and Grewell (2011) provide detailed information on estuarine flora observed at the site that includes 123 currently known taxa from 34 plant families. In addition, we provide a framework for future monitoring, research, and adaptive conservation management.

Geomorphic and Historic Development of Vegetation Paleoecology and Historical Ecology In addition to spatial scale differences in hydrology, past land use exerts great influence on hydrology and thus on present plant community structure. Paleoecological reconstructions of geology, climate, sedimentation, and vegetation change of the northern San Francisco Bay Estuary during the past 7,000 years have included site-specific studies of tidal wetlands ringing the Potrero Hills (including Rush Ranch) that are the ecological heritage of modern vegetation. The oldest tidal brackish and salt marsh sediments in the northern San Francisco Bay Estuary are associated with a slowing of postglacial sea level rise rates as modern sea level was approached. This initial deceleration of sea level rise began 6,000 years before present (BP), and by approximately 4,000 years BP initiation of most modern tidal marsh plains began, although some emer-

gent fresh-brackish estuarine marshes deposited discontinuously earlier (Wells et al. 1997; Malamud-Roam and Ingram 2004; Malamud-Roam et al. 2007). Studies of the stratigraphic record of microfossils (pollen, diatoms, foraminifera) and organic and inorganic sediments at Rush Ranch indicate that the wetland vegetation at Rush Ranch developed and has been subjected to fluctuations in environmental variability over millennial and centennial scales, as well as climate-driven changes in hydrology and aqueous salinity (Wells et al. 1997; Malamud-Roam and Ingram 2004; Malamud-Roam et al. 2007). These climate variations occurred in a background of relatively slow and stable sea level rise rates. They corresponded with marked fluctuations in the composition of tidal marsh dominant vegetation, indicated by reversals in relative abundance and composition of pollen assemblages corresponding with low- and high-salinity regimes (Watson and Byrne 2009). Empirical reconstruction of Rush Ranch paleoecology clearly indicates that the existing mature marsh plain and sloughs, and corresponding development and evolution of marsh plant communities, have a relatively brief geologic existence—less than 2,000 years—and underwent profound fluctuations in vegetation dominance and salinity regimes, as well as in precipitation (Byrne et al. 2001). The stratigraphic and pollen records do not support the assumption of an “equilibrium” or steady “natural” state in either Suisun Marsh or Rush Ranch (Byrne et al. 2001; Goman et al. 2008). These records have specific implications for special-status species conservation, particularly endemic Suisun Marsh species. Suisun Marsh historic endemic species, some of which are now endemic to Rush Ranch alone or nearly so, either persisted in refugial habitats within local salinity gradients of Suisun Marsh or underwent range shifts very rapidly between Suisun Marsh, the western estuary, and the Delta. Stable suitable habitat likely did not persist at any one location at Rush Ranch for more than 1,000 years.  

 

Early Anthropogenic Influences on Estuarine Vegetation Rush Ranch is located near some of the largest prehistoric Patwin (Wintun) village sites recorded in the Suisun Marsh region (Kroeber 1925; John-

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son 1978; Fulgham Archaeological Resource Service 1990). Patwin and California Indians inhabiting the Estuary’s margins utilized annual burning of grasslands and after-seed harvest in the lowland valleys for hunting, maintenance of favorable seed (pinole), and bulb production (Bean and Lawton 1973; Lewis 1973; Johnson 1978; Lightfoot and Parrish 2009). Annual burns likely influenced the character of tidal marsh edges and stream valleys, particularly in limiting the development of woody scrub. In addition to burning activities, digging, stem cutting, and burning of tule stands to enhance growth also likely altered productivity and the ecology of tidal marsh vegetation in California (Anderson 2005). Schoenoplectus acutus (hardstem bulrush), S. californicus (California bulrush), Juncus species, Carex barbarae (basket sedge), and other sedge beds were harvested for textiles (house construction, reed boats, clothing, footwear, duck decoys, and basketry) as well as food (Johnson 1978; Anderson 2005). In the 1700s, Spanish explorers also introduced both nonnative plants and fire into the system. All of these activities potentially influenced the structure of tidal marsh vegetation at Rush Ranch. Photographs and other records from the late nineteenth and early twentieth centuries suggest that historic anthropogenic influences on Rush Ranch tidal marshes include regional and local diking (e.g., Suisun Slough and partial diking within marsh plain, Second Mallard Branch drainage), ditching to drain tidal marsh plains (mosquito ditching), haying and livestock grazing in tidal marsh, creation of tidal marsh pans and ponds, construction of slough dams and partial levees along marsh perimeters, and introduction of nonnative animal and plant species. These alterations contributed to indirect ecological alterations such as increased terrigenous sedimentation from gullies and seasonal streams and slope failures of adjacent hillslopes, both subject to overgrazing. Diking of historic tidal marsh in the Suisun region progressed from the late 1870s through the 1970s. The construction of full and incomplete dikes at Rush Ranch along slough borders of tidal marshes likely contributed significantly to local declines in tidal slough bank vegetation (including rare endemic plants) that was regionally decimated by early twentieth-century diking. Diking and ditching, and cattle manure in the 116



tidal marsh, also likely facilitated the spread of invasive nonnative species into the marsh. Diking of historic tidal marsh has greatly affected estuarine ecotone transitions in the San Francisco Estuary by creating sharp boundaries between wetlands and terrestrial grasslands (Josselyn 1983; Fieldler and Zebell 1995). Mason (1972) and George et al. (1965) report accounts from “old timers” that prior to diking of wetlands in Suisun Marsh, there were extensive tidal marshes “where water stood on the land,” and tall tules lined the margins in deeper water. This pattern of vegetation was also reported by DeAnza as he first explored Suisun by water in 1776. Historical reports also note that high marsh plains on Grizzly Island were covered with Distichlis spicata, which was dominant but associated with salt-tolerant species, including Sarcocornia pacifica (syn. Salicornia virginica, Salicornia pacifica, perennial pickleweed), in poorly drained areas (George et al. 1965; Mason 1972). In the nineteenth century, D. spicata (salt grass) and Schoenoplectus americanus (chairmaker’s bulrush) were both harvested as commodities and utilized as packing material by the Gladdin McBean Pottery Works in Lincoln (Frost, not dated). Salt grass hay bales were also loaded onto schooners at sites such as Rush Landing and transported for sale as cattle feed (Mason 1957, 1972; George et al. 1965; and Frost, not dated). Haying directly in tidal marshes also likely had acute and prolonged inhibitory effects on reproduction of what are now rare endemic high tidal marsh plants. Grazing most likely had impacts similar to those of haying. Grazing in marshes would likely have been most intensive in early summer, when hillslopes are dry and green forage is restricted to wetlands. Intensive grazing likely occurred during peak f lowering periods of Cirsium hydrophilum var. hydrophilum (Suisun thistle) and Chloropyron molle ssp. molle (soft bird’s beak), for example. Cattle grazing has been officially excluded from tidal wetland areas of Rush Ranch since the Open Space Preserve was established (Wetland Resource Associates 1990), but in the 1980s, prior to transfer of ownership to Solano Land Trust (SLT), grazing within the tidal wetland was pervasive (P. Moyle, personal observation). Following removal of cattle from the marsh, the population of endangered C. h. var. hydrophilum (presumed extinct) recruited and

Ecology: Organisms

spread along tidal creek channels (B.J. Grewell, personal observation). Since SLT ownership, there has been both unintentional and intentional grazing of cattle within tidal marsh areas; the practice has recently (2011) been reestablished (Poerner of SLT, personal communication). Cattle grazing has directly affected endangered plant populations in the tidal wetlands, and resultant trampling has destroyed a historic population of endangered C. molle on the marsh (Grewell et al. 2003; Grewell 2005). Historic hunting influences on Rush Ranch and surrounding private hunting clubs and public wildlife areas also influenced tidal wetland vegetation within Rush Ranch. Historically, much of the abundant native vegetation (e.g., Sarcocornia pacifica and Distichlis spicata) in Suisun Marsh was considered “undesirable” for waterfowl (Rollins 1981), and early management of diked wetlands focused on production of nonnative plants and some native species—particularly Bolboschoenus maritimus (alkali bulrush), Scirpus robustus misapplied—that were not naturally dominant in the region (Miller et al. 1975). Several plant species or novel genotypes of local species were introduced by duck clubs in California, who primarily purchased seed for waterfowl habitat from eastern and southern U.S. sources (Mason 1957). Releases of exotic ring-necked pheasants and other game birds on adjacent hunting lands may explain the high density of pheasants at Rush Ranch where they are protected from hunting. Pheasants rely on plant seeds and insect food sources, and their foraging effects on Rush Ranch vegetation and native wildlife food webs are unknown. Feral pigs (Sus scrofa), relative of the European boar, are nonindigenous to North America and were introduced for hunting; in recent years, they have invaded Rush Ranch tidal wetlands. Their impacts on tidal marsh vegetation are quite visible, but ecological effects have not been studied at Rush Ranch. Rooting and wallowing activities of feral pigs are a major source of unnatural disturbance in the marshlands. For example, large sections of D. spicata– dominated areas of marsh plains have been especially affected at the preserve (authors’ personal observation). Habitat destruction by feral pigs is a major threat to the long- and short-term viability of endangered soft bird’s beak (Grewell et al. 2003) and endangered Suisun thistle (Fiedler et al. 2007) at Rush Ranch.  

 

 

Exotic Plant Introductions The introduction of exotic plant species and their subsequent spread and colonization as invasive weeds has degraded tidal wetlands of the San Francisco Estuary, and Rush Ranch has not been excluded from this impact. Interactions between exotic and native species alter the structure and function of wetland plant communities, profoundly affect the diversity and abundance of native flora, and pose significant challenges to the integrity and sustainability of current and proposed wetland restoration projects. At Rush Ranch, Lepidium latifolium, Apium graveolens, and a suite of winter annual grasses—Hainardia cylindrica (barbgrass), Parapholis incurva (sicklegrass), Polypogon monspeliensis (rabbitsfoot grass, annual beard grass)—have been particularly problematic and directly affect endangered native flora (Grewell et al. 2003; Grewell 2005; Fiedler et al. 2007; Grewell et al. 2007). Tidal wetland restoration sites are highly susceptible to weed invasion due to hydrochorous dispersal of weed diaspores and the disturbed condition of newly restored sites and also because the implementation of restoration projects is proceeding prior to regional eradication of weeds to manageable levels. Exotic, invasive plant species of particular concern at Rush Ranch are discussed below. Additional exotic plants with potential for increased spread and impact are listed in Baye and Grewell 2011.  

 

Lepidium latifolium L. latifolium (perennial pepperweed, white top) was first discovered in California in 1936 (Robbins 1941). L. latifolium began rapidly and aggressively expanding its range from 1986 to 1996 as water management and land use practices in the Delta changed dramatically (Mooney et al. 1986; Howald 2000). In the early 1990s L. latifolium invaded and spread in tidal wetlands, along ephemeral stream corridors, and in disturbed upland areas at Rush Ranch (B.J. Grewell, personal observation). By 1995, L. latifolium had aggressively displaced formerly dense stands of endangered Cirsium hydrophilum var. hydrophilum (B.J. Grewell, personal observation). At Rush Ranch, several plant and animal species, including endangered endemic taxa, coexist with the weed as understory species (Spautz and Nur 2004; Reynolds and Boyer 2010). In 2003, L. latifo-

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lium was the third most frequent plant associate (85% frequency) of endangered C. h. var. hydrophilum at Rush Ranch (Fiedler et al. 2007). By 2005, L. latifolium had invaded 12% of a population of endangered soft bird’s beak that had been reintroduced at Rush Ranch in 2000 (Grewell 2005). This aggressive weed threatens the viability and recovery of endangered plant populations at Rush Ranch and elsewhere in San Francisco Estuary (Grewell 2005; Fiedler et al. 2007).

Apium graveolens A. graveolens (celery), a horticultural/garden escapee native to Europe, has invaded estuarine emergent wetland plant communities at Rush Ranch, greater Suisun Marsh, and the Carquinez Straits. Jepson (1923) and Mason (1957) noted the naturalization of A. graveolens in marshes and along streams in the Sacramento Valley and Southern California. The species was described as common in the San Francisco Estuary more than 30 years ago (Atwater et al. 1979), but invasive spread has been recent. In its native European range (Spain), relative cover and elevational amplitude of A. graveolens are low relative to other salt marsh plant community members, and the plant is restricted to high marsh (Sánchez et al. 1996). At Rush Ranch, A. graveolens is often closely associated with Lepidium latifolium, but it has a broader ecological amplitude than its co-invader and occupies a broader range of hydrogeomorphic settings than reported from its native range (B.J. Grewell, personal observation). Within 4 years of an experimental restoration of Chloropyron molle ssp. molle to the Spring Branch restoration site at Rush Ranch, A. graveolens had invaded C. molle ssp. molle subpopulations, and its frequency of occurrence was 18% (Grewell 2005). The frequency of A. graveolens with endangered Cirsium hydrophilum var. hydrophilum was as high (85%) as that of it co-invader, L. latifolium, at Rush Ranch (Fiedler et al. 2003). The invasive populations in the Potrero Hills region may be a source for new invasions westward in the Estuary. In 2009, A. graveolens first colonized the Southampton marsh preserve (Benicia State Recreation Area) in the Carquinez Straits, suggesting the need for greater recognition of this problematic invasive plant, and management and reduction of upstream source populations (Grewell 2010). 118



Exotic Annual Grasses A suite of exotic, winter annual grasses are invasive on the high marsh plain near the terrestrial ecotone, and also in seasonal wetlands, at Rush Ranch. Polypogon monspeliensis is native to Europe, Asia, and Africa. Evidence from adobe brick remains place P. monspeliensis introduction to California at the mid-nineteenth century (Frenkel 1977). Seasonally low salinity levels imposed by winter and anthropogenic runoff into estuarine wetlands control the distribution and abundance of P. monspeliensis (Callaway and Zedler 1998) because germination percentages of seeds decrease with increasing salinity. Thus, salt applications may be a practical control method (Kuhn and Zedler 1997). Hainardia cylindrica (syn. Monerma cylindrica, thintail, hardgrass) and Parapholis incurva (curved sicklegrass) are taxonomically similar European introductions. H. cylindrica is locally abundant in terrestrial ecotone and turf pans, Hill Slough, and Rush Ranch tidal marshes (P. Baye and B.J. Grewell, personal observations). P. incurva is less common at Rush Ranch and other tidal wetlands ringing the Potrero Hills, but locally co-occurs with H. cylindrica (P. Baye and B.J. Grewell, personal observations). The exotic cool-season grasses all have a C3 photosynthetic pathway, and the inherent lower photosynthetic rate suggests they will be competitively inferior in interactions with C 4 grasses such as native Distichlis spicata (Waller and Lewis 1979). At Rush Ranch, D. spicata is obviously more abundant than these exotic cool-season grasses. However, competitive superiority and relative abundance are not the only criteria by which exotic species should be considered in a management context. At Rush Ranch, seeds of these exotic grasses germinate in late November to February, and the exotic, annual grasses complete their annual growth cycle by late spring to early summer. During the pre-reproductive growth phase of the exotic annuals, the endangered hemiparasitic herb Chloropyron molle ssp. molle germinates and emerges as a seedling in exotic-grass-occupied habitat and forms parasitic connections with the roots of the exotic grasses; the exotic hosts die back when C. m. ssp. molle is in seedling stage (Grewell 2004). In a field study at Rush Ranch and Hill Slough, nearest neighboring plant species (potential hosts) were shown to greatly affect C. m. ssp. molle seedling survi-

Ecology: Organisms

vorship, and the presence of winter exotic grasses (particularly H. cylindrica) in the community was highly correlated with premature mortality of the endangered plant seedlings, while survivorship was highest when native D. spicata and Sarcocornia pacifica were nearest neighbors (Grewell et al. 2003; Grewell 2004). These results suggest that in estuarine vegetation at Rush Ranch and elsewhere, control of exotic winter grasses prior to restoration attempts is essential for sustainable populations of C. m. ssp. molle (Grewell 2004, 2005). The negative impacts of nonnative host plants suggest recovery planning for endangered hemiparasites must consider the costs of noncompetitive mechanisms when nonnative species removal is a priority in estuarine wetlands (Fellows and Zedler 2005; Grewell 2005).

Phragmites australis Cosmopolitan P. australis (common reed) is a large, perennial grass with creeping rhizomes and stolons that is found worldwide. Two recognized subspecies of P. australis (one native, the other exotic) are among the most misunderstood plant taxa in Suisun Marsh and at Rush Ranch. Fossil records dating to the Cretaceous and additional archeological records confirm a long presence of P. australis in North America as a minor native component of tidal wetland plant communities (Orson et al. 1987). In the past 150 years, a dramatic expansion of P. australis in North America has occurred to the point that it is considered a nuisance in many estuaries. This aggressive spread by vegetative growth may have both environmental and genetic causes, and multiple karyotypes are involved (Chambers et al. 1999). Molecular studies have confirmed that native, introduced, and Gulf Coast North American Phragmites lineages are genetically distinct, and invasive introduced populations do not represent a hybrid population type (Saltonstall 2003a). Native individuals persist in many midwestern and western states, including California, but introduced populations are also present, and recently introduced genotypes are largely dominant in the Atlantic coast region (Saltonstall 2003b). The typically noninvasive genotype—Phragmites australis (Cav.) Steud. ssp. berlandieri (E. Fourn.) Saltonstall and Hauber—native to California is present at Rush Ranch, nearby Peytonia Slough Ecological Reserve, and other tidal wet 

 

lands in the Delta and Suisun Marsh (B.J. Grewell and A.M. Shapiro, personal observations). This native taxa serves as host plant for Ochlodes yuma (Yuma skipper), which is associated with only the native genotype, while the exotic, invasive genotype P. australis ssp. americanus has been adopted by Poanes viator (broad-winged skipper, a large eastern Lepidopteran species) (Shapiro and Manolis 2007). In disturbed environments, both native and exotic genotypes can spread and displace competing macrophytes, though aggressive spread is more typical of the more recent, exotic invader. Differences between the two subspecies can be subtle and may partially depend on ecological conditions, but there are distinguishing morphological characters (Swearingen and Saltonstall 2010). The assumption that all P. australis present is of invasive taxa can be problematic, as some stands at Rush Ranch have persisted for decades. The presence of the native genotype at Rush Ranch that supports native insect species should be considered in management plans. Importance of Vegetation Presence and Type Coastal wetlands and their ecotones provide key ecological services and ecosystem functions (Emmett et al. 2000; Levin et al. 2001; Weslawski et al. 2004). Many of these services and functions are dependent on the composition and structure of plant communities (Bruno and Bertness 2001). In estuarine communities, vascular plants act as the major modifiers of the physical environment, provide primary energy and nutrient sources, and form most of the structural environment for other organisms. Critical marsh functions (such as nursery habitat provision, bank stabilization, runoff filtration, and trophic support) are directly and indirectly tied to the presence of vascular plants (Gleason et al. 1979; Warren and Niering 1993). At Rush Ranch and elsewhere within San Francisco Bay Estuary, vegetation type and structure, as well as marsh size and surrounding land use, are important in determining the distribution of multiple bird species (Spautz et al. 2006) and macroinvertebrate trophic relationships at nearby sites at Grizzly Island (de Szalay and Resh 1996). Thus, understanding and documenting the location and distribution of plants through time at properties like Rush Ranch is essential to effective management and preservation of these ecologically important habitats.

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A

b

C

D

Contemporary Vegetation Patterns Relative to Hydrogeomorphic Settings A wide range of environmental factors (i.e., hydroperiod, nutrient regimes, disturbance levels) and their interactions control the structure and composition of estuarine vegetation (Levine et al. 1998; Keddy 2000). Tidal submergence is a complex measure that serves as the primary control of the elevational ranges of tidal marsh plant species (Hinde 1954; Atwater et al. 1979; Macdonald 1988; Watson et al. 2011). Plant functional traits that convey stress-avoidance or stress-tolerance ability combine with competitive and facilitative interactions among plant species to influence estuarine plant species’ presence and abundance across environmental gradients (Keddy 1990; Bertness 1992; Pennings and Bertness 2001; Grewell et al. 2007; Grewell 2008a). At a local scale, environmental heterogeneity associated with hydrogeomorphic complexity combines with past land use and location to support distinct plant communities and assemblages. Vegetation patterns in oligohaline to brackish marshes like Rush Ranch are often more patchy (Crain 2008) than zonal, compared with tidal salt marshes. Thus we will discuss vegetation patterns in a geomorphic landscape unit context. Geomorphic units are planning areas delineated on the basis of integrated topographic, vegetation, and hydrologic features. These landforms can serve as the basis of conceptual physical models for soil/vegetation distribution and dynamics and are the major controls of habitat quality and spatial pattern of habitats over time. In addition, hydrogeomorphic units provide an appropriate context for description of the azonal nature of modern vegetation at Rush Ranch. Rush Ranch includes four major estuarine geomorphic units: subtidal channel beds, fring-

ing tidal marsh, tidal marsh plain, and tidal marsh-terrestrial ecotone (Figure 8.1A,B). It also includes three major terrestrial geomorphic units: hillslopes and inactive and active alluvial fans. Our focus is on the diverse array of estuarine wetland vegetation and ecotonal vegetation at the margins of tidelands at Rush Ranch. Here we describe the plant communities of subtidal channel beds, fringing tidal marsh, tidal marsh plain, tidal marsh ecotones, and tidal-terrestrial ecotones (including alluvial fans).

Tidal Channels Submerged aquatic vegetation (SAV) is made up of rooted flowering plants that grow primarily below the water surface. The primary Stuckenia pectinata (syn. Potamogeton pectinata, sago pondweed) and Ruppia maritima (widgeongrass) beds in San Francisco Estuary are around islands and other shallow areas in Honker Bay, Suisun Cutoff, and Suisun Bay (Schaeffer et al. 2007; California State Coastal Conservancy 2010). S. pectinata and R. maritima have long been recognized as important waterfowl food plants in managed wetlands throughout Suisun Marsh (George et al. 1965; Miller et al. 1975). Important food plants from out-of-state sources were planted extensively by duck club managers on Honker and Suisun Bay islands and throughout Suisun Marsh (Miller et al. 1975), and novel genotypes of S. pectinata and other waterfowl food plants may have been introduced and dispersed into Bay shallows (Mason 1957). Most SAV in the vicinity of Rush Ranch occurs in diked managed wetlands with perennial ponds and ditches, which in some years support substantial stands of S. pectinata, R. maritima, and Zannichellia palustris (horned pondweed) (Mason 1972). S. pectinata typically

Figure 8.1 (opposite).  Fringing high marsh and terrestrial ecotone at Rush Landing, Suisun Slough (A), where the high marsh plain is dominated by diverse patches of clonal forbs and graminoids, including Juncus arcticus subsp. balticus (Baltic rush), Potentilla anserina subsp. pacifica (silverweed), Grindelia stricta (gumplant), and Euthamia occidentalis (western goldenrod). Terrestrial ecotone shifts toward dominance by clonal perennials Distichlis spicata (salt grass), Leymus triticoides (creeping wildrye), and Ambrosia psilostachya (western ragweed). Fringing low marsh at channel’s edge is dominated by tall bulrushes, Schoenoplectus californicus and S. acutus. Also shown is the subtidal habitat that supports submerged aquatic vegetation (SAV) communities; these SAV patches are often found around shallows of small islands in the backwater areas of sloughs. The high marsh plain (marsh platform) of Rush Ranch, show in aerial view (B), is a complex and highly dynamic mosaic of vegetation patches. Tidally drained marsh plains near channels (C) support diverse assemblages of both tall and low-growing forbs (i.e., rushes, bulrushes, gumplant). Poorly drained marsh in the interior portions of the plain (D), remote from channels, supports three-square bulrush, salt grass, Baltic rush, sea arrowgrass, and low-stature pickleweed assemblages.

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dies back when water salinity exceeds 15 parts per thousand (ppt) (Kantrud 1990) but reappears with the return of oligohaline conditions. R. maritima is an opportunistic species that thrives in warm and less saline water (Kantrud 1991; Koch and Dawes 1991) yet also tolerates salinity fluctuations and marine conditions. In 2010, extensive beds of S. pectinata appeared in open subtidal beds of Suisun Slough near Goat Island, possibly in relation to declining suspended sediment supply and turbidity (Ganju and Schoelhammer 2010) and aqueous salinity (Moyle et al. 2010).

Fringing Tidal Marsh The fringing tidal marsh is positioned immediately above typically unvegetated subtidal channel beds (Figure 8.1A). This tidal marsh landform occurs as narrow bands on low edges of channels or at the edges of the marsh between “uplands” (hillslopes, scarps, alluvial fans) and tidal sloughs and appears to provide wave-damping, peatforming, and sediment-deposition functions comparable to those of the high fringing marshes investigated in Maine (Morgan et al. 2009). This landform supports plant species diversity and richness similar to (and in some locations, exceeding) vegetation of many high marsh plain areas at Rush Ranch. In contrast to the wave-attenuating marsh plains, fringing marshes of Rush Ranch are generally exposed to wind waves from open slough fetch from the west and northwest. Fringing marsh occurs as narrow bands along large, tidal sloughs at Suisun Slough and Hill Slough, particularly where the sloughs abut levee banks. Fringing marsh is also present where tidal sloughs border the neighboring hills with abrupt changes in slope that preclude development of tidal marsh plains (e.g., the reach of Suisun Slough immediately north of Rush Landing). In fringing or narrow tracts of tidal marsh, sinuous, complex tidal drainage networks are not able to develop because of the insufficient area available and the proximity of relatively steeper drainage gradients to the adjacent sloughs. Fringing marshes at Rush Ranch have developed extensively along the upland transition. Fringing marsh is also found in small, discontinuous segments that are directly exposed to wave action, forming dynamic peat slumps and scarps along slough edges. Fringing marsh may also be buffered by wave-damping, tule-dominated 122



low marsh. These tule-dominated areas may also serve as a barrier that can limit the growth and spread of adjacent marsh vegetation species toward tidal sloughs. Fringing marshes at Rush Ranch appear to have no history of ditching or diking and are generally composed of mostly organic (peat or muck) fine sediment, except at edges of active or recently active alluvial fans where better-drained mineral sediments are found. In some locations, substantial sediment has deposited along the exterior of artificial levees, allowing vegetation to colonize and expand outward for large distances into the slough. At Rush Ranch, fringing tidal marsh banks adjacent to dikes are steep scarps composed primarily of fine-grained peaty sediments. Fringing marshes adjacent to active alluvial fans or subject to slow current can support limited natural levees with overbank deposits. These natural levees include better-drained sediments that support vegetation less tolerant of long hydroperiods. The fringing marsh at Rush Ranch is bordered by inundation-tolerant Schoenoplectus/Typha/ Carex (bulrush/cattail/sedge) associations at lowest elevation. At middle to higher elevations, the fringe vegetation usually is composed of subshrubs, creeping perennial forbs and rushes, and grasses as well as bunchgrasses that are tolerant of brackish salinity conditions. Tall, shrubby Grindelia stricta var. angustifolia (syn. G. hirsutula, marsh gumplant) often borders tidal creek banks and provides dense emergent cover in mature tidal marshes (Baye 2007); it is also abundant in high fringing marsh. The vegetation typical of widespread well-drained, high banks of mature tidal creeks of Rush Ranch is also abundant in high fringing marsh, including tall, densegrowth forms of Sarcocornia pacifica, Frankenia salina (alkali heath), Potentilla anserina var. pacifica (syn. Argentina egedii, silverweed cinquefoil), Glaux maritima (sea milkwort), Distichlis spicata, Deschampsia cespitosa (L.) ssp. holciformis (tufted hairgrass), Juncus arcticus ssp. balticus (Baltic rush), and other associated subshrubs and forbs. In addition to these broadly distributed species, the high fringing marsh supports species with small habitat ranges and narrow salinity tolerances, including Sium suave (water parsnip), the rare Cicuta bolanderi (Bolander’s water hemlock), Helenium puberulum (sneezeweed), Eryngium heterophyllum (coyote thistle), and Oenanthe sarmentosa (water parsley), as well as more widespread

Ecology: Organisms

perennial brackish wetland forbs like Euthamia occidentalis (western goldenrod), Ambrosia psilostachya (western ragweed), and Pluchea odorata (marsh fleabane). Slumps and scarps of waveimpacted fringing marsh locally support opportunistic colonizers Lilaeopsis masonii (Mason’s lilaeopsis, western grasswort) and Isolepis cernua (low club rush). Below freshwater seeps in wavecut low bluffs, fringing marsh at Rush Landing supports distinctive stands of freshwater marsh species discussed in the Freshwater Seepage Sites section below. In recent years, Lepidium latifolium has invaded fringing marsh at Rush Ranch, but at present it is infrequent in this geomorphic setting.

Tidal Marsh Plains Typically, tidal marsh plains (platforms) at Rush Ranch are wide tidal landforms dissected by complex, sinuous dendritic channels. Compared with surrounding, diked areas in Suisun Marsh, Rush Ranch channels have been altered significantly less. However, regional and on-site ditching, partial diking, and dam constructions have reduced sediment supply and thus altered the historic channel sinuosity. Internal landforms and vegetation zones of the tidal marsh plain (tidal creek banks and natural levees, artificial channels and dikes, tidally vs poorly drained plains, ponds, and turf pans) will be discussed individually. Tidal Creek Banks and Natural Levees Tidal creeks are a key feature of natural estuarine wetlands that dissect marsh plains and range from large creeks that rarely drain, to small creeks that are covered with vegetation (Leopold et al. 1993). The banks of these creeks are a distinct landform at Rush Ranch, formed by gradual overbank accretion of sediment and debris at stable bank positions. The vegetation present on these banks and levees is strongly influenced by subsurface drainage of the adjacent creeks. The tidal creek banks are regularly subjected to brackish tidewater and support tall emergent graminoids and forbs such as Schoenoplectus spp., Carex spp., and Typha spp. (cattails), as well as more diminutive plants such as Lilaeopsis masonii, Isolepis cernua, Hydrocotyle verticillata (water pennywort), and Triglochin striata (three-ribbed arrowgrass). Hill Slough’s low creek banks support extensive colonies of Carex lyngbyei (Lyngbye’s sedge), an

oligohaline tidal marsh species typical of the Pacific Northwest. This is a disjunct population apparently unique in the San Francisco Estuary. At Rush Ranch, it is established at the lower end of the low tule marsh zone. The more elevated upper creek banks are often a habitat for tall forbs and subshrubs such as Grindelia stricta, providing dense flood refuge and cover for marsh wildlife. Upper creek banks also support rare or endangered plants, such as Cirsium hydrophilum var. hydrophilum, Cicuta bolanderi, and Lathyrus jepsonii ssp. jepsonii (Jepson’s Delta tule pea). Invasive nonnative clonal forbs such as Lepidium latifolium also occupy this habitat. At Rush Ranch, the spread of L. latifolium is frequently along these tidal channels and the upland margin of other tidal marshes near Potrero Hills (Grossinger et al. 1998; Boyer and Burdick 2010). In fresh to brackish tidal areas, small Lilaeopsis masonii and Triglochin striata are found in the marsh ground layer below the canopy of tall emergent macrophytes along tidal sloughs and slumping banks of in-channel islands. The macrophytes may include Schoenoplectus californicus, S. acutus, Typha domingensis, T. angustifolia, T. latifolia, and Phragmites australis, either in mixed or in monospecific stands. Extensive marsh plains within the brackish marsh are dominated by Distichlis spicata. Where tidal creeks introduce complexity, we also find Sarcocornia pacifica, Limonium californicum (hyphenate sea lavender), Atriplex prostrata (common spearscale), Glaux maritima, Jaumea carnosa (fleshy jaumea), Triglochin maritima (seaside arrowgrass), Isolepsis cernua (club rush), I. carinata (keeled club rush), Helenium bigelovii, Pluchea odorata, Deschampsia cespitosa, and Oenanthe sarmentosa. Other rarer plants such as Symphyotrichum lentum (Suisun Marsh aster), Lathyrus jepsonii ssp. jepsonii, Cicuta bolanderi, and Eleocharis parvula (dwarf spikerush) also occur in this zone. When the depth and duration of flooding increase during wet years, midzone diversity is reduced in Suisun Marsh as mosaics of more flood-tolerant Juncus arcticus ssp. balticus and Schoenoplectus americanus expand (B.J. Grewell, unpublished data). Artificial Channels and Levees The creation of channels and dikes at Rush Ranch most likely began in the late nineteenth

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century for agricultural purposes. Ditching of tidal marsh plains with poor drainage (mosquito ditches) or tidal marsh pans and ponds such as in the “Mallard Slough” vicinity created ponded habitats attractive to dabbling ducks (Wetlands and Water Resources 2011). At Rush Ranch, nonengineered ditching extended and connected the distal ends of small tidal creeks. In addition, slough dams and partial dikes were constructed along the marsh perimeters on branches of Second Mallard Slough, Suisun Slough, and Hill Slough. The resultant steeply elevated berms line rectilinear channels, creating crests above the marsh plain that are only flooded at the most extreme high tides. Throughout most of Suisun Marsh, dikes (artificial levees) adjacent to tidal marshes have replaced much of the natural flood refuge habitats formerly provided by natural marsh levees or terrestrial vegetation. In this artificial and constantly changing ecotone, ruderal species, such as Lepidium latifolium, Annagalis arvensis (pimpernel), Brassica spp. (wild mustards), Raphanus spp. (wild radish), Foeniculum vulgare (fennel), Helminthotheca echioides (bristly ox tongue), Conium maculatum (poison hemlock), and Rubus armeniacus (Himalayan blackberry) thrive. These ruderal species are characterized by high reproductive abilities, fast growth rates, and short life spans and are thus capable of thriving in a frequently changing depositional area (Grime 1977). Tidally Drained Marsh Plain This zone of the tidal marsh is influenced by surface and subsurface drainage of adjacent creeks and ditches, which limits soil waterlogging, salt accumulation due to evapotranspiration, and ponding, in contrast to poorly drained marsh plain (discussed below). The tidally drained marsh plain at Rush Ranch is extensive, compared with other relict tidal marshes with small to no remnant tidal plain habitats (Figure 8.1B,C). This brackish landform at Rush Ranch is dominated by mixed creeping subshrub and graminoid species, such as Distichlis spicata, Sarcocornia pacifica, and Lepidium latifolium. Plants such as Glaux maritima and Senecio hydrophilus, previously common in this zone throughout the Estuary, persist at Rush Ranch yet rarely occur on other similar properties. Within the well-drained marsh plain, there are middle- and high-elevation areas (high 124



and mid marsh) occurring in a patchy mosaic distribution indicative of altered hydrology. Traditionally, high marsh is defined as the area from approximately mean higher high water to extreme high water (occurring on spring tidal cycles) (Josselyn 1983; Peinado et al. 1994). Much of the marsh native plant and animal biodiversity, including regionally rare and endangered species, is found, in particular, in the high marsh. The common holoparasitic vine Cuscuta pacifica var. pacifica (salt marsh dodder) and the endangered root hemiparasite Chloropyron molle ssp. molle suppress perennial dominants in the community and enhance plant species richness in the high marsh (Grewell 2008a,b). The marsh at Rush Ranch is primarily high marsh plain dominated by Distichlis spicata, Sarcocornia pacifica, and Frankenia salina and locally abundant C. p. var. pacifica and Grindelia stricta var. angustifolia or G. × paludosa. At Rush Ranch, Arthrocnemum subterminale (Parish’s glasswort) is found in upper Spring Branch Creek; in the emphemeral drainage in Suisun Hill Hollow below Suisun Hill spring, stock pond, and Grizzly Island Road; and as small rare patches near termini of firstorder tidal creeks associated with Second Mallard Branch Slough. Other co-occurring species include introduced Cotula coronopifolia (brass buttons), in areas where water occasionally pools, and Atriplex prostrata. Endangered C. m. ssp. molle also occurs along high marsh ecotones, at drainage divides within marsh plains, and near high-order tidal creeks. Similar to levee and berm habitats, the high marsh is also susceptible to invasion by many nonnatives, including L. latifolium, Apium graveolens, Lotus corniculatus (bird’s foot trefoil), Bromus diandrus (ripgut brome), Hainardia cylindrica (barbgrass), Parapholis incurva, and Polypogon monspeliensis. Rumex crispus and R. pulcher (curly and fiddle docks) have been reported at the edges of brackish high marshes at Rush Ranch but are not believed to be invasive (Baye et al. 2000). In slight depressional areas of the marsh plain that experience more extended hydroperiods, a number of marsh plants co-occur in a patchy mosaic distribution that reflects subtle changes in sediment characteristics and hydrology. Plant species here include Jaumea carnosa, Frankenia salina, Cuscuta pacifica var. pacifica, Triglochin maritima, Distichlis spicata, Juncus arcticus ssp. balticus, and Sarcocornia pacifica. Glycyrrhiza

Ecology: Organisms

lepidota (wild licorice), rare in tidal wetlands, is associated with J. a. ssp. balticus in the marsh plain adjacent to Hill Slough. In this zone, S. pacifica is often in its tallest, most robust form among all habitats within the marsh (Suisun Ecological Workgroup 1996). Although Watson and Byrne (2009) found that D. spicata had been nearly replaced by Schoenoplectus americanus and Bolboschoenus maritimus in the mesohaline marshes, including Rush Ranch, their sampling was limited to a single season. Historically and in modern times, D. spicata is the main dominant in this zone, also reaching its maximum height form in this habitat (Mason 1972). Well-drained peat sediments bordering first-order tidal creeks and mosquito ditches dissecting the plain support the endemic, federally endangered Suisun thistle (Cirsium hydrophilum var. hydrophilum) (Fiedler and Keever 2003; Fiedler et al. 2007), as well as Senecio hydrophilus (alkali marsh ragwort), Pluchea odorata, and Grindelia spp. The five most dominant plant species (measured as canopy cover) associated with endangered C. h. var. hydrophilum bordering tidal creeks have been native Potentilla anserina var. pacifica, Schoenoplectus americanus, Juncus arcticus ssp. balticus, and Grindelia stricta, plus exotic Lepidium latifolium (Fiedler et al. 2003). In addition to these dominant species, results of a marsh-wide census at Rush Ranch indicate, exotic Apium graveolens, Atriplex prostrata, and Rumex crispus also frequently occur with the endangered thistle, though they are not dominant in the association (Fiedler et al. 2003). It is interesting to note that the dominant plant species associated with endangered C. h. var. hydrophilum are also recognized to be key indicator species for California black rail breeding habitat (particularly S. americanus) and California clapper rail breeding habitat (particularly Grindelia stricta) in Suisun and North Bay marshes (Evens and Nur 2002; Evens 2010). Poorly Drained Marsh Plain Contrasting with tidally drained marsh plain, the poorly drained marsh plain habitat is remote from tidal channel drainage and is inefficiently drained primarily by slow overland sheeting flow or very slow infiltration and evapotranspiration (Figure 8.1B,D). Elevated groundwater and soil salt accumulation due to evapotranspiration are important structuring processes in this marsh plain zone.

Similar to the well-drained marsh habitats, this habitat is dominated by mixed creeping subshrub and graminoid species, such as Distichlis spicata, Sarcocornia pacifica, and recently, Lepidium latifolium. At Rush Ranch, patchy sections of nonnative Phragmites australis occur in the poorly drained portions of the marsh plain habitats. Of all the zones on the well-drained marsh plain, the D. spicata–dominated zone is least invaded by L. latifolium, potentially because of this poor drainage and resultant long hydroperiod (C.R. Whitcraft, unpublished data).  

Marsh Plain Ponds The ecogeomorphic origins of tidal marsh plain ponds (variously termed pools, ponds, pans, or pannes in different regions and times: Harshberger 1916; Pethick 1974; and Adamowics and Roman 2005) in Suisun Marsh are not known but may have structure and secondary origins similar to those in mature high peat tidal marshes of the U.S. Northeast (Wilson et al. 2009, 2010). These ponded depressions in the tidal marsh plain are isolated from drainage networks, allowing them to maintain permanent standing water, except where they have been degraded or destroyed by marsh ditching (MacDonald et al. 2010). As a rare vegetation habitat at Rush Ranch, the ponds support Stuckenia pectinata and epiphytic green algae. The tidally restricted ponds of Goat Island Marsh (diked marsh) also support stands of S. pectinata and the floating-leaved Potamogeton nodosus (pondweed). High Marsh Pans (Turf Pans) These high marsh pans occur from the upper edges of the high marsh to the lower edges of the alluvial fan. They are often poorly drained during high winter tides and dry in the summer neap tides. The vegetation on these pans is dominated by turf-like low or prostrate perennial and annual graminoids, forbs, and subshrub vegetation. They are similar in structure to playas (shallowflooded, seasonally desiccated, and hypersaline wetlands in arid or semiarid flats or basins) or saline vernal pools and dominated by annual forbs, graminoids, perennial grasses, and prostrate subshrubs. These species include Polypogon monspeliensis, Hainardia cyclindrica, Lasthenia glabrata ssp. glabrata (goldfields), and Juncus bufo-

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nius (toad rush). Plants such as Triphysaria versicolor ssp. versicolor (butter and eggs) appear rarely in turf pans of south Rush Ranch tidal marsh in association with L. glabrata, Isolepsis cernua, and J. bufonius, as well as prostrate Sarcocornia pacifica. This is ecologically distinctive as the only reported occurrence of T. versicolor in a brackish tidal marsh; the species is typically found in the region within seasonal wetlands and alkali grasslands. Lepidium latifolium appears to be consistently excluded from the summer-desiccated high turf pans, potentially by high pore water salinity or low pore water in general. In contrast to many estuarine marshes, the high intertidal zone of San Francisco Estuary brackish wetlands can support the greatest richness of plant species in the marsh. However, at limited locations in the Suisun Marsh where the highest marsh elevation zone is still intact, tides are muted, summer temperatures can exceed 38° C, soil pore water can be hypersaline (>40 ppt, and in places >100 ppt) (Grewell et al. 2007; Grewell 2008a), and only Arthrocnemum subterminale, Sarcocornia pacifica, and Cressa truxillensis (alkali weed) are found.

Freshwater Seepage Sites Unique communities can occur on the upper edge of brackish marshes, where salt water rarely reaches or where salt is diluted by freshwater seepage. Oenanthe sarmentosa can be relatively abundant in wet years, particularly in or near freshwater seepages adjacent to low shoreline bluff scarps or drainages from upland swales. Other predominantly freshwater marsh species, including Mimulus guttatus (monkeyflower) and small-fruited sedge, appear anomalously in the middle marsh zone of fringing marshes at Rush Landing below seeps in the high marsh zone that support Sisyrinchium bellum (blue-eyed grass) and mixed stands of Carex barbarae and Leymus triticoides (creeping wild rye).

Seasonal Wetlands The primary seasonal wetland at Rush Ranch occurs along the Spring Branch corridor between the South Pasture Trail and Grizzly Island Road. Prior to nineteenth-century alterations, this area was an extension of the historic Holocene tidal wetland. The freshwater input at 126



this site is inhibited by a stock pond and the road upstream of the site; in addition, the tidal flow within this area is restricted by a berm and by a culvert (pipe) under the trail at the west end of the area. Despite these water flow restrictions, this area retains some plant species typical of a seasonal wetland habitat. This habitat is dominated by nonnative grasses: Hordeum marinum ssp. gussoneanum (Mediterranean barley), Lolium multiflorum (Italian ryegrass), Polypogon monspeliensis, Hainardia cylindrica, and Parapholis incurva. However, there is muted tidal influence from First Mallard Branch Slough, and in wetter years the soil is inundated and saline, as indicated by the presence of obligate wetland plants: Arthrocnemum subterminale, Sarcocornia pacifica, Cressa truxillensis, Frankenia salina, Juncus arcticus, and Bolboschoenus maritimus. These plant species persist within this habitat, occupying remnant channels and floodplain. Ephemeral vernal flora along the terrestrial ecotone at Upper Spring Branch also includes Lepidium oxycarpum (forked pepperweed), Muilla maritima (common muilla), Lasthenia glabrata, Triphysaria eriantha ssp. eriantha, T. versicolor ssp. faucibarbata, and in some years the rare Lasthenia conjugens (Contra Costa goldfields).

Diked Marsh Fringing marshes throughout the Suisun Marsh were frequently converted to diked marshes for hunting. At Rush Ranch, a muted-tidal impoundment of tidal marsh (diked marsh south of Goat Island) and hunter’s cabin were added by the Rush family prior to 1900, and partial diking within the marsh plain also supported hunting pursuits on-site. Today, the impoundment includes a more complete but low levee along Suisun Slough, and two water control structures (at the north and south ends) allow limited inundation from the neighboring slough. Although the dikes and water control structures at Rush Ranch have not been thoroughly maintained, water levels within the diked marsh do not fluctuate to the full extent of the surrounding undiked marsh. There is no levee on the eastern/landward side of the diked marsh. Here, tall, robust stands of Sarcocornia pacifica and Distichlis spicata transition to grassland. While diked marshes share some of the dominant plant species with natural marsh areas, the

Ecology: Organisms

altered hydrological conditions in the diked, nontidal marshes do not support many of the rare or uncommon plant and animal species found in the more natural tidal marshes. Such is the case at Rush Ranch. The diked marsh is dominated by native cattails (Typha latifolia, T. domingensis) and bulrushes (Schoenoplectus californica, S. acutus var. occidentalis, S. pungens, and S. americanus) (personal observation). In addition, Phragmites australis has colonized the more disturbed areas along the south edge, with observed spread into the more interior regions. The artificial levee around the diked marsh supports an abrupt break in vegetation across a short and artificially steep slope, bordering a narrow fringing marsh. However, there are narrow bands of middle and high brackish marsh vegetation on the levee, including Distichlis spicata, Sarcocornia pacifica, and Grindelia stricta var. angustifolia, as well as Euthamia occidentalis (western goldenrod) and Calystegia sepium (morning glory). The upper zones of dikes are typically weedy and support a variety of introduced and invasive species, including Rubus armeniacus that frequently weakens the structure of the levee itself. The dike at the Goat Island Marsh at Rush Ranch is no exception and contains a community dominated by R. armeniacus, annual forbs, nonnative forbs including Lepidium latifolium in small patches, and Raphanus sativus (cultivated radish), as well as large colonies of Phragmites australis extending from the adjacent slough and diked marsh. Potential plans to restore tidal inundation to this area would dramatically alter the existing vegetation patterns.

Tidal Wetland: Terrestrial Ecotones Lowland Grassland (Sedge Rush Meadow) Since grazing began to be restricted at Rush Ranch in the 1990s, the lowlands (sandy to silty alluvial fan edges near sea level) have regenerated extensive stands of a dominant native clonal perennial grass, Leymus triticoides, along the ecotone between alluvial fan edges and tidal marsh. L. triticoides forms extensive, spreading clonal colonies that coalesce and extend up to the fence line that restricts grazing (currently less than 10–20 m above the highest tide lines). This grass also extends down to intergrading stands of Sarcocornia pacifica, Distichlis spicata, Cressa truxillensis, and Frankenia salina. Leymus  

triticoides is abundant to dominant in floodplain grasslands and lowland swales and was likely a dominant element of lowland mesic or seasonal wetland grasslands in California (Holstein 2001). Its recent spread in areas where grazing pressure has been reduced at Rush Ranch, and elsewhere where agricultural crop production was abandoned, suggests that it was a widespread, if not dominant, element of tidal marsh ecotones with lowland grasslands. The stands of L. triticoides at Rush Ranch may represent the most extensive and phenotypically diverse of any remnant tidal marshes in Suisun Marsh and the greater San Francisco Estuary (P. Baye and B.J. Grewell, personal observations). Other clonal, graminoid species of seasonal wetland sedge meadows and grasslands have regenerated extensive, locally dominant stands at the tidal marsh ecotone of Rush Ranch following local restriction of intensive cattle grazing, including Carex praegracilis (field sedge, locally abundant at southeast Rush Ranch tidal marsh edges) and Carex barbarae, particularly near seeps or swales with seasonally saturated or mesic soils. Riparian Bluff The north-aspect bluffs (wave-cut or channelcut scarps in low sandstone hillslopes) of Suisun Slough and Hill Slough support remnants of native woody riparian scrub that are otherwise very scarce in swales of Rush Ranch and Potrero Hills, which are heavily grazed. The steep bluffs are inaccessible to cattle and likely have provided a natural refuge from grazing where slopes approach vertical and support shallow seeps. The ground layer of the riparian bluffs includes lowland grassland/sedge meadow elements (Leymus triticoides, Carex barbarae), and the patchy woody shrub thickets are dominated locally by Rosa californica (California rose), Sambucus mexicana (elderberry), Toxicodendron diversilobum (poison oak), and Baccharis pilularis (coyote brush). The lower branches of riparian scrub in some locations provide structural support for vines of the rare Lathyrus jepsonii var. jepsonii established in the upper tidal marsh edge below the bluffs. Large patches of Rubus armeniacus also occur in riparian bluffs of Hill Slough. Glycyrrhiza lepidota (wild licorice), a riparian subshrub, is locally abundant along upper tidal edges of low bluffs along Hill Slough, locally spreading into high marsh.

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Active Alluvial Flats The “hollows” of Suisun Hill and Spring Branch Creek (shallow ephemeral creeks and swales draining grasslands) develop low-gradient lower reaches that form braided alluvial fan distributaries with disturbed, fine, slightly saline sands and silts (derived from marine sandstones), grading into tidal marsh edges. The alluvial flats are for the most part intensively grazed and trampled, and they include barrens as well as herbaceous lowland grassland assemblages similar to those of tidal marsh ecotones, including Lolium perenne (ryegrass), Distichlis spicata, Frankenia salina, and Cressa truxillensis. Near low channel scarps and relatively well-drained edges, stands of Hemizonia fitchii (spikeweed) and Arthrocnemum subterminale are locally common.

Modern Transformations: Climate Change It is particularly noteworthy that the Estuary’s regional climate in the historic period (post-1850) has been relatively stable compared with the majority of the tidal marsh stratigraphic record, with most of the historic change in the salinity signal due to water diversion in the Delta (Byrne et al. 2001). The primarily fresh-brackish phase of Rush Ranch tidal marshes known from the early historic period is not a permanent or prevailing condition but a long freshwater phase that began only 750 years BP. Most significantly, perhaps, is that the entire geomorphic and ecological history of Rush Ranch tidal marsh plains occurred under a regime of slow sea level rise and gradual accretion of marsh peat (1.3 mm/y) during the formation of the mature marsh plain (Byrne et al. 2001). No part of the marsh’s history reflects the conditions that are expected in the twenty-first century: accelerated sea level rise rates significantly greater than 2 mm/y, prolonged warmer climate with reduced Delta outflows, and seasonal Delta outflow limited to the wet season because of reduced or absent Sierra snowpack. Modern operation of state and federal water projects damp seasonal and annual outflow and salinity variability, yet climate change is the most powerful driver of long-term variability at Rush Ranch and regionally (Enright and Culberson 2009). Over the past 30 years, the large annual ranges 128



of channel salinity in Suisun Marsh have also had considerable temporal and spatial variation. This high interannual variability in salinity is likely key to a productive ecosystem that supports native biota (Atwater et al. 1979; Fox et al. 1991; Peterson et al. 1995; Byrne et al. 2001; MalamudRoam et al. 2007; Moyle et al. 2010). Summer salinity is projected to increase in the Suisun Marsh because increasing spring air temperature is causing snowmelt runoff into the estuary to occur earlier in the year (Knowles and Cayan 2002). Recent projections of areas vulnerable to sea level rise suggest variable effects at Rush Ranch that correspond to the magnitude of increases in water elevation (Knowles 2010). For example, with sea level increases of 50–150 cm relative to mean lower low water (MLLW), it is projected based on present-day elevations that wetland elevations of the diked wetland and Hill Slough regions of Rush Ranch will drop to below MLLW tidal datum, while the tidal marsh associated with Suisun and Cutoff Slough will remain above MLLW. Projections of wetland elevation increases in the range of 100–150 cm above MLLW suggest Rush Ranch tidal wetlands will be among extremely rare and isolated wetlands above MLLW relative to a largely inundated Suisun Marsh. However, these projections ignore the potential for vertical accretion and lateral migration of wetlands (Knowles 2010). Certainly, understanding how Rush Ranch vegetation may respond to predicted sea level rise will depend on understanding sediment supply and accretion (Orr et al. 2003; Callaway et al. 2007). At Rush Ranch, the average marsh accretion rate over the last 750 years has been approximately 1.5 mm/y, close to the average rate of sea level rise at San Francisco for the period AD 1855–1986 (Lyles et al. 1988; Byrne et al. 2001). The actual changes in salinity and inundation regimes at sites such as Rush Ranch are difficult to predict, and heterogeneous effects could result in increased plant species evenness (Watson and Byrne 2009). Some studies suggest that increases in salinity and submergence of wetlands associated with sea level rise in Suisun Marsh will prompt local-scale declines in plant species richness and productivity (Callaway et al. 2007). If Rush Ranch tidal marsh plain accretion rates fall below rates of accelerated sea level rise and hydroperiods increase, large-scale marsh vegetation zonation changes and domi-

Ecology: Organisms

 

 

 

nance shifts within both the marsh plain and terrestrial ecotone would be expected (Watson and Byrne 2009). High marsh and terrestrial ecotone assemblages would be likely to shift landward and invade low-gradient stream valleys like Spring Branch Creek and Suisun Hill Hollow. Expansion of lower tidal marsh assemblages tolerant of prolonged flooding, such as tules, bulrushes, or sedges, would be expected to displace salt grass, rush, and perennial forb assemblages within the tidal marsh platform. High marsh assemblages dominated by tall perennial forbs along tidal creek banks, internal to the marsh plain, would also be at risk of conversion to more flood-tolerant wetland graminoid assemblages. Failed levees and expanded subtidal basins in the vicinity of Suisun and Montezuma Sloughs can result in a reduction of tidal range due to tidal prism increase. If such tidal damping interactions with sea level rise are significant at Rush Ranch, dominance by flood-tolerant wetland graminoid vegetation may be intensified. This condition may have parallels with the earliest vegetation history of Rush Ranch, evident in sediment cores showing foundering Cyperaceaeand Poaceae-dominated marsh and mudflat in its early stages of formation prior to 1,750 years BP, before formation of the high marsh platform (Byrne et al. 2001).

Synthesis and Future Directions Assembling information about the history and current status of vegetation at Rush Ranch highlights gaps in our knowledge and enables predictions of future vegetation trends. Enumeration of such research gaps has been identified by the San Francisco Bay National Estuarine Research Reserve as a management priority in their 2011– 2016 management plan in order to provide ideas for researchers, especially graduate students, as well as to facilitate cooperation among researchers. We see one major area of focus as the need to determine the types and levels of ecological impacts resulting from different management actions of the Solano Land Trust (i.e., grazing, fencing locations, and stock pond management) that have the potential to affect estuarine vegetation communities on a large geographic scale. In addition, development of standardized and regularly occurring monitoring of submerged aquatic vegetation communities and rare plant  

populations would improve the ability to manage for their success and continued recruitment. In parallel with monitoring rare plant populations, it would be ideal to conduct research to support predictive modeling of nonnative plants within the property and among neighboring properties. Successful modeling of potential ranges of a given plant species requires growth parameters for each invasive plant, obtained through controlled field or greenhouse experiments. Several ecological restoration projects are being considered for Rush Ranch while surrounding tidal wetlands are heavily invaded with exotic plant species. Given the high numbers of exotic species in the estuarine flora at Rush Ranch, research is needed to support ecologically based, comprehensive (multiple species) weed management strategies that will promote recovery of sustainable native plant communities. Restoration projects should be paired with research on short- and long-term responses of target weeds, native plant indicator species, and native plant communities to restoration and management actions. Research is needed to evaluate the range of variability in estuarine conditions (i.e., salinity, tidal flows) that will be needed to support more heterogeneous, native vegetation associated with specific hydrogeomorphic landform units (see Moyle et al. 2010). Focusing on these data gaps can inform adaptive management planning and actions for conservation and recovery of native plant communities at Rush Ranch. Climate is changing across a range of scales, from local to global, yet ecological consequences of the predicted changes are difficult to understand and predict. Accurate predictions of the future impacts of climate change (e.g., sea level rise, sediment transport, and salinity regimes) on plant diversity and distribution are critical to the development of conservation strategies and management plans. Incorporation of climate change factors, such as sea level rise and subsidence, is essential to accurate predictions. As discussed throughout this chapter, paleoecological reconstructions of geology, climate, sedimentation, and vegetation at Rush Ranch demonstrate similarities to potential future climate scenarios in terms of high sea level rise rates, low sediment transport rates, increased summer peak salinity, and a landward migration of high salinity levels. In addition, climate change is also predicted to interact with other drivers of biodiversity change,

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such as habitat destruction and fragmentation or the introduction of nonnatives. Thus, there is a critical need for manipulative experiments that address how life history strategies and functional traits of plant species that influence ecological processes (e.g., dispersal, invasion, nutrient cycling) might drive the ability of plant species to respond to climate perturbations. Acknowledgments We would like to thank the Solano Land Trust and San Francisco Bay National Estuarine Research Reserve for property access and personal expertise. We thank Simon Malcomber, Matt Ferner, and two anonymous reviewers for critiquing our draft manuscript. This publication was partially supported by the CALFED postdoctoral research fellowship to C. Whitcraft and D. Talley and by academic funding to C. Whitcraft from California State University–Long Beach.  

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chapter Nine

Invertebrates Past and Current Invasions Elizabeth D. Brusati

contents

1986; Peterson and Vayssieres 2010). The Bay is the largest estuary in western North America, and it contains the largest area of marshes and other wetlands (Callaway and Zedler 2009). However, 93% of the wetlands in the Bay have been destroyed in the past century and a half, mostly by filling for development (Zedler and Callaway 2009). Other chapters of this book discuss these alterations more thoroughly. Invasive species are only one facet of the human-caused changes in San Francisco Bay and other estuaries. As a heavily urbanized estuary, the Bay has more potential pathways for invasive species and experiences invasion at a greater frequency than nearby bays. Like the physical environment of the Bay itself, the invertebrate community differs greatly from that which existed prior to the 1849 Gold Rush. Terminology used in biological invasions can be confusing. In general, introduced, exotic, nonnative, and nonindigenous are interchangeable terms that refer to organisms that did not evolve in a particular location but arrived there through human actions. In North America, introduced usually refers to species that arrived after European settlement. Invasive refers specifically to those introduced species that cause ecological or economic harm (National Invasive Species Council 2001). By this definition, not all introduced species are invasive. In addition, cryptogenic refers to species whose origin is not clear and which may be either native or introduced (Carlton 1996). Fouling species refers to species that attach to hard substrates such as rocks, docks, or boat

Pathways of Introduction General Patterns Ecosystem Engineers Competition and Predation Interactions among Invaders Ongoing Invasions Synthesis and Future Directions

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stuarine invertebrates represent a wide taxonomic range, from crabs and snails crawling across mudflats, to tiny clams and annelid worms buried within the sediment. They must survive physical stresses such as tidal fluctuations, variations in salinity, and lack of oxygen within waterlogged sediment (Pennings and Bertness 2001). They must also contend with predation by the more conspicuous marsh inhabitants, such as shorebirds and fish. However, these species provide important functions by altering sediment structure, affecting small-scale hydrology, cycling nutrients, and providing food for higher trophic consumers (Mitsch and Gosselink 1993). Salt marshes around the world face humancaused stresses as well, including filling of marshes for development or agriculture, pollution, and rising sea levels due to climate change. San Francisco Bay has experienced these as well as many other alterations (Nichols et al.





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hulls and includes species such as bryozoans, tunicates, and algae. Introduced species can influence both physical and biological interactions within their new locations. They may change ecosystem processes such as nutrient cycling or hydrology, change disturbance regimes, or alter habitat structure (Jones et al. 1997; Vitousek et al. 1997). They may compete with, prey upon, or facilitate native species. Most introduced species that have been well studied in coastal or estuarine environments have demonstrable negative impacts on native communities (Grosholz 2002; Williams and Grosholz 2008). This chapter focuses on epifaunal (surfacedwelling) and benthic infaunal (belowground) organisms introduced to San Francisco Bay marshes and associated mudflats. Little information was found on insects. Josselyn (1983) provides a good overview of invertebrate communities in San Francisco Bay marshes. Virtually no information exists about the invertebrates within San Francisco Bay salt marshes before human activities altered the physical and biological environment of the Bay. The Bay is frequently called “the most invaded estuary in the world,” with an estimated 234 introduced species throughout the ecosystem (all taxa, including plants), an additional 125 cryptogenic species, and more continuing to invade (Cohen and Carlton 1998). These numbers include 15 annelid worms, 27 mollusks, and 51 arthropods living in salt or brackish water. Surveys by the California Department of Fish and Game recorded more introduced species in San Francisco Bay than in any of California’s other six major harbors, with 190 introduced and 45 cryptogenic species (Resources Agency 2008, results available online at www.dfg.ca.gov/ospr/Science/invasive_­species .aspx). Within the San Francisco Estuary, introduced species represent 40%–100% of the common species, 97% of the total number, and 99% of the biomass (Cohen and Carlton 1998) (Table 9.1). Comparisons with other bays are complicated by differences in size, level of human activities, potential invasion pathways, and survey methods. However, Chesapeake Bay, which is larger and has a much longer history of urbanization, contains 170 introduced species (again, all taxa), including 16 marine invertebrates shared with San Francisco Bay (Fofonoff et al. 2009). Elkhorn Slough, a much smaller, less-developed estuary located south of  

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San Francisco, contains at least 56 introduced estuarine invertebrates (Wasson et al. 2001). The large number of cryptogenic species in San Francisco Bay (Cohen and Carlton 1998) makes it difficult to separate the effects of most introduced species from the overall marsh community. In some cases, genetic studies or other evidence have reclassified native species as likely introductions (Carlton 1979). For example, one of the most common clams in San Francisco Bay mudflats is Macoma petalum (Poulton et al. 2004; Brusati and Grosholz 2006), previously described as M. balthica or M. inconspicua. Genetic and shell studies revealed that there are two species of Macoma, neither native to San Francisco Bay, with the species in the Bay most likely native to the northwestern Atlantic (Cohen and Carlton 1995).

Pathways of Introduction Invertebrates cover a diverse range of organisms in many phyla and have arrived in new locations through a range of transportation methods. Globally, the most important pathways for introductions of marine and estuarine species include fouling organisms attached to ships; ballast water; and species imported for aquaculture, fishing bait, or the aquarium trade (Ruiz et al. 1997). Ballast water refers to water contained in tanks within ships to stabilize them (Resources Agency 2008). The same pathways brought species to California (Resources Agency 2008). After the discovery of gold in 1849, San Francisco Bay filled with ships from across the globe—the beginning of its international shipping trade. An estimated 9.1 million metric tons of ballast water was released in California harbors in 2005 (Resources Agency 2008). In 1999, California law began mandating ballast water management, requiring ships arriving from foreign ports to empty and then refill ballast tanks in the open ocean before entering ports (Resources Agency 2008). Commercial shipping in San Francisco Bay may indirectly affect neighboring bays by providing a source of introduced species that are then moved regionally (Wasson et al. 2001). Recreational boats may unintentionally transport attached species between bays (Wasson et al. 2001; Davidson et al. 2010). Intentional imports of some species may transport many others as hitchhikers. In San Francisco Bay, oysters were the most serious example of

Ecology: Organisms

 

Table 9.1 Some Common Invertebrates in San Francisco Bay Salt Marshes That Are Known or Assumed to Be Introduced

Scientific name (common name)

Native range

Method of introduction

Annelida Heteromastus filiformis

W. Atlantic

Oysters or ballast water

Nereis succinea (pile worm)

W. Atlantic

Oyster cultivation

Polydora cornuta

W. Atlantic

Multiple introductions?

Pseudopolydora kempi

W. Pacific? 

Oysters or ballast water

Streblospio benedicti

W. Atlantic, Europe

Oyster cultivation, ship fouling, ballast water

Japan

Ballast water

a

Mollusca Corbula amurensis (Asian clam) Gemma gemma (gem clam)

W. Atlantic

Oyster cultivation

Geukensia demissa (ribbed mussel)

W. Atlantic

Oyster cultivation

Ilyanassa obsoleta (eastern mudsnail)

W. Atlantic

Oyster cultivation

Macoma balthica / M. petalum (Baltic clam)

NW. Atlantic

Oyster cultivation

Musculista senhousia (Japanese mussel)

Japan

Oyster cultivation

Mya arenaria (soft-shell clam)

W. Atlantic

Oyster cultivation

Myosotella myosotis

N. Atlantic

Oyster cultivation or ballast water

Mytilus trossolus / M. galloprovincialis

Europe

Hybridb

Urosalpinx cinerea (Atlantic oyster drill)

W. Atlantic

Oyster cultivation

Venerupis philippinarum (Japanese littleneck clam)

Japan

Oyster cultivation

Carcinus maenas (European green crab)

Europe

Bait worms?

Grandidierella japonica

Japan

Oysters or ballast water

Sphaeroma quoyanum

Australia, New Zealand

Ship fouling

Crustacea

source: Compiled from Brusati 2004, Cohen and Carlton 1995, Carlton 2007, and Neira et al. 2007. Many more species are described in detail by Cohen and Carlton (1995). a

Listed as introduced in Cohen and Carlton 1995, but Carlton 2007 suggests it may not be.

b

Native M. trossolus hybridizes with European M. galloprovincialis.

this. Attempts to establish an industry of eastern oysters (Crassostrea virginica) in the Bay, starting in 1869, brought in a host of other species unintentionally (Carlton 1979; Ruiz et al. 1997). (Meanwhile, overharvesting, along with other anthropogenic impacts, decimated native Olympia oysters, Ostrea lurida.) Species that likely arrived in oyster shipments from the Atlantic coast of North America include the Atlantic mudsnail

Ilyanassa obsoleta, oyster drill Urosalpinx cinerea, clams Gemma gemma and Mya arenaria, ribbed mussel Geukensia demissa, and polychaete worms Streblospio benedicti and Polydora ligni (Carlton 1979; Cohen and Carlton 1995; Miller et al. 2007; Hoos et al. 2010). I. obsoleta is described in more detail later in this chapter. Shipments of Pacific oysters (Ostrea gigas) from Japan, meanwhile, brought the Japanese littleneck clam, Venerupis

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philippinarum, and Japanese mussel, Musculista senhousia (Carlton and Cohen 2007). Of 93 species of bivalves and gastropods that arrived with oysters between 1869 and1940, 12 successfully established populations (Miller et al. 2007). Species most likely to survive were abundant in the source region of the oysters, were tolerant of low salinity, and reproduced by direct development (Miller et al. 2007). These conclusions held true in comparisons with other estuaries. Other pathways of unintentional introductions include fishing bait, the aquarium trade, and live seafood. Bait worms imported from the Atlantic coast also transport other species. Bait and pile worms are packed in seaweed that contains many nonnative invertebrates (Resources Agency 2008). Juvenile European green crabs, Carcinus maenas (described below), may have reached California this way (Carlton and Cohen 2007). Live rock (coral) imported for saltwater aquaria may also harbor introduced species that could be released (Resources Agency 2008).

General Patterns Introduced species tend to be generalists that can survive a wide range of habitat conditions. Introduced invertebrates reaching the Bay have already survived harsh conditions during transport in ballast water tanks, in seaweed, or attached to ships. After arriving, they must contend with other difficult conditions. During a 10-year study on mudflats in the southern Bay, three introduced species dominated the benthic invertebrate assemblage: the gem clam (Gemma gemma), an amphipod (Ampelisca abdita), and a polychaete worm (Steblospio benedicti) (Nichols and Thompson 1985). Opportunistic reproductive strategies and tolerance of minor disturbance to the mudflat habitat appear to give these species an advantage in this habitat (Nichols and Thompson 1985). Invasive species may also be more tolerant to pollution. In one study of fouling species, introduced species were better able to tolerate exposure to copper sulfate than were native species (Crooks et al. 2011). Conversely, native species adapted to a particular set of physical conditions may not be able to adjust to changes such as have occurred in the Bay. Freshwater inflow into the Bay can vary by a factor of 10 among years, because of both precipitation differences and water diversions, leading to 138



fluctuations in salinity (Peterson and Vayssieres 2010). Species that tolerate these fluctuations have a better chance of survival. Twenty-seven years of monitoring in benthic habitats of the upper San Francisco Estuary (San Pablo Bay to the Lower Sacramento River) showed that shifts in invertebrate assemblages correlate with changes in hydrologic years and salinity (Peterson and Vayssieres 2010). Benthic invertebrates shift down the Estuary in years with more freshwater flow from the Delta, and up the Estuary in years with less flow (Peterson and Vayssieres 2010). Benthic habitats in San Pablo Bay are dominated by introduced clams Corbula amurensis, Macoma balthica, and Mya arenaria (Poulton et al. 2004). Introduced species that rely on hard substrate may also benefit from changes that have occurred in the Bay. Hard substrate is more prevalent now than before the Bay was developed, thanks to concrete walls lining marshes, tires scattered across the mudflats, nearby breakwaters, and so on. While there probably always was some hard substrate, such as logs washed into the marshes, the armoring of many shorelines has opened up habitat for species that require hard substrate.On the other hand, such structure provides habitat for some native species, such as Olympia oysters, bay mussels, and algae. Around the Bay, there are efforts to review existing structures and remove hard substrate where possible and to move toward using more natural materials and structures (such as living seawalls) for shoreline protection. Such efforts may help reduce habitats favored by some invasive invertebrates. Many introduced species fail to establish populations in their new locations. One example of a species that has not established populations despite repeated introductions is Littorina littorea, a marine snail native to Europe. L. littorea had been found in San Francisco Bay and other bays on the West Coast occasionally for several decades; the first sizable populations were found at two sites within the Bay in 2002 (Chang et al. 2011). All individuals were adults, with no evidence of recruitment into the population. Genetic analysis comparing these populations to natives and other introduced species indicated that the snails came from the East Coast, rather than from their native range in Europe. This species is sold live in Asian food markets, and the locations where they were found were near markets or bait shops (Chang et al. 2011).

Ecology: Organisms

Reproduction seems to be the main factor restricting its spread. L. littorea is a broadcast spawner; its planktonic larvae would need to settle near other snails in order to establish a population (Chang et al. 2011). However, time will tell whether continued propagule pressure can overcome the factors restricting the snail’s establishment. Some introduced species show a “lag phase” where they are present at low densities for many years before suddenly expanding populations. This example shows that even when physical conditions seem favorable, other factors may restrict introduced species from establishing populations. Now that introduced species dominate the estuarine habitats of the Bay, it is difficult to discern what existed prior to its urbanization. Did introduced species replace similar native ones, or did they fill an empty ecological niche? It is likely that both occurred, although only a few species have been studied in sufficient detail to provide clear conclusions. Specific examples will be described later in this chapter. Introduced species now represent much of the available prey for shorebirds, waterfowl, and fish that forage along marsh edges and in mudflats. Macoma petalum (M. balthica) now forms a significant part of the diet of fish, diving ducks, and California clapper rails (Rallus longirostris levipes) (Cohen and Carlton 1995), but in some parts of the Bay it has been overtaken by another invader, Corbula amurensis (Richman and Lovvorn 2004; Poulton et al. 2004). Ilyanassa obsoleta, Gemma gemma, Neanthes succinea (polychaete), Geuken­ sia demissa, and Mya arenaria are major food items for shorebirds (Recher 1966). One study of clapper rail diets lists three introduced species (Macoma petalum, Geukensia demissa, and Ilyanassa obsoleta) and only one native species (Hemigrapsus oregonensis) as major prey items (Moffitt 1941). These results may not indicate deliberate selection of these introduced invertebrates but may instead reflect their abundance and availability. One of the stranger recorded impacts is that of ribbed mussels, Geukensia demissa, on the California clapper rail (Rallus longirostris obsoletus), a bird that nests within marshes and forages in creeks and on mudflats. Many clapper rails in San Francisco Bay are missing toes; this has been attributed to the mussel’s valves catching the feet of rails as the birds walk over them (Moffitt

1941; Cohen and Carlton 1995). One captured rail lost part of its beak when a mussel attached to it (Cohen and Carlton 1995). However, rails also eat the mussels (Moffitt 1941). The examples below describe in more detail some specific impacts caused by introduced invertebrates and their interactions with native and introduced species.

Ecosystem Engineers Ecosystem engineers are species that change the physical environment in such a way that it influences the space, food, or other resources available to other species either directly or indirectly (Jones et al. 1997). These changes may either benefit or harm native or introduced species. Introduced species may serve as ecosystem engineers themselves or respond to the effects of other ecosystem engineers. One such ecosystem engineer is the isopod Sphaeroma quoyanum, a small crustacean with potentially large impact on the salt marsh habitat. First recorded in California in 1893, it has spread to several estuaries. Native to New Zealand and Australia, it may have arrived burrowed into the wooden hulls of nineteenth-century ships (Cohen and Carlton 1995). S. quoyanum burrows into walls of tidal creeks and at the edge of marshes. Its numerous burrows weaken tidal banks, making the banks more prone to slumping. In San Francisco Bay marshes, Sphaeroma can occur in densities of up to 6,000 per square meter in December and 11,000 per square meter in July, and most sites have at least 34% cover of burrows (Talley et al. 2001). The positive relationship between the number of isopod burrows and the amount of undercutting on creek banks suggests that this isopod contributes to marsh erosion (Talley et al. 2001). Experiments within the marsh showed that experimental cages with isopods caused a significant loss of sediment. However, isopods are only one of several factors causing erosion, exacerbating other effects such as the wakes from ferries and other boats (Talley et al. 2001). Given that most of the marsh area in the San Francisco Estuary has already been destroyed, the invasion of this species may contribute to continued degradation of the remaining habitat. Some introduced species create hard substrate, such as the dense beds of the mussel Geukensia

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demissa, which form around the roots of cordgrass. The creation of hard substrate may in turn facilitate other invaders. In an unusual example of morphological plasticity, the nonnative bryozoan Schizoporella errata, a common fouling species that normally grows as single or multilayer crusts on hard substrate, forms free-living ballshaped colonies and reef-like structures on the mudflats in South San Francisco Bay (Zabin et al. 2010). These structures in turn provide threedimensional hard substrate for dozens of other native and nonnative species, many of which are hard-substrate dependent. Introduced species interact with other native and nonnative organisms through competition, predation, and facilitation. A few species in the Bay have been studied well enough and early enough in their invasion that some of the mechanisms behind their impacts can be seen.

Competition and Predation The invasion of Ilyanassa obsoleta into San Francisco Bay marshes is an example of competition and physiological differences interacting to affect behavioral interactions between a native and an introduced species. The native California horn snail, Cerithidea californica, is an herbivore that used to be common in tidal creeks, marsh pannes (shallow ponds), and mudflats along the Pacific coast. In San Francisco Bay, unlike the rest of its range, C. californica is confined to marsh pannes (Race 1982). I. obsoleta, an Atlantic scavenger and facultative omnivore, arrived in San Francisco with shipments of oysters from the East Coast in the early 1900s (Carlton 1979). I. obsoleta also inhabits tidal creeks and mudflats (Race 1982). However, the habitat use of the two snails varies seasonally. Both snails are dormant during the winter, C. californica burying itself in the mud under vegetation while I. obsoleta stays in the mudflats. In the spring, C. californica moves into tidal creeks, but by late summer I. obsoleta is in the creeks and C. californica remains only in the marsh pannes. Through a combination of field and laboratory observations, Race (1982) concluded C. californica starts in creek banks when it emerges from winter dormancy, but after I. obsoleta emerges, competition between the two snails confines C. californica to marsh pannes. Competition for space may be exacerbated by direct predation: I. obsoleta 140



will eat the eggs of C. californica, while the native snail, as an herbivore, does not consume those of I. obsoleta (Race 1982). Physiological differences also affect the balance of competition between the snails. During very high tides, I. obsoleta was observed moving into the marsh pannes usually inhabited by C.  californica. When the tides receded, many of the I.  obsoleta marsh pannes died, apparently from desiccation and thermal stress. This hypothesis was supported by laboratory experiments where C.  californica survived as long at 15 days out of water while no I. obsoleta survived past 2 days (Race 1982). More recent work suggests that another dynamic may influence this interaction, with an addition of nitrogen changing the populations of diatoms that the snails consume and, differentially, the two snails’ survival (H. Weiskel, unpublished data). The European green crab (Carcinus maenas) is one of the few introduced invertebrates in the Bay that has been tracked and studied since early in its invasion. Green crabs became abundant in the Bay within a year of their first detection in 1990 (Cohen et al. 1995). Omnivorous predators, adult green crabs grow much larger than the native crabs Hemigrapsus oregonensis and Pachygrapsus crassipes.Green crabs are native to Europe and have also invaded the east coast of the United States. The method of introduction to California is unclear. Larvae might have been transported in ballast water, and juvenile crabs may have been transported in the seaweed used to pack live bait worms or lobsters shipped from the East Coast. Genetic studies indicate that California populations originated in eastern North America and that a single introduction into San Francisco Bay formed the genesis of all invasive populations on the west coast (Geller et al. 1997; Tepolt et al. 2009). Green crabs inhabit protected marine and estuarine environments such as mudflats, eelgrass (Zostera marina) beds, and salt marshes (Cohen et al. 1995). Protected lagoons within the Bay may have provided “incubators” to allow populations to establish and spread around the Bay (Cohen et al. 1995). Green crabs have high fecundity, producing up to 200,000 eggs per mature female annually in the native range (Cohen et al. 1995). Green crabs’ greatest effect may be their impact on the food web as a result of the breadth and quantity of prey they consume, which ranges from algae

Ecology: Organisms

to worms (polychaetes and oligochaetes) to small fish. The crabs’ diets encompass 158 different genera representing 5 plant or protist phyla and 14 animal phyla (Cohen et al. 1995). While no studies have examined the specific effects of green crabs in San Francisco Bay, the crabs are implicated in declines in 20 invertebrate species in Bodega Bay, 100 miles to the north (Grosholz et al. 2000). Within 3 years of the green crab’s introduction in Bodega Bay, the native clams Nutricola confusa and Nutricola tantilla, as well as the native crab Hemigrapsus oregonensis, declined by 80%–90%. Subsequently, the numbers of other organisms that compete with clams for space increased, indicating an indirect effect as these organisms were released from competition with the crabs’ prey. However, shorebird numbers did not change during this time period, indicating that changes in the birds’ prey base did not seem to affect the birds themselves. Interestingly, predation by larger native crabs (Cancer spp.), in the form of limb damage to green crabs, appears to restrict C. maenas’ distribution in estuaries in northern California estuaries (Jensen et al. 2007), although this has not been tested for San Francisco Bay. Jensen et al. (2007) concluded that green crabs living within marsh experience less predation than those on mudflats. This could potentially increase predation pressure from green crabs on species within marshes if the crabs are restricted there. The Asian clam Corbula (formerly Potamocorbula) amurensis is another invader with wide-ranging effects. First discovered in the San Francisco Estuary in 1986, it soon became the numerically dominant species in the northern estuary from San Pablo Bay to the Lower Sacramento River (Carlton et al. 1990; Peterson and Vayssieres 2010). It is now the main prey for wintering ducks. Greater and lesser scaup (Aythya marila and A. affinis) consume many C. amurensis but fewer M. balthica than they did previously. This may result from the abundance of C. amurensis and its location at a shallow depth in the sediment that makes it more accessible to scaup than are other, more deeply buried clams (Poulton et al. 2004; Richman and Lovvorn 2004). Compared with M. balthica, C. amurensis clams and shells have higher energy content, higher nitrogen content, and better digestibility, but its shells are more difficult for ducks to crush and it accumulates higher levels of toxic contaminants  

(Richman and Lovvorn 2004). Predation by wintering scaup may cause C. amurensis populations to decline between fall and spring. C. amurensis is also an example of a species able to tolerate the Bay’s fluctuating environment. Its initial invasion occurred at the beginning of a multiple-year drought. C. amurensis tolerates a wide range of salinities, which may have helped it establish populations (Carlton et al. 1990). Its invasion resulted in significant changes in benthic assemblages, including the displacement of most other common species, many of which were themselves introduced (Nichols et al. 1990). Species richness increased and several other introduced species became more abundant (Peterson and Vayssieres 2010). In addition, direct predation by large numbers of C. amurensis as they filter feed Bay water has greatly reduced zooplankton abundance (Kimmerer et al. 1994; Peterson and Vayssieres 2010).

Interactions among Invaders Invertebrates respond to the habitat structure within marshes. The roots of salt marsh provide attachment sites, while aboveground vegetation reduces desiccation at low tide by shading the soil (Pennings and Bertness 2001). Many benthic invertebrates feed on the detritus produced as vegetation decays. Therefore, changes in the amount or type of vegetation of a marsh can have significant impacts on invertebrate communities. The spread of hybrid cordgrass (Spartina alterniflora × foliosa) during the latter part of the twentieth century transformed marshes and mudflats in San Francisco Bay. Its removal may create even more alterations. Starting in the 1970s, salt marshes and mudflats in San Francisco Bay were invaded by Atlantic cordgrass (Spartina alterniflora), which interbred with native Pacific cordgrass (Spartina foliosa) to create hybrid S. alterniflora × foliosa (Ayres et al. 2004). The hybrid replaces Spartina foliosa, grows up into the pickleweed (Sarcocornia pacifica, formerly Salicornia virginica) zone, and spreads across previously unvegetated mudflats. In contrast to Pacific cordgrass, hybrid Spartina produces taller stems and greater plant biomass both above- and belowground while occupying a much wider tidal range, thereby transforming open mudflats to vegetated habitat (Brusati and Grosholz 2006).

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The large changes hybrid Spartina brings to marshes and mudflats suggest that it could change the composition of the invertebrates in the marshes, alter the flow of vegetative material to the food web, and perhaps facilitate other introduced species, such as those from S. alterniflora’s native range. S. foliosa marshes contain significantly higher densities of benthic infauna than adjacent mudflats, while hybrid Spartina areas do not (Brusati and Grosholz 2006). Overall macrofaunal invertebrate densities can be 75% lower in hybrid Spartina sediments than in adjacent mudflats, while biomass is 57% lower (Neira et al. 2006). S. foliosa appears to produce a moderate level of structure that can facilitate benthic invertebrates, whereas hybrid Spartina’s dense roots appear to exclude them (Brusati and Grosholz 2006). The large number of cryptogenic species, as well as the difficulty of identifying many of the smallest but most abundant infaunal species, makes it difficult to determine whether hybrid Spartina facilitates introduced species at the expense of native ones or whether the observed patterns depend on other factors. Hybrid Spartina may facilitate certain introduced invertebrates. For instance, green crab densities are five times higher in hybrid Spartina than in adjacent mudflats (Neira et al. 2005). Macoma petalum, Mya arenaria, Geukensia demissa, Ilyanassa obsoleta, and Urosalpinx cinerea are also more abundant at the edge of the marsh than on mudflats (Gros­ holz et al. 2009). However, it is unclear whether the organisms respond to the presence of vegetation, which may provide protection from desiccation and a refuge from predators, or whether hybrid Spartina itself provides a unique environment that is attractive to these species. Many of these species are native to the Atlantic coast of North America, so coevolution with S. alterniflora may make them better suited to the cordgrass in its new range. One of the few documented insect introductions to San Francisco Bay marshes is Trigonotylus uhleri, an herbivore specialist on S. alterniflora in its native range, now often found on Pacific cordgrass in San Francisco Bay. It was first collected in the 1990s and probably arrived with introduced S. alterniflora (Cohen and Carlton 1995). Structural changes brought about by this ecosystem engineer, specifically the increased biomass and subsequent increased production of 142



detrital material, do not necessarily make hybrid Spartina a major contributor to marsh food webs (Brusati and Grosholz 2009). Neither native nor hybrid Spartina appears to be a significant food source for M. petalum in the Bay (Brusati and Grosholz 2009). Experiments with isotope tracers show that some amphipods and polychaetes ingest hybrid Spartina but that these differ from the organisms that feed on S. foliosa (Neira et al. 2006). A major eradication effort by the State Coastal Conservancy has greatly reduced the extent of hybrid cordgrass, raising the question of how the infauna will respond as the marshes change yet again.

Ongoing Invasions New species continue to enter the Bay. One recent invader is the Japanese mud snail Batillaria attramentaria, discovered in 2003 at a marina in San Pablo Bay (H. Weiskel, personal communication). It has been present in estuaries north and south of the Bay for several decades, giving a possible preview to its effects in San Francisco Bay. B. attramentaria has replaced or appears in the process of replacing C. californica in several northern California estuaries (Byers 2000). While some research from Tomales Bay, California, suggested that B. attramentaria and C. californica can coexist (Whitlatch and Obrebski 1980), other studies found a marked decline in C. californica, attributed in part to the introduced snail’s ability to withstand hypoxia (Byers 2000). C. californica is a host of numerous trematode parasites, which provides linkages in the food web, while B. attramentaria hosts only one of these parasites, so the decline of the native snail could have broader implications (Lafferty and Kuris 2009). Changes in the number and types of parasites can alter the behavior of parasite hosts, with potential effects on predator-prey interactions elsewhere in the food web (Lafferty and Kuris 2009). In an effort to eradicate the new invasion in San Pablo Bay, teams of volunteers removed the snails by hand (Prado 2008). B. attramentaria is capable of floating to new locations (Whitlatch and Obrebski 1980), so only time will tell whether this laborintensive strategy was effective. Most of the introduced invertebrates in San Francisco Bay are here to stay. They join other ongoing perturbations to the physical and bio-

Ecology: Organisms

logical environment of the Bay. Numerous marsh and other estuarine restoration projects are in progress around the Bay, such as the South Bay Salt Pond Restoration Project (the largest tidal wetlands restoration project on the West Coast, www.southbayrestoration.org) and attempts to reestablish native oyster beds. In terrestrial environments, restoration often incorporates eradicating unwanted introduced species. Few large-scale control efforts have been attempted in marine and estuarine environments. In California, two invasive algae were eradicated in the early 2000s: Caulerpa taxifolia in two Southern California harbors (Anderson 2005) and Ascophyllum nodosum in San Francisco Bay (Miller et al. 2004), although A. nodosum has since returned. Eradicating invertebrates is difficult because of their small size and their reproductive strategies, which often include far-traveling larvae. Just as the effects on native fauna due to many introduced invertebrates are unclear, so are the potential effects of removing the introduced invertebrates now that they dominate the San Francisco Estuary. As for any other form of introduced species, preventing new invasions is more cost-effective than attempting large-scale removal of established populations (Anderson 2005; Williams and Grosholz 2008). Efforts to address pathways such as ballast water and hull fouling, along with increased public education, may slow the entry of invasive invertebrates but probably cannot stop the flow entirely.

Synthesis and Future Directions In general, most introduced species have little impact on their new environment, and only a small number become truly invasive with serious negative effects. It is likely that the introduced invertebrates of San Francisco Bay follow a similar pattern. However, determining which species will cause problems and how to address them raises both scientific and natural management questions.This chapter has discussed a few well-studied introduced invertebrates that can be categorized as ecosystem engineers, competitors, or predators and that have demonstrated negative effects on native species. The lack of knowledge about invertebrate communities in the Bay makes it difficult to know just how much of an impact these species have had. Invertebrates tend to be less visible than other introduced species such as

plants or fish.The human changes to the physical and biological environment of San Francisco Bay discussed in earlier chapters have influenced the types of species that have invaded and the ways in which they have invaded. This will continue as climate change drives species into new interactions. Many scientific and resource management questions remain unanswered. For instance, which factors most influence the abundance and spread of introduced invertebrates? Existing research has focused mostly on direct changes caused by a few easily studied species. Questions about more indirect influences are more difficult to examine experimentally but may be valuable, especially now that so many introduced species are established and abundant. Another area for future research is predicting which species that are problem invaders in other parts of the world are likely to reach San Francisco Bay, what pathways are most likely to bring them, and which species should be cause for concern. A review of other estuaries that have similar climates and are connected through potential pathways of introduction might provide valuable information. From a management perspective, questions include which species are worth trying to control, what forms of control are most effective, and which species or invasion pathways should be targeted for early detection. Ballast water is one major known pathway of introduction and is starting to be addressed. Where is the balance between controlling invasive species that are here already and stopping future invasions? The answers to these questions may require policy decisions at the state or national level. Acknowledgments Thank you to Chela Zabin of the Smithsonian Ecological Research Center for writing the section on invaders of hard substrates and providing helpful comments on an earlier draft, and to Heidi Weiskel at the University of California–Davis for sharing her dissertation research. The comments of three anonymous reviewers also improved the manuscript.  

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Ayres, D. R., D. L. Smith, K. Zaremba, S. Klohr, and D. R. Strong. 2004. Spread of exotic cordgrasses and hybrids (Spartina sp.) in the tidal marshes of San Francisco Bay, California, USA. Biological Invasions 6: 221–231. Brusati, E. D. 2004. Effects of native and hybrid cordgrass on benthic invertebrate communities and food webs. PhD dissertation, University of California–Davis. Brusati, E. D., and E. D. Grosholz. 2009. Does invasion of hybrid cordgrass change estuarine food webs? Biological Invasions 11:917–926. Brusati, E. D., and E. D. Grosholz. 2006. Native and introduced ecosystem engineers produce contrasting effects on estuarine infaunal communities. Biological Invasions 8:683–695. Byers, J. E. 2000. Competition between two estuarine snails: Implications for invasions of exotic species. Ecology 81:1225–1239. Callaway, J. C. and J. B. Zedler. 2009. Salt marsh conservation along the leading edge of the continent. In Human Impacts on Salt Marshes, edited by B. Silliman, E. Grosholz and M. Bertness, 285–307, University of California Press, Berkeley, CA. Carlton, J. T. 1996. Biological invasions and cryptogenic species. Ecology 77:1653–1655. Carlton, J. T. 1979. Introduced invertebrates of San Francisco Bay. In San Francisco Bay: The urbanized estuary, edited by T. J. Conomos, 427–444. Pacific Division of the American Association for the Advancement of Science, San Francisco. Carlton, J. T., ed. 2007. The Light and Smith manual: Intertidal invertebrates from central California to Oregon. 4th ed. University of California Press, Berkeley. Carlton, J. T., and A. N. Cohen. 2007. Introduced estuarine and marine invertebrates. In The Light and Smith manual: Intertidal invertebrates from central California to Oregon, 4th ed., edited by J. T. Carlton, 28–31. University of California Press, Berkeley. Carlton, J. T., J. K. Thompson, L. E. Schemel, and F. H. Nichols. 1990. Remarkable invasion of San Francisco Bay (California, USA) by the Asian clam Potamocorbula amurensis. I. Introduction and dispersal. Marine Ecology Progress Series 66:81–94. Chang, A. L., A. M. H. Blakeslee, A. W. Miller, and G. M. Ruiz. 2011. Establishment failure in biological invasions: A case history of Littorina littorea in California, USA. PLoS ONE 6(1):e16035. Cohen, A. N., and J. T. Carlton. 1998. Accelerating rate of invasion in a highly invaded estuary. Science 279:555–558. Cohen, A. N., and J. T Carlton. 1995. Nonindigenous aquatic species in a United States estuary: A case study of the biological invasions of the San Francisco  

 

 

 

 

 

 

 

 

 

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Ecology: Organisms

chapter Ten

Invertebrates A Case Study of China Camp State Park, Marin County April Robinson, Andrew N. Cohen, Brie Lindsey, and Letitia Grenier

CONTENTS

tion of physiological limitations and ecological interactions (Tomanek and Helmuth 2002). Within tidal marshes, distinct subhabitats— from large, high-order channels to small, loworder channels, to marsh plain and natural levee—are found adjacent to each other along the tidal gradient, sometimes changing abruptly from one meter to the next. Marsh invertebrate communities vary by subhabitat, with many species showing a preference for particular elevations, vegetation zones, or substrate types (Teal 1962; Davis and Grey 1966; Levin and Talley 2000). Invertebrates constitute much of the secondary productivity in tidal marshes (Teal 1962) and play a critical role in transferring primary productivity up the food web, forming a substantial part of the diet of many resident marsh vertebrates (Grenier and Greenberg 2005). As there are few seeds and fruits in the marsh available for terrestrial vertebrates to forage on (Greenberg et al. 2006), the distribution and diversity of invertebrates largely determines the food resources available for secondary consumers and influences their foraging behaviors. Invertebrates constitute a substantial portion of the diets of many common marsh fish species as well (Visintainer et al. 2006). This chapter provides original data on the distribution of macroinvertebrates across a tidal gradient and reviews what is known about the diversity, distribution, and abundance of inter-

Tidal Marsh Invertebrate Study Rocky Intertidal Invertebrates Invertebrate Diversity Invertebrate Distribution Invertebrates as Food Resources Synthesis and Future Directions

 

 

I

ntertidal habitats present a harsh physical environment for resident invertebrates. Twice-daily tides subject terrestrial invertebrates to the risk of drowning and aquatic invertebrates to the risk of desiccation. Inundation periods and sediment properties vary across the intertidal gradient, and environmental conditions change rapidly with inundation and exposure. Physical and biological conditions change over small spatial scales, as slight changes in elevation translate to large changes in hydrology, geomorphology, and vegetation (Collins et al. 1986; Pennings and Callaway 1992). The distribution of rocky intertidal invertebrates varies over both large and small spatial scales as a result of differences in dispersal, recruitment, and response to changes in microhabitat among species (Underwood and Chapman 1996). The small-scale zonation of rocky intertidal invertebrates results from a combina-





147

tidal invertebrates at China Camp State Park in Marin County, California, a National Estuarine Research Reserve site. Data from two studies, one of tidal marsh invertebrates and the other of rocky intertidal invertebrates, are presented here (Robinson et al. 2011). Most of the previously available invertebrate data from China Camp focus on predation of invertebrates (Dean et al. 2005; Visintainer et al. 2006) rather than on their diversity and distribution. The implications of invertebrate distribution and diversity on the behavioral ecology of their predators is also briefly discussed. The data presented in this chapter demonstrate the unequal distribution of invertebrates across intertidal subhabitats at China Camp State Park in San Francisco Bay. Relatively few species made up the majority of the invertebrate biomass in the tidal marsh, and the majority of both the rocky intertidal invertebrates and the tidal marsh invertebrates identified to species were exotic. The strong association of certain invertebrate groups to specific subhabitats suggests that predators with different feeding specializations may forage primarily in one part of the marsh or another.

Tidal Marsh Invertebrate Study Invertebrates at China Camp marsh were collected from the channels, marsh plain, and natural levees as part of a food web study reported in greater detail by Grenier (2004). Invertebrates were collected to investigate which taxa were available as potential prey items for the San Pablo song sparrow (Melospiza melodia samuelis), a tidal marsh obligate, and other marsh vertebrates and to determine how macroinvertebrates were distributed across the tidal gradient. Because no single method was sufficient to account for all invertebrate locomotion types and habitat preferences, multiple trapping methods were used. The study was conducted in a 3.3 ha plot within the portion of the marsh that has accreted since the mid-1800s, characterized by the simple, less sinuous channels typical of a rapidly formed marsh. Sampling was conducted at low tide from May to July 2001 and consisted of five capture methods: pit trap, sweep net, snail count, mud core, and sticky trap. Equal sampling effort was expended along high-order and low-order chan148



nels. For each channel type, random sampling locations were stratified across three subhabitats: within the channel, on the natural levee adjacent to the channel, and on the nearby marsh plain. No samples were taken in standing water. The plant species within 10 cm of each trap were recorded. Pit trap, sweep net, and snail count methods were conducted with equal effort in each of the subhabitats. Pit traps were cylindrical plastic containers, 11 cm in diameter and 11 cm deep, buried in the sediment with the top of the trap level with the ground and no space between the container and the surrounding sediment. Traps were open for at least 3 hours. Sweep net sampling consisted of 10 strokes with a 15-inch-diameter sailcloth net, sweeping new vegetation with each stroke. Snail counts consisted of counting all snails within a 22 cm × 22 cm quadrat. Mud core and sticky trap methods were used only in the channels because (1) on the natural levees and marsh plains, pilot mud core samples consisted of dry, hard-packed sediment devoid of macroinvertebrates and (2) pilot sticky trap samples replicated results from pit traps and sweep nets. Cores were 7.0 cm in diameter and 10 cm deep, and organisms were collected from them with a 0.5 mm mesh sieve. For each core, the relative abundance of roots was recorded on a scale of 0–3, with 0 indicating no roots and 3 indicating very dense roots. Each sticky trap was a thin layer of Tanglefoot adhesive spread onto a 20 cm × 10 cm sheet of plastic that was placed on the sediment. The traps were set for at least 3 hours and checked frequently as the tide rose; if the traps were in jeopardy of flooding, they were moved to adjacent higher ground. Common invertebrates were identified to the lowest feasible taxonomic level with assistance from experts (see Acknowledgments). Average biomass was determined for large or common taxa (greater than 10 individuals per trap method) by weighing between 9 and 115 individuals per taxon, after drying at 55° C until a constant weight was achieved. Snails were weighed without their shells. Because planthoppers (Prokelisia marginata) had such low mass, they were weighed in groups of 10 individuals at a time. Masses for araneid spiders were estimated from lycosid spiders of similar size. Catch per unit effort (CPUE) was calculated as the number of invertebrates of the same taxon  

Ecology: Organisms

caught per trap hour for pit traps and sticky traps, and it was calculated as invertebrates per trapping event for all other capture methods. Differences in CPUE among subhabitats were examined using nonparametric ANOVA (­K ruskal-Wallis), which was also used to determine the relationship between CPUE and presence of roots, and CPUE and plant community composition. The relationship between CPUE and plant community composition was examined separately for each of the subhabitats along the tidal gradient because the vegetation varied dramatically among subhabitats. Plant-invertebrate relationships in the channel subhabitat were tested separately for large and small channels because Spartina foliosa was found only in large channels. A total of 4,597 invertebrates were captured in 787 trapping events, representing 7 taxonomic classes and at least 14 orders (Table 10.1). Six of the 7 taxa identified to species (85.7%) were exotic (most of the arthropods were not identified to species). As expected, community composition of invertebrates differed notably by capture method, and one taxon usually dominated captures for each trapping method. The amphipod Traskorchestia traskiana comprised 77% of the individuals caught by pit trap, while the planthopper Prokelisia marginata comprised 64% of the individuals caught by sweep net. Oligochaete and polychaete worms made up 67% of mud core captures, and dolichopodid flies made up 83% of individuals caught by sticky trap. The abundance of common taxa differed by subhabitat (Table 10.2). Channel size also influenced invertebrate community composition, with several common taxa being more abundant near either low-order or high-order channels (Table 10.3). The burrowing amphipod Corophium alienense was the only species whose abundance was related to the density of plant roots, being more abundant in areas with lower root density (Kruskal-Wallis, H = 14.57, n = 72, p = 0.02). The abundance of Corophium amphipods, Macoma petalum clams, and Prokelisia planthoppers was related to plant distribution. The burrowing amphipods and clams were more likely to be found in large channels where Spartina foliosa was not present (C. alienense: Mann-Whitney, U = 223.5, n = 36, p < 0.001; M. petalum: Mann-Whitney, U = 223, n = 36, p < 0.001), while planthoppers were more likely to be found in channels

where S. foliosa was present (Mann-Whitney, U = 94, n = 36, p < 0.001). Mass (+ 1 SD) of common taxa ranged from 0.26 (+ 0.07) mg/individual for Prokelisia marginata to 8.65 (+ 6.70) mg/individual for Traskorchestia traskiana. Pit trap biomass was dominated by one species across all subhabitats, while sweep net biomass was dominated by a different taxon in each subhabitat. Average sweep net biomass per trapping event was calculated to be greatest in the channel subhabitat (5.2 mg), followed by the natural levee (4.6 mg), and lowest in the marsh plain (3.4 mg). P. marginata constituted the largest fraction of sweep net biomass in the channel subhabitat (2.5 mg, 47.6%), while spiders of the family Araneidae made up the greatest percentage of sweep net biomass on the natural levee (2.7 mg, 58.8%). Sweep net biomass on marsh plain was not clearly dominated by any one taxon. Sweep net biomass was greater near high-order than low-order channels (5.9 mg and 2.9 mg, respectively). Pit trap biomass was dominated by T. traskiana, which accounted for 94.3% of the biomass. Average pit trap biomass per trap hour was higher near low-order channels than near high-order channels (10.7 mg vs 5.1 mg). Average pit trap biomass per trap hour was highest in the marsh plain subhabitat (15.0 mg), followed by the natural levee (9.8 mg), and lowest in the channel subhabitat (3.1 mg). The mean biomass per quadrat for the snail Myosotella myosotis was 15.7 mg on the marsh plain and 25.5 mg on the natural levee, with no snails observed in the channels. In addition to the taxa above, several invertebrates that had been seen but not captured during the quantitative tidal marsh study were handcollected for identification. These taxa included the European green crab (Carcinus maenas), the yellow shore crab (Hemigrapsus oregonensis), two species of shrimp (Palaemon macrodactylus and Crangon franciscorum), the eastern mud snail (Ilyanassa obsoleta), stinkbugs in the family Pentatomidae, and mites in the family Tetranychidae. Other invertebrates commonly observed at China Camp (A. Cohen, personal observation) include the isopod Sphaeroma quoiana, whose pencil-diameter burrows riddle the channel banks and may contribute to their slumping and erosion, and the small comensal isopod Iais californica, which lives on Sphaeroma’s ventral surface. Both of these species are from ­Australia.

Invertebrates: A Case Study of China Camp

149

Arthropoda

Mollusca

Oligochaeta

Annelida

Amphipoda

Veneroida

Bivalvia

Crustacea

Basommatophora

Other Polychaeta

Phyllodocida

Order

Gastropoda

Polychaeta

Class

Phylum

Table 10.1

Aquatic amphipod

Grandidierella japonica Traskorchestia traskiana

Corophiidae Talitridae

-

98

195

Aquatic amphipod

Tellinidae

Corophium alienense

1

131

2

655

Mud core

Corophiidae

European marsh snail

Polychaete worm

Polychaete worm

Oligochaete worm

Common name

39

Myosotella myosotis

Allita succinea

Genus and species

Macoma petalum

Ellobiidae

Nereidae

Family

Number of Invertebrates Collected in the Tidal Marsh Study by Each Capture Method Bold indicates exotic species.

3

3

602

-

-

-

-

-

Pit trap

-

-

-

4

22

-

-

-

Sweep net

-

-

-

-

886

-

-

-

Snail count

Number of individuals collected (by capture method)

4

-

-

-

2

-

-

-

Sticky trap

Insecta

Arachnida

Insect

Other Insecta

-

-

-

Moth

Leaf hopper

Planthopper

-

-

-

-

Lepidoptera

Other Homoptera

Delphacidae

Other Diptera

Picture-wing fly

Otitidae

1

Beetle adult Beetle larva

Other Coleoptera Long-legged fly

Spotted cucumber beetle

Chrysomelidae

Dolichopodidae

1 14

Ground beetle

Bembidion

-

Weevil

Curculionidae

41

2

Mud-living beetle

-

-

Heteroceridae

Prokelisia marginata

Wolf spider

Lycosidae Spider

Orb spider

Araneidae

Hemiptera

Homoptera

Diptera

Coleoptera

Other Arachnida

Araneae

4

-

5

-

-

-

-

-

-

1

1

52

-

65

3

7

26

1

7

6

4

9

11

703

86

25

116

-

17

13

-

38

2

44

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

1

99

-

544

-

-

-

-

-

-

1

-

-

Table 10.2 Catch per Unit Effort (CPUE) by Subhabitat for Pit Trap and Sweep Net Samples P-values are from Kruskal-Wallis tests (alpha = 0.05). Bold text indicates the zone with the highest CPUE.

Total count (no. trap hours or trapping events) Capture method

Taxon (order)

Pit trap

Lycosidae (Araneae) Bembidion (Coleoptera)

Sweep net

Channel

Marsh plain

Natural levee

0 (234)

13 (230)

13 (236)

0.001

0 (234)

44 (230)

21 (236)