Climate Change and Estuaries (CRC Marine Science) [Team-IRA] [1 ed.] 0367647524, 9780367647520

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Climate Change and Estuaries Climate change is having an increasing impact on coastal, estuarine, and marine environments worldwide. This book provides state-of-the-art coverage of climate change effects on estuarine ecosystems from local, regional, and global perspectives. With editors among the most noted international scholars in coastal ecology and estuarine science and contributors who are world-class in their fields, the chapters in this volume consist of comprehensive studies in coastal, estuarine and marine sciences, climate change, and coastal management and provide an extensive international collection of data in tabular, illustrated, and narrative formats useful for coastal scientists, planners, and managers. Comprised of three sections: (1) physical-chemical aspects; (2) biological aspects; and (3) management aspects, the book not only examines climatic and non-climatic drivers of change affecting coastal, estuarine, and marine environments but also their interactions and effects on populations of organisms, communities, habitats, and ecosystem structure and function. Pulling together today’s most salient issues and key literature advances for those concerned with coastal management, it allows the reader to see across direct and indirect interactions among disciplinary and ecosystem boundaries. Climate Change and Estuaries meets the research needs of climate scientists, estuarine and marine biologists, marine chemists, marine geologists, hydrologists, and coastal engineers, while students, professors, administrators, and other professionals will also find it an exhaustive reference.

Marine Science Series The CRC Marine Science Series is dedicated to providing state-of-the-art coverage of important topics in marine biology, marine chemistry, marine geology, and physical oceanography. The series includes volumes that focus on the synthesis of recent advances in marine science. CRC MARINE SCIENCE SERIES Series Editors Michael J. Kennish, Ph.D. and Judith S. Weis, Ph.D. PUBLISHED TITLES The Biology of Sea Turtles, Volume I, edited by Peter L. Lutz and John A. Musick Pollution Impacts on Marine Biotic Communities, Michael J. Kennish Coastal Ecosystem Processes, Daniel M. Alongi Intertidal Deposits: River Mouths, Tidal Flats, and Coastal Lagoons, Doeke Eisma Estuary Restoration and Maintenance: The National Estuary Program, Michael J. Kennish Handbook of Marine Mineral Deposits, David S. Cronan Seagrasses: Monitoring, Ecology, Physiology, and Management, edited by Stephen A. Bortone Artificial Reef Evaluation with Application to Natural Marine Habitats, edited by William Seaman, Jr. Handbook for Restoring Tidal Wetlands, edited by Joy B. Zedler Eutrophication Processes in Coastal Systems: Origin and Succession of Plankton Blooms and Effects on Secondary Production in Gulf Coast Estuaries, Robert J. Livingston Ecology of Seashores, George A. Knox Marine Chemical Ecology, edited by James B. McClintock and Bill J. Baker Trophic Organization in Coastal Systems, edited by Robert J. Livingston Estuarine Research, Monitoring, and Resource Protection, edited by Michael J. Kennish Estuarine Indicators, edited by Stephen A. Bortone Practical Handbook of Estuarine and Marine Pollution, Michael J. Kennish Practical Handbook of Marine Science, First Edition, edited by Michael J. Kennish Practical Handbook of Marine Science, Second Edition, edited by Michael J. Kennish Practical Handbook of Marine Science, Third Edition, edited by Michael J. Kennish Practical Handbook of Marine Science, Fourth Edition, edited by Michael J. Kennish Restoration of Aquatic Systems, Robert J. Livingston Chemical Oceanography, Third Edition, Frank J. Millero Coastal Pollution: Effects on Living Resources and Humans, Carl J. Sindermann Acoustic Fish Reconnaissance, I.L. Kalikhman and K.I. Yudanov Coastal Lagoons: Critical Habitats of Environmental Change, edited by Michael J. Kennish and Hans W. Paerl Ecology of Marine Bivalves: An Ecosystem Approach, Second Edition, Richard F. Dame Climate Change and Coastal Ecosystems: Long-Term Effects of Climate and Nutrient Loading on Trophic Organization, Robert J. Livingston Habitat, Population Dynamics, and Metal Levels in Colonial Waterbirds: A Food Chain Approach, Joanna Burger and Michael Gochfeld Living Shorelines: The Science and Management of Nature-Based Coastal Protection, edited by Donna Marie Bilkovic, Molly M. Mitchell, Megan K. La Peyre, and Jason D. Toft Fishes Out of Water: Biology and Ecology of Mudskippers, edited by Zeehan Jaafar and Edward O. Murdy A Blue Carbon Primer: The State of Coastal Westland Carbon Science, Practice and Policy, edited by Lisa-Marie Windham-Myers, Stephen Crooks, and Tiffany G. Troxler Climate Change and Estuaries, edited by Michael J. Kennish, Hans W. Paerl, and Joseph R. Crosswell For more information about this series, please visit: https://www.crcpress.com /CRC-Marine-Science/book-series/CRCMARINESCI?page=&order=pubdate&size=12&view =list&status=published,forthcoming

Climate Change and Estuaries

Edited by Michael J. Kennish, Hans W. Paerl, and Joseph R. Crosswell

First edition published 2024 by CRC Press 6000 Broken Sound Parkway NW, Suite 300, Boca Raton, FL 33487-2742 and by CRC Press 4 Park Square, Milton Park, Abingdon, Oxon, OX14 4RN © 2024 selection and editorial matter, Michael J. Kennish, Hans W. Paerl and Joseph R. Crosswell; individual chapters, the contributors CRC Press is an imprint of Taylor & Francis Group, LLC Reasonable efforts have been made to publish reliable data and information, but the author and publisher cannot assume responsibility for the validity of all materials or the consequences of their use. The authors and publishers have attempted to trace the copyright holders of all material reproduced in this publication and apologize to copyright holders if permission to publish in this form has not been obtained. If any copyright material has not been acknowledged please write and let us know so we may rectify in any future reprint. Except as permitted under U.S. Copyright Law, no part of this book may be reprinted, reproduced, transmitted, or utilized in any form by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying, microfilming, and recording, or in any information storage or retrieval system, without written permission from the publishers. For permission to photocopy or use material electronically from this work, access www​.copyright​.com or contact the Copyright Clearance Center, Inc. (CCC), 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400. For works that are not available on CCC please contact mpkbookspermissions​@tandf​.co​​.uk Trademark notice: Product or corporate names may be trademarks or registered trademarks and are used only for identification and explanation without intent to infringe. ISBN: 9780367647520 (hbk) ISBN: 9780367647537 (pbk) ISBN: 9781003126096 (ebk) DOI: 10.1201/9781003126096 Typeset in Times by Deanta Global Publishing Services, Chennai, India

Contents Foreword................................................................................................................................................................................ix Preface...................................................................................................................................................................................xi About the Editors.................................................................................................................................................................. xv List of Contributors.............................................................................................................................................................xvii

Section 1  Physical–Chemical Aspects Chapter 1 Introduction to Climate Change and Estuaries................................................................................................. 3 Michael J. Kennish, Hans W. Paerl, and Joseph R. Crosswell Chapter 2 Climate Change in the Earth System.............................................................................................................. 23 David R. Easterling and Kenneth E. Kunkel Chapter 3 Estuaries: Origin, Historical Development, and Classifications..................................................................... 43 Michael J. Kennish Chapter 4 Sea-Level Rise and Estuaries.......................................................................................................................... 55 John A. Church and Xuebin Zhang Chapter 5 Anthropogenic Drivers of Estuarine Change.................................................................................................. 75 Michael J. Kennish Chapter 6 Climate Change and Saltwater Intrusion in Estuaries....................................................................................99 Arnoldo Valle-Levinson and Ming Li Chapter 7 Biogeochemical Changes in Estuaries.......................................................................................................... 113 Nicholas D. Ward, Thomas S. Bianchi, Christopher L. Osburn, and Allison Myers-Pigg Chapter 8 Hypoxia and Climate Change in Estuaries................................................................................................... 143 Jeremy M. Testa, Jacob Carstensen, Arnaud Laurent, and Ming Li Chapter 9 Estuarine Acidification under a Changing Climate...................................................................................... 171 Wei-Jun Cai Chapter 10 Global Change and Estuarine Carbon Dynamics......................................................................................... 183 Charles Hopkinson, Nathaniel Weston, and Wei-Jun Cai Chapter 11 Blue Carbon in a Changing Climate and a Changing Context..................................................................... 203 Lisamarie Windham-Myers



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Chapter 12 Effects of a Changing Climate on the Physics of Estuaries.......................................................................... 215 L. Fernando Pareja-Roman and Robert J. Chant Chapter 13 Climatic Drivers of Estuarine Sediment Dynamics...................................................................................... 231 Neil K. Ganju Chapter 14 Climate Change Effects on Intertidal and Subtidal Environments: Impacts, Projections, and Management........................................................................................................................................... 249 Kerrylee Rogers, Janine Adams, Nicole Cormier, Jeffrey Kelleway, and Neil Saintilan

Section 2  Biological Aspects Chapter 15 Estuarine and Coastal Marine Organism Responses to Climate Change: An Introduction......................... 277 Alexa Fredston and Benjamin S. Halpern Chapter 16 Microbial Ecology in a Changing Climate................................................................................................... 289 Jennifer L. Bowen Chapter 17 Climate Change, Phytoplankton, and HABs................................................................................................ 315 Hans W. Paerl Chapter 18 Responses of Marine Macroalgae to Climate Change Drivers..................................................................... 335 Yan Ji and Kunshan Gao Chapter 19 Effects of Climate Change on Salt Marshes................................................................................................. 355 Ivan Valiela, Javier Lloret, and Kelsey Chenoweth Chapter 20 Mangrove Forests and Climate Change: Impacts and Interactions.............................................................. 381 Daniel A. Friess, Luzhen Chen, Nicole Cormier, Ken W. Krauss, Catherine E. Lovelock, Jacqueline L. Raw, Kerrylee Rogers, Neil Saintilan, and Frida Sidik Chapter 21 Estuarine Seagrass and Climate Change...................................................................................................... 401 Kenneth A. Moore and Jessie C. Jarvis Chapter 22 Estuarine Benthos and Climate Change....................................................................................................... 431 Jeffrey S. Levinton Chapter 23 Estuarine Shellfish and Climate Change....................................................................................................... 451 Stephen J. Tomasetti and Christopher J. Gobler Chapter 24 Climate Change Effects on Fish Populations................................................................................................ 475 Alan K. Whitfield, Bronwyn M. Gillanders, and Kenneth W. Able Chapter 25 Estuarine and Coastal Birds, Climate Change, and Sea-Level Rise............................................................. 507 Joanna Burger

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Chapter 26 Climate Change and Invasive Species.......................................................................................................... 535 Edwin D. Grosholz Chapter 27 Animal Response to Hypoxia in Estuaries and Effects of Climate Change................................................. 545 Brad A. Seibel

Section 3  Management Aspects Chapter 28 Perspectives on Managing Estuaries while Addressing the Climate Crisis................................................. 565 Donald F. Boesch and Natalie Snider Chapter 29 Sea-Level Rise Risk and Adaptation in Estuaries........................................................................................ 581 Jochen Hinkel, Mark Schuerch, Jon French, and Robert J. Nicholls Chapter 30 Managing for Resilience of Estuarine and Coastal Marine Environments to Climate Change...................603 Heather M. Leslie, Melissa L. Britsch, Marina Cucuzza, Kara E. Pellowe, Sarah C. Risley, and Joshua S. Stoll Chapter 31 Climate Change Adaptation of Engineering Infrastructure in Estuarine Environments............................. 619 W. L. Peirson, R. J. Cox, and K. A. Bishop Chapter 32 Conserving and Managing Estuaries during Climate Change..................................................................... 633 J. K. O’Leary, E. E. Bockmon, M. Goodman, G. Grimsditch, M. A. Madej, A. Mohammed, and J. Tyburczy Index................................................................................................................................................................................... 655

Foreword Among all the features of nature that a climate scientist ponders, estuaries exercise a special grip on me that is as much personal as professional. I grew up less than three miles from Long Island Sound and did my first shell fishing there as a child (long before the US Clean Water Act was adopted; the oysters were inedible due to pollution). I have lived most of my adult life within walking distance of the Hudson River estuary – and I start most of my days with an early walk along its banks (where the water quality has improved markedly despite dense urban settlement on both banks). I studied the effects of nitrogen pollution originating from acid rain on the Chesapeake Bay estuary at a time when climate change was starting to emerge as an issue for governments. One of my early professional writings on the subject noted the special vulnerability of estuaries as the new threats from warming and sea-level rise interact with a pre-existing stew of environmental insults. Estuaries were then and still are locations of heavy urban and industrial development that remakes their contours. Estuaries are the places where major rivers discharge to the ocean the sewage, agricultural nutrients, atmospheric pollution, and toxins that drain from urban and industrial pipes, farmland, and the sky. The consequences have, in many cases, made the surviving fish too contaminated to be eaten and swimming unpleasant and unhealthy.

The pattern is repeated worldwide, despite improvements resulting from the determined efforts of scientists, citizen-activists, and some governments. From a scientific perspective, the situation is complex with many detrimental stresses abounding, so that the chief cause of deterioration is often difficult to determine. At a time when the health of some estuaries is improving, climate change and sealevel rise are accelerating. We need to learn a lot more, very fast. Climate Change and Estuaries takes a critical step in that direction by laying out the challenges and identifying ways to accommodate the effects of climate change while governments struggle to rein in the emissions of the greenhouse gases that are causing the warming. If someone, or just happenstance, desecrates a great work of art, as fire did Notre Dame, we do not throw it away or let it deteriorate until unrecognizable – we learn from experience, reduce the future threat, and call in the restoration crews to do their magic. Saving estuaries is more difficult than saving a masterpiece, but there is no other option. Michael Oppenheimer Albert G. Milbank Professor of Geosciences and International Affairs and the High Meadows Environmental Institute Princeton University

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Preface Estuaries are vital transitional environments where the land meets the sea and freshwater runoff mixes with seawater, resulting in salinity gradients, a continuum of biogeochemical cycles and biota from freshwater to marine, supporting an array of distinctive habitats. Collectively, they cover more than 400,000 km of global coastline with an estimated surface area of 1.1 × 106 km2 (Dürr et al. 2011; Lonsdale et al. 2022). Estuaries rank among the most productive ecosystems on earth, generating exceptional ecological and economic provisional services for contemporary society. However, their location in the coastal zone, often downstream of highly developed and densely populated watershed areas with escalating human pressures, renders them vulnerable to numerous anthropogenic and environmental stressors that can threaten ecosystem structure, function, and sustainability while concurrently reducing ecosystem services and the resilience of coastal habitats and their biotic communities. Indeed, much of the world’s coastal population, infrastructure, industry, fisheries, and tourism is located around estuaries, and the potential impacts of climate change on engineering infrastructure alone are substantial. Because of this, rising sea-levels threaten many coastal communities and metropolitan centers through a range of interconnected hazards and risks. Furthermore, anthropogenic forcing in the coastal zone, supporting increasing population growth and development concomitant with greater regional and global effects of climate change, is modulating physical–chemical and ecological conditions in estuarine ecosystems. This poses ever-increasing challenges for management and conservation programs. Climate change is a major driver of shifting estuarine dynamics. For example, the mechanisms driving biogeochemical cycling in estuaries are being altered by disturbances and stressors associated with climate change. Among the most profound consequences of climate change in estuaries are warming and altered hydrologic and biogeochemical conditions resulting from more intense wet/ dry cycles, including a rise in intense storms. These shifting conditions have altered watershed nutrient and contaminant delivery, resulting in accelerating eutrophication, proliferation of harmful algal blooms, and depletion of dissolved oxygen, exacerbating natural and anthropogenic hypoxia detrimental to biotic communities and their habitats. In addition, climate change can play a significant role in the establishment, spread, and impact of non-native invasive species commonly leading to the degradation of ecosystem components. Increasing sea-level rise, water temperature, salinity variation, altered freshwater inputs and circulation, nutrient enrichment, deoxygenation, acidification, invasive species, and resulting changes in physical, chemical, and biotic characteristics of estuarine environments are outcomes of escalating climate change.

The IPCC (2012) has defined climate change as regional or global changes in mean climate state or in patterns of climate variability over decades to millions of years often identified using statistical methods and sometimes referred to as changes in long-term weather conditions. Increasing atmospheric concentrations of greenhouse gases (primarily carbon dioxide, nitrous oxides, and methane) due to the burning of fossil fuels, deforestation, and changing land use have led to significant changes in the Earth’s climate system, more rapidly now during the Anthropocene than at any point in human history. The resulting outcomes have generally been detrimental to estuarine and coastal marine environments. There is an array of climate change stressors affecting estuarine organisms and habitats. Responses to these stressors are typically non-linear because they do not occur in isolation but have synergistic, cumulative, or antagonistic interactions that may increase, add to, or decrease the effect of individual stressors alone. Multiple direct anthropogenic drivers of change often interact synergistically with climatic drivers (e.g., higher intensity storms, increasing temperatures, heatwaves, rising sea-levels, extreme floods, and droughts) to amplify impacts in these environments. The effects can be transformative. In fact, some local anthropogenic activities, such as land reclamation, flow diversions, shoreline hardening, and channel dredging, are dominant drivers of hydrodynamic and other changes in estuarine systems that significantly affect biotic communities and habitats. Estuaries occur across diverse environmental settings, ranging, for example, from low-relief tropical coastlines to deep glaciated fjords. These characteristics influence how estuaries are variably impacted by climate change, and how they affect the role that estuaries play as a feedback in the global carbon cycle (Crosswell et al. 2022). Sea-level rise driven by climate change is likely to have the most profound impact on the structure and function of intertidal environments, which are a dominant feature along gently sloping coastal margins. By contrast, steeper coastlines like fjords may be impacted more by altered freshwater input and meltwater cycles resulting from climate change. Other climate change drivers, such as increasing air and water temperatures and extreme storm events, will interact with and amplify the effects of sea-level rise on these environments. Interactive drivers, in particular, have accelerated estuarine ecosystem degradation and the loss of ecosystem services by altering basic properties, physical and chemical characteristics, and ecological processes (Kennish 2021). Climatic and non-climatic drivers of change vary significantly in space and time, influencing physiological responses of microbes, flora and fauna, their behavior, and trophic interactions, frequently culminating in altered demographic characteristics, biotic community structure, and ecosystem function. Biotic responses to these drivers xi

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of change include marked changes in organism abundance, distribution, diversity, community composition, productivity, phenology, predator–prey interactions, and food web dynamics. Some responses are overtly damaging to estuarine environments, such as the development and proliferation of opportunistic harmful (toxic, hypoxia-generating, and food web altering) bloom-forming algal taxa (HABs) that impair estuarine water quality and degrade biotic communities and habitats (Paerl et al. 2018). The loss and fragmentation of estuarine habitats including salt marshes, mangroves, and seagrasses due to climatic forcing factors cause acute changes in biotic communities of coastal wetlands. These blue carbon ecosystems are important for the sequestration of carbon and the mitigation of climate change effects, as significant amounts of atmospheric CO2 are stored in wetland soils and sediments. Rising sea-levels and the consequent submergence of salt marshes reduce carbon sequestration processes in these critical coastal wetlands. Efforts to rehabilitate, revegetate, and restore these essential habitats are ongoing, although with variable success over protracted time intervals. Climate change presents major challenges in understanding the responses of planktonic, benthic, and nektonic organisms. Changes in estuarine circulation affect egg and larval transport, as well as organism recruitment, with significant consequences on estuarine population and community-level dynamics. Climate-mediated differential effects on reproduction, growth, and mortality will have an increasingly greater influence on the demographics of estuarine organisms during the 21st century. Of great concern is the sustainability of fisheries’ resources in the face of climate change and the conservation practices needed to support them. The cumulative effects of climate change impacts will increase the vulnerability of finfish and shellfish species, communities, and estuarine ecosystems. Aquaculture is emerging as a driving force for shellfish restoration and conservation in estuaries. Natural oyster reefs are vulnerable to biogeochemical fluxes, sediment inputs, and increased storm intensity, and aquaculture can augment shellfish production in estuarine waters. Because of concerns about fisheries and other biotic resources in estuarine and coastal marine environments, more investigations are also underway to assess climate-mediated effects on fish, mammals, and avian populations to determine the totality of climate effects on higher trophic-level organisms in these environments. Difficulties in assessing interactive drivers of change in estuarine environments are complicating the prediction and projected outcomes of ecologic and economic impacts important for the development of coastal management programs and their remedial measures (e.g., restoration and conservation initiatives) to address them and ultimately strengthen ecosystem resilience and sustainability. Integrated ecosystem-based management, which considers the connectivity of land, atmosphere, estuary, and ocean, as well as ecological components and ecosystem services, is an effective approach to protecting and maintaining these

Preface

coastal environments. Managing for their future requires an appreciation of the socio-economic-ecologic context, decision-making frameworks, management options, and strengthening of resilience in response to the climate crisis itself. Climate Change and Estuaries examines the drivers of change in estuarine environments, notably the interactive effects and impacts of multiple anthropogenic climatic and non-climatic drivers that affect the structure and function of estuarine ecosystems. It also assesses the management approaches implemented to mitigate the impacts and increase adaptation, resilience, and sustainability of these environments as well as built coastal communities. An example in estuaries is the sequestration of carbon in wetlands as a pathway to mitigate climate change impacts. Because of the danger that climate change poses to natural and built environments along the coast, conservation and restoration programs supporting the goals of long-term sustainability and resilience of coastal habitats and communities in the face of rising sea-levels and other climatic-driven effects must also be important elements of a holistic management program. Examples include the siting of Marine Protection Areas (MPAs), shoreline stabilization, wetland drainage restoration, and thin-layer sediment application to increase marsh elevation. The tasks to achieve management goals to halt or reverse the effects of climate change on estuaries over the succeeding decades of the 21st century will be difficult because of ongoing greenhouse gas emissions and the persistence of elevated CO2 concentrations in the atmosphere. This book consists of three sections: (1) physical–chemical aspects; (2) biological aspects; and (3) management aspects. It not only examines climatic and non-climatic drivers of change affecting estuarine environments but also their interactions and effects on populations of organisms, communities, habitats, and ecosystem structure and function. The management of climate change effects in estuaries considers both natural and built communities and includes mitigation, adaptation, and resilience programs. The vulnerability and sustainability of estuaries to climate change are also assessed in response to temperature increases, precipitation patterns, extreme storm and drought events, rising sea-levels, and other drivers of change. Chapter 1 is an overview of Climate Change and Estuaries. Chapter 2 covers the dynamics of climate change in the earth system. Chapter 3 provides a description of estuaries, their historical development, and classifications. Ensuing chapters examine specific aspects of climate change and interactive factors: (Chapter 4) sea-level rise; (Chapter 5) anthropogenic drivers of estuarine change; (Chapter 6) saltwater intrusion; (Chapter 7) biogeochemical changes; (Chapter 8) hypoxia; (Chapter 9) acidification; (Chapters 10) carbon dynamics; (Chapter 11) blue carbon; (Chapter 12) circulation and hydrological responses; (Chapter 13) sediment dynamics; (Chapter 14) intertidal and subtidal environments; (Chapter 15) organism responses; (Chapter 16) microbial ecology; (Chapter 17) nutrients, phytoplankton,

Preface

and harmful algal blooms (HABs); (Chapter 18) macroalgae; (Chapter 19) salt marshes; (Chapter 20) mangroves; (Chapter 21) seagrasses; (Chapter 22) benthic communities; (Chapter 23) shellfish; (Chapter 24) fish; (Chapter 25) estuarine and coastal birds; (Chapter 26) invasive species; (Chapter 27) faunal response to estuarine deoxygenation; (Chapter 28) estuarine management, mitigation, and future conditions; (Chapter 29) sea-level rise risk and adaptation in estuaries; (Chapter 30) resilience of estuarine and coastal marine environments to climate change; (Chapter 31) climate change adaptation of engineering infrastructure; and (Chapter 32) conservation and management strategies dealing with climate change impacts in estuaries.

REFERENCES Crosswell, J. R., F. Bravo, I. Pérez-Santos, G. Carlin, N. Cherukuru, C. Schwanger, R. Gregor, and A. Steven. 2022. Geophysical controls on metabolic cycling in three Patagonian fjords. Prog. Oceanogr. 207: 102866. Dürr, H. H., G. G. Laruelle, C. M. van Kempen, C. P. Slomp, M. Meybeck, and H. Middelkoop. 2011. Worldwide typology of nearshore coastal systems: Defining the estuarine filter of river inputs to the oceans. Estuar. Coasts 34(3): 441–458.

xiii Intergovernmental Panel on Climate Change (IPCC). 2012. Managing the risks of extreme events and disasters to advance climate change adaptation. pp. 339–329. In: C. B. Field, V. Barros, T. F. Stocker, D. Qin, D. J. Dokken, E. L. Ebi, M. D. Mastrandrea et al. (Eds.), A Special Report of Working Groups I and II of the Intergovernmental Panel on Climate Change. Cambridge: Cambridge University Press. Kennish, M. J. 2021. Drivers of change in estuarine and coastal marine environments: An overview. Open J. Ecol. 11(3): 224–239. Lonsdale, J.-A., C. Leach, D. Parsons, A. Burkwith, A. Manson, and M. Elliott. 2022. Managing estuaries under a changing climate: A case study of Humber Estuary, UK. Environ. Sci. Policy 134: 75–84. Paerl, H. W., T. G. Otten, and R. Kudela. 2018. Mitigating the expansion of harmful algal blooms across the freshwater-to-marine continuum. Environ. Sci. Technol. 52(10): 5519–5529.

Michael J. Kennish Hans W. Paerl Joseph R. Crosswell

About the Editors Dr. Michael J. Kennish is a professor emeritus in the Department of Marine and Coastal Sciences, School of Environmental and Biological Sciences, Rutgers University, New Brunswick, New Jersey. His career in coastal, estuarine, and marine sciences spans 50 years and has included extensive multidisciplinary research on coastal, estuarine, and marine ecosystems. He has taught coastal and marine science classes at Rutgers University for many years, while also supervising undergraduate and graduate students. In addition, he has been active for decades in the outreach of science to coastal communities and K–12 schools. As a member of the Climate Institute at Rutgers University, Dr. Kennish has been involved in the study of long-term climate change impacts on the New Jersey coast and elsewhere. He was an expert reviewer of the Sixth Assessment Report of the Intergovernmental Panel on Climate Change (IPCC) published in 2021 (WGI) and 2022 (WGII and WGIII). Dr. Kennish is the author or editor of 16 scholarly books on various topics in coastal, estuarine, and marine sciences, the author of more than 200 research articles in science journals and other publications, and the editor of 9 peer-reviewed compendium science journal special issues. Dr. Kennish has maintained a wide range of research interests in marine ecology and marine geology. He has been most actively involved in leading research teams investigating estuarine and coastal marine environments in New Jersey. Much of this research has involved the development and application of innovative methods to determine the condition and ecosystem health of coastal ecosystems in the state. Dr. Kennish is widely known for his work on the human impacts of coastal, estuarine, and marine environments and has served on environmental panels and workgroups assessing these problems in New Jersey, the mid-Atlantic region, and nationwide, while concomitantly collaborating extensively with state and federal government agencies to remediate degraded water quality and habitats. Most notably, he has been heavily engaged in investigations of impairment and remediation of impacted estuarine and coastal marine environments. These include studies of the natural and anthropogenic stressors that effect change in coastal ecosystems as well as the dynamics of environmental forcing factors that generate imbalances in biotic community structure and ecosystem function. His research, which has been funded by the USEPA, NOAA, USDA, state environmental agencies, and other federal and state sources, is multidisciplinary in scope. It has addressed an array of nationally significant problems, such as habitat loss and alteration of aquatic systems, nutrient enrichment

and eutrophication, hypoxia and anoxia, organic pollution, chemical contaminants, climate change, sea-level rise, overfishing, invasive species, watercraft effects, dredging and dredged-material disposal, freshwater diversions, calefaction of estuarine waters, entrainment and impingement of electric generating stations, and the effects of watershed development on coastal systems. In addition, he has examined the effects of the construction and operation of industrial facilities, maintenance of shorelines and waterways, and human use of coastal space and aquatic systems. He has also studied the biology and geology of mid-ocean ridge and deep-sea hydrothermal vent systems as a member of the Center for Deep-Sea Ecology and Biotechnology at Rutgers University. Dr. Kennish is the recipient of many awards, including the 2008 Guardian of the Barnegat Bay Award (Barnegat Bay National Estuary Program/USEPA), 2009 NOAA/ NERRA National Award for outstanding contributions to the National Estuarine Research Reserve System of NOAA, 2010 Graham Macmillan Award of the American Littoral Society for significant contributions to marine science and conservation, 2010 Sierra Club Award for outstanding environmental accomplishments, 2011 Pearl S. Schwartz Environmental Award of the League of Women Voters for work on New Jersey’s coastal environments, 2013 Frank Oliver Award of the New Jersey Environmental Lobby for contributions to the protection of New Jersey’s environments, and the 2017 Albert Nelson Marquis Lifetime Achievement Award for dedication to the environmental and oceanographic sciences. Dr. Hans W. Paerl is the Kenan Professor of Marine and Environmental Sciences at the University of North Carolina’s Institute of Marine Sciences. He holds a joint appointment with the Departments of Earth, Marine and Environmental Sciences and Environmental Sciences and Engineering. His collaborative research addresses microbially mediated nutrient cycling and primary production dynamics, environmental controls, and management of harmful algal (specifically cyanobacterial) blooms. Dr. Paerl’s research spans freshwater lakes, reservoirs (including ones used as drinking water supplies), estuarine, and coastal marine waters in the USA and globally (see: https://paerllab​.web​.unc​.edu​ /research/). He has published over 350 peer-reviewed articles and book chapters on these subjects. His work has been supported by the NSF, EPA, NIH, NOAA/NC Sea Grant, USDA, the NC Water Resources Research Institute, the UNC Collaboratory, the California Bay Delta Science Program, xv

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various state environmental agencies and private foundations, the Chinese Academy of Sciences and Ministry of Science and Technology, and the Netherlands Academy of Sciences. He also manages several estuarine water quality monitoring and assessment programs, including the Neuse River Estuary (NC) Modeling and Monitoring Program, ModMon (https://paerllab​.web​.unc​.edu​/projects​/modmon/) and the Ferry-based Water Quality Monitoring Program for the Pamlico Sound System, FerryMon (https://paerllab​.web​.unc​.edu​/projects​/ferrymon/). Dr. Paerl has supervised over 70 graduate students, 12 postdocs, and advises undergraduate students at UNC-CH as well as other institutions. He received the 2003 G. Evelyn Hutchinson Award from the Association of the Sciences of Limnology and Oceanography (ASLO), and the 2011 Odum Award from the Coastal and Estuarine Research Federation (CERF) for addressing the causes, consequences, and controls of eutrophication and harmful algal blooms in aquatic ecosystems. In 2015, he was named a Fellow of the American Geophysical Union (AGU), and in 2022 he received the Lifetime Achievement Award from the US Harmful Algal Bloom Committee. He is a fellow of the Royal Dutch Academy of Sciences and holds honorary joint faculty positions at the Hohai University and the Nanjing Institute of Geography and Limnology, Chinese Academy of Sciences, both located in Nanjing, China. Dr. Joseph (Joey) Crosswell is a senior research scientist at the Commonwealth Scientific and Industrial Research Organization (CSIRO). Dr. Crosswell’s core research examines the connectivity of coastal systems, particularly carbon and nutrient cycling between sediment, ocean, and atmosphere (see https://people​.csiro​ .au​/c​/j​/joey​-crosswell). He is particularly interested in exploring diverse and remote coastal systems, ranging from

About the Editors

mangroves to mesoscale eddies and from arid tropical estuaries in northern Australia to fjords in southern Chile. Dr. Crosswell has served as Chief Scientist on 22 research voyages at CSIRO, leading cruises to more than 70 estuaries and coral reefs in the Great Barrier Reef ecosystem as well as voyages to other remote regions of Australia, Patagonia, Fiji, Papua New Guinea, and Indonesia. He has supervised seven graduate students, two postdocs, and several undergraduate students at institutions spanning six countries. A key element of Dr. Crosswell’s research involves developing new tools and methods for coastal observations. These include instrument platforms for measuring carbonate chemistry and physical processes in estuaries, which have been deployed over the past 15 years in waters along the mid-Atlantic coast of the USA, the east Australian coast, and in Chilean fjords. More recently, Dr. Crosswell has collaborated with CSIRO researcher-engineers on the application of computer vision to detect crown of thorns starfish outbreaks in the Great Barrier Reef as well as machine learning models for mapping seagrass across the Indo-Pacific. Dr. Crosswell plays an active role in advancing national and international research strategy and collaboration. Since 2020, he has been co-leader of the research domain Coastal Mapping and Monitoring within CSIRO’s Environment Business Unit. He has served as a nominated author for expert scientific assessments in the USA and Australia, such as the 2nd State of the Carbon Cycle Report. He is a founding member of an international working group focused on feedback between carbon cycling and extreme events, e.g., tropical cyclones, floods, and wildfires. He leads research and field studies for multi-lateral initiatives exploring blue carbon as a resource for climate action and sustainable livelihoods across the Indo-Pacific. Dr. Crosswell has received an Innovators Fellowship from the Deshpande Foundation, a Collaborative Research Award from the Burroughs Wellcome Fund, a Research Impact Award from the University of North Carolina, and the 2022 CSIRO Collaboration medal.

List of Contributors Kenneth W. Able Department of Marine and Coastal Sciences Rutgers University Marine Field Station Tuckerton, New Jersey, United States

Jacob Carstensen Department of Ecoscience Aarhus University Roskilde, Denmark

Janine Adams Department of Botany Nelson Mandela University Port Elizabeth, South Africa

Robert J. Chant Department of Marine and Coastal Sciences School of Environmental and Biological Sciences Rutgers University New Brunswick, New Jersey, United States

Thomas S. Bianchi Department of Geological Sciences University of Florida Gainesville, Florida, United States K. A. Bishop Aquatic Ecologist Sugar Creek Road Bungwahl, Australia Donald F. Boesch University of Maryland Center for Environmental Science Annapolis, Maryland, United States E. E. Bockmon Chemistry and Biochemistry Department California Polytechnic State University San Luis Obispo, California, United States Jennifer L. Bowen Department of Marine and Environmental Sciences Marine Science Center Northeastern University Nahant, Massachusetts, United States Melissa L. Britsch Maine Department of Marine Resources Augusta, Maine, United States Joanna Burger Department of Biological Sciences Rutgers University New Brunswick, New Jersey, United States Wei-Jun Cai School of Marine and Science Policy University of Delaware Newark, Delaware, United States

Luzhen Chen Key Laboratory of the Ministry of Education for Coastal and Wetland Ecosystems Xiamen University Xiamen, China Kelsey Chenoweth The Ecosystems Center Marine Biological Laboratory Woods Hole, Massachusetts, United States John A. Church Climate Change Research Centre University of New South Wales Sydney, Australia and Australian Centre for Excellence in Antarctic Science (ACEAS) University of New South Wales NSW, Australia Nicole Cormier Department of Earth and Environmental Sciences Macquarie University Sydney, Australia R. J. Cox Australian Climate Change Adaptation Research Network for Settlements and Infrastructure University of New South Wales Sydney Sydney, Australia Joseph R. Crosswell Oceans and Atmosphere Commonwealth Scientific and Industrial Research Organisation (CSIRO) Brisbane, Australia

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List of Contributors

Marina Cucuzza South Central Climate Adaptation Science Center Norman, Oklahoma, United States

G. Grimsditch United Nations Environment Programme Nairobi, Kenya

David R. Easterling National Climate Assessment Technical Support Unit National Centers for Environmental Information National Oceanic and Atmospheric Administration Asheville, North Carolina, United States

Edwin D. Grosholz Department of Environmental Science and Policy University of California, Davis Davis, California, United States

Alexa L. Fredston Department of Ocean Sciences University of California, Santa Cruz Santa Cruz, California, United States Jon French Coastal and Estuarine Research Unit Department of Geography University College London London, UK Daniel A. Friess Department of Earth and Environmental Sciences Tulane University New Orleans, Louisiana, United States Neil K. Ganju U.S. Geological Survey Woods Hole Coastal and Marine Science Center Woods Hole, Massachusetts, United States Kunshan Gao State Key Laboratory of Marine Environmental Science College of Ocean and Earth Sciences Xiamen University Xiamen, China and Co-Innovation Center of Jiangsu Marine Bio-industry Technology Jiangsu Ocean University Lianyungang, China Bronwyn M. Gillanders Southern Seas Ecology Laboratories School of Biological Sciences University of Adelaide Adelaide, Australia Christopher J. Gobler School of Marine and Atmospheric Sciences Stony Brook University Stony Brook, New York, United States M. Goodman Hopkins Marine Station Stanford University Pacific Grove, California, United States

Benjamin S. Halpern Bren School of Environmental Science and Management University of California, Santa Barbara Santa Barbara, California, United States Jochen Hinkel Global Climate Forum Berlin, Germany and Resource Economics Group Albrecht Daniel Thaer-Institute and Berlin Workshop in Institutional Analysis of Social- Ecological Systems (WINS) Humboldt University Berlin, Germany Charles S. Hopkinson Department of Marine Sciences University of Georgia Athens, Georgia, United States Jessie C. Jarvis Department of Biology and Marine Biology University of North Carolina, Wilmington Wilmington, North Carolina, United States Yan Ji State Key Laboratory of Marine Environmental Science College of Ocean and Earth Sciences Xiamen University Xiamen, China and School of Biological and Chemical Engineering Qingdao Technical College Qingdao, China Jeffrey Kelleway School of Earth, Atmospheric and Life Sciences University of Wollongong Wollongong, Australia Michael J. Kennish Department of Marine and Coastal Sciences School of Environmental and Biological Sciences Rutgers University New Brunswick, New Jersey, United States

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List of Contributors

Ken W. Krauss U.S. Geological Survey Wetland and Aquatic Research Center Lafayette, Louisiana, United States Kenneth E. Kunkel Cooperative Institute for Satellite Earth System Studies North Carolina Institute for Climate Studies North Carolina State University Asheville, North Carolina, United States Arnaud Laurent Department of Oceanography Dalhousie University Halifax, Nova Scotia, Canada Heather M. Leslie University of Maine Darling Marine Center Walpole, Maine, United States and School of Marine Sciences University of Maine Orono, Maine, United States

Allison Myers-Pigg Marine and Coastal Research Laboratory Pacific Northwest National Laboratory Sequim, Washington, United States and Department of Environmental Sciences University of Toledo Toledo, Ohio, United States Robert J. Nicholls Tyndall Centre for Climate Change Research University of East Anglia Norwich, UK J. K. O’Leary Wildlife Conservation Society Mombasa, Kenya Christopher L. Osburn Department of Marine, Earth, and Atmospheric Sciences North Carolina State University Raleigh, North Carolina, United States

Jeffrey S. Levinton Department of Ecology and Evolution Stony Brook University Stony Brook, New York, United States

L. Fernando Pareja-Roman Department of Marine and Coastal Sciences School of Environmental and Biological Sciences Rutgers University New Brunswick, New Jersey, United States

Ming Li Horn Point Laboratory University of Maryland Center for Environmental Science Cambridge, Maryland, United States

Hans W. Paerl Institute of Marine Sciences University of North Carolina at Chapel Hill Morehead City, North Carolina, United States

Javier Lloret The Ecosystems Center Marine Biological Laboratory Woods Hole, Massachusetts, United States

Kara E. Pellowe University of Maine Darling Marine Center Walpole, Maine, United States

Catherine E. Lovelock School of Biological Sciences The University of Queensland St Lucia, Queensland, Australia

School of Marine Sciences University of Maine Orono, Maine, United States

M. A. Madej Coastal Ecosystems Institute of Northern California Humboldt, California, United States

Stockholm Resilience Centre Stockholm, Sweden

and

and

A. Mohammed United Nations Environment Programme Nairobi, Kenya

William Peirson New College University of New South Wales Sydney Sydney, Australia

Kenneth A. Moore Department of Biological Sciences The Virginia Institute of Marine Science College of William and Mary Gloucester Point, Virginia, United States

Jacqueline L. Raw Department of Botany and Institute for Coastal and Marine Research Nelson Mandela University Port Elizabeth, South Africa

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Sarah C. Risley University of Maine Darling Marine Center Walpole, Maine, United States and

List of Contributors

Stephen J. Tomasetti Department of Natural Sciences University of Maryland Eastern Shore Princess Anne, Maryland, United States

School of Marine Sciences University of Maine Orono, Maine, United States

J. Tyburczy California Sea Grant Humboldt, California, United States

Kerrylee Rogers School of Earth, Atmospheric and Life Sciences University of Wollongong Wollongong, Australia

Ivan Valiela The Ecosystems Center Marine Biological Laboratory Woods Hole, Massachusetts, United States

Neil Saintilan Department of Earth and Environmental Sciences Macquarie University Sydney, Australia

Arnoldo Valle-Levinson Department of Civil and Coastal Engineering University of Florida Gainesville, Florida, United States

Mark Schuerch Department of Geography University of Lincoln Lincoln, UK

Nicholas D. Ward Marine and Coastal Research Laboratory Pacific Northwest National Laboratory Sequim, Washington, United States

Brad A. Seibel College of Marine Science University of South Florida St. Petersburg, Florida, United States

Nathaniel Weston Department of Geography and the Environment Villanova University Villanova, Pennsylvania, United States

Frida Sidik Institute for Marine Research and Observation Ministry of Marine Affairs and Fisheries Perancak, Bali, Indonesia

Alan K. Whitfield South African Institute for Aquatic Biodiversity Somerset Street Grahamstown/Makhanda, South Africa

Natalie Snider Environmental Defense Fund Washington, D.C., United States

Lisamarie Windham-Myers U.S. Geological Survey Water Resources Mission Area Menlo Park, California, United States

Joshua S. Stoll University of Maine Darling Marine Center Walpole, Maine, United States and School of Marine Sciences University of Maine Orono, Maine, United States Jeremy M. Testa Chesapeake Biological Laboratory University of Maryland Center for Environmental Science Solomons, Maryland, United States

Xuebin Zhang Climate Science Centre Commonwealth Scientific and Industrial Research Organisation (CSIRO) Hobart, Australia

Section 1 Physical–Chemical Aspects

1

Introduction to Climate Change and Estuaries Michael J. Kennish, Hans W. Paerl, and Joseph R. Crosswell

CONTENTS Abstract................................................................................................................................................................................... 3 1.1 Introduction................................................................................................................................................................... 3 1.2 Plan of the Volume........................................................................................................................................................ 4 1.3 Climate Change Drivers................................................................................................................................................5 1.3.1 Temperature Increases.......................................................................................................................................5 1.3.2 Sea-Level Rise................................................................................................................................................... 7 1.3.3 Other Climatic Drivers...................................................................................................................................... 8 1.4 Anthropogenic Non-climatic Drivers.......................................................................................................................... 10 1.5 Interactive Drivers of Change...................................................................................................................................... 10 1.6 Ecological Impacts of Climate Change....................................................................................................................... 14 1.7 Management Strategies................................................................................................................................................ 16 1.8 Conclusions.................................................................................................................................................................. 16 Acknowledgments................................................................................................................................................................. 17 References............................................................................................................................................................................. 18

ABSTRACT Estuaries are highly variable, complex coastal environments that rank among the most productive aquatic ecosystems on earth. They have great value to humans because of their diverse and extensive services. However, estuaries are particularly susceptible to human impacts because of burgeoning population growth and development in the coastal zone as well as multiple climatic stressors. Climate change affects estuaries in a myriad of ways, most notably by increasing water temperature, sea-level, tropical cyclone intensity, storm surge, extreme precipitation, and freshwater flux interspersed with record droughts, resultant major salinity shifts, as well as altered biogeochemical cycling and circulation processes, accelerating shoreline erosion, and degrading wetlands habitat; all of which is leading to major alterations of the structure and function of biotic communities (e.g., species abundance, distribution, diversity, reproduction, phenology, production, and trophic interactions) worldwide. Climatic drivers often interact with anthropogenic non-climatic drivers, including nutrient over-enrichment, sediment loads, and xenobiotic pollutant inputs to exacerbate adverse effects in estuaries. The overall impact of climate change on estuarine and coastal marine environments is increasing, and thus effective management strategies are needed to mitigate these adverse effects while also effectively increasing resilience and environmental sustainability. Restoration efforts must also be implemented for the recovery of estuaries that are already severely impacted and no longer sustainable. DOI: 10.1201/9781003126096-2

Key Words:  estuaries, climatic drivers, non-climatic drivers, anthropogenic impacts, ecological changes, mitigation, adaptation, resilience, sustainability

1.1 INTRODUCTION Estuaries are complex and dynamic coastal ecosystems. They are transitional environments where the land meets the sea and freshwater runoff mixes with seawater, resulting in salinity gradients, a continuum of biogeochemical cycles and biota from freshwater to marine, as well as an array of distinctive habitats. They rank among the most productive ecosystems on earth, rivaling those of coral reefs and tropical rain forests. About 40% of the world’s more than 8 billion human population lives within 100 km of the coast, with many depending on the provision of estuarine environments (Kennish 2019). The global value of estuarine ecosystem services and that of adjoining coastal wetlands is estimated to be in the trillions of dollars and includes broad and essential ecological, economic, and societal benefits (Barbier et al. 2011; Costanza et al. 2014). Among the most highly valued human services associated with estuaries are recreational and commercial fisheries, tourism, aquaculture, electric power generation, oil and gas operations, transportation, and shipping, as well as a source of natural substances used in the production of specialty chemicals and foods, medicines, and pharmaceuticals. Many large and vibrant cities of the world are located near estuaries, and hence the real estate value bordering these coastal 3

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water bodies is typically high (Thrush et al. 2014). They are important venues for coastal tourism, being heavily utilized by swimmers, fishers, boaters, hunters, and birdwatchers. Clearly, the world economy relies heavily on the human use of estuaries and the rich resources and services they provide. Numerous environmental factors affect the structure (biotic community composition, species abundance, and biodiversity) and function (biotic production of organic matter, biogeochemical cycling, including the transformation of nutrients, filtration of pollutants, capture and storage of carbon, and trapping of sediments) of estuarine ecosystems (Kennish 2016, 2017). An array of habitats supports numerous populations of estuarine and marine organisms many of recreational and commercial importance. As noted by Kennedy et al. (2002), about 75% of commercially harvested marine fish and shellfish species representing up to 90% of the recreational harvest of marine species in the USA depend on estuaries for reproduction and spawning, nursery and refuge habitats, feeding grounds, and migration routes between spawning and feeding habitats. The buffering capacity of coastal wetlands and estuaries helps protect and sustain upland habitats and infrastructure against the insidious and acute impacts of climate change, including sea-level rise, storm surge, shoreline erosion, inundation, flooding, and salinity intrusion. Because coastal development is escalating with millions of new inhabitants living in watersheds surrounding estuaries, management strategies are being developed to ensure the long-term sustainability and resilience of coastal habitats and communities in the face of rising sea-levels and other climate-driven effects (Kennish 2021, 2022). Since 1950, the world population has more than tripled (~2.5 billion in 1950 to >8 billion in 2022; United Nations Department of Economics and Social Affairs Population Dynamics 2022) with a major fraction of the total growth occurring in coastal regions. In the USA alone, more than 130 million people (~42% of the nation’s population) inhabited coastal shoreline counties in 2013 (Kildow et al. 2016; Payne et al. 2018). These population figures are similar to those of other countries. Human population growth and urbanization are increasing substantially in coastal zones, with an estimated 75% of the world’s population expected to live there by 2025 (Bianchi 2007; Bianchi and Allison 2009). Not surprisingly, anthropogenic stressors are major drivers of change in estuarine environments. Coastal erosion coupled with rising sea-levels impacts beaches, dunes, estuarine basins, and shoreline habitats worldwide. In the USA, for example, Hapke et al. (2011) reported long-term erosion rates of 0.5 ± 0.09 m/yr along the Mid-Atlantic and New England coasts. More than 80% of the beaches on the east coast of the USA are now eroding. Beach renourishment projects are on the rise in many coastal regions to help maintain these habitats. The objective of this book is to examine and assess the interactive effects and impacts of multiple anthropogenic climatic and non-climatic drivers that affect and alter the

Climate Change and Estuaries

structure and function of estuarine ecosystems, as well as the management approaches to mitigate the impacts and increase sustainability and resilience of these ecosystems and coastal communities. Climatic drivers include “any climate-induced factor that directly or indirectly causes a change” (Wong et al. 2014). Climate change is defined as regional or global changes in mean climate state or in patterns of climate variability over decades to millions of years often identified using statistical methods and sometimes referred to as changes in long-term weather conditions (IPCC 2012). Climate change effects in estuaries must be assessed in conjunction with direct anthropogenic drivers in coastal watersheds that can exacerbate overall impacts, including removal of natural vegetation, erosion of soils, mobilization of sediments, associated nutrients and other pollutants, construction of impervious surfaces and hardened shorelines, application of fertilizers, diversion of freshwater, and infrastructure expansion, as well as estuarine basin effects, such as dredging, marina and harbor construction, shipping, and aquaculture (Table 1.1). Improved land-use policies and water management plans are necessary to reduce interactive effects of anthropogenic climatic and non-climatic drivers of change that are adversely affecting estuaries and other coastal environments (Breitburg et al. 1998, 2018; Scavia et al. 2002; Crain et al. 2008; Monbaliu et al. 2014; Robins et al. 2016; Bindoff et al. 2019). Management strategies are focusing on the mitigation of anthropogenic climatic and non-climatic stressor impacts, adaptation to changing conditions, an increase in resilience of coastal communities, and sustainability of habitats.

1.2 PLAN OF THE VOLUME Climate Change and Estuaries consist of three major sections: (1) physical–chemical aspects; (2) biological aspects; and (3) management aspects. The book not only examines climatic and non-climatic drivers of change affecting estuarine environments but also their interactions and effects on populations of organisms, communities, habitats, and ecosystem structure and function. The management of climate change effects in estuaries considers both natural and built communities and includes mitigation, adaptation, and resilience programs. The vulnerability and sustainability of estuaries to climate change are also assessed in response to temperature increases, altered precipitation amounts and patterns, more extreme storm and drought events, rising sea-levels, and other drivers of change. Chapters in the book discuss how climate change affects estuaries by modulating water temperature, salinity, sealevel, storm intensity and storm surge, precipitation and freshwater flux, as well as biogeochemical and circulation processes, shoreline erosion, sediment delivery and deposition, and other phenomena. For example, warming waters and density changes, shifting currents and water masses, reduced dissolved oxygen and acidification, and other factors are affecting the biogeochemical cycling of carbon and

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Introduction to Climate Change and Estuaries

TABLE 1.1  Major anthropogenic drivers of change in estuaries Drivers Class 1 (Degrade Water Quality)   Nutrient Enrichment/Eutrophication    Organic Carbon and Thermal Loading   Biogeochemical Alteration   Chemical Contaminants   Sediment/Particulate Inputs   Sewage Inputs   Pathogens Class 2 (Impact Habitat)    Watershed Development    Watershed Impervious Cover    Dredging and Dredged-Material Disposal   Shoreline Hardening   Lagoon Construction    Land Reclamation and Impoundments   Coastal Subsidence Class 3 (Alter Biotic Communities)    Human-altered Hydrological Regimes   Overfishing   Intensive Aquaculture   Invasive/Introduced Species   Floatables/Plastics/Debris Class 4 (Climate Linked)    Climate Change Drivers    CO2, CH4, NO2, Chlorofluorocarbons, (Greenhouse Gases)     Warming Temperatures     Precipitation and Land Runoff     Extreme Events      Hurricanes and Other Major Storms      Storm Surges      Tornadoes      Droughts Source: Modified from Kennish (2021).

nutrients (N, P, Si, Fe, and micronutrients), and from the organismal perspective, abundance, distribution, diversity, reproduction, phenology, species interactions, food web dynamics, and community structure of estuarine and marine organisms. Climate-driven biotic invasions and extinctions are increasing. Fishery yields are changing in different regions, some driven by overfishing and others by water quality and habitat declines. Rising sea-level, mainly attributable to the thermal expansion of the oceans and melting of ice sheets, is responsible for the increasing loss of essential coastal habitat (e.g., salt marshes, seagrasses, and mangroves) and an array of diminishing ecosystem services. Rising sea-level reduces the resilience and sustainability of many coastal wetlands and other essential estuarine and coastal marine habitats that provide a protective buffer for

natural and developed communities from extreme weather events, inundation, and flooding. Increasing frequencies of extreme rainfall and flooding events, and associated accelerated loading of nutrients and organic matter in runoff, are accelerating eutrophication, increasing HABs, and expanding hypoxia in receiving waters (Figure 1.1). Chapter 1 is an overview of Climate Change and Estuaries. Chapter 2 covers the dynamics of climate change in the earth system. Chapter 3 provides a description of the origin, historical development, and classifications of estuaries. Ensuing chapters examine specific aspects of climate change and interactive factors: (Chapter 4) sealevel rise; (Chapter 5) anthropogenic drivers of estuarine change; (Chapter 6) saltwater intrusion; (Chapter 7) biogeochemical changes; (Chapter 8) hypoxia; (Chapter 9) acidification; (Chapter 10) carbon dynamics; (Chapter 11) blue carbon; (Chapter 12) hydrological responses and circulation changes; (Chapter 13) sediment dynamics; (Chapter 14), intertidal and subtidal environments; (Chapter 15) organism responses; (Chapter 16) microbial ecology; (Chapter 17) nutrients, phytoplankton, and harmful algal blooms (HABs); (Chapter 18) macroalgae; (Chapter 19) salt marshes; (Chapter 20) mangroves; (Chapter 21) seagrasses; (Chapter 22) benthic communities; (Chapter 23) shellfish; (Chapter 24) fish; (Chapter 25) avifauna; (Chapter 26) invasive species; (Chapter 27) deoxygenation and estuarine fauna; (Chapter 28) estuarine management; (Chapter 29) risks and adaptation; (Chapter 30) resilience; (Chapter 31) effects on engineering infrastructure; and (Chapter 32) conservation and management strategies.

1.3 CLIMATE CHANGE DRIVERS 1.3.1 Temperature Increases The global mean surface temperature (GMST) of the earth is increasing. It increased by 1.09°C between 1850–1900 and 2011–2020 (Gulev et al. 2021). The 20th century was the warmest in more than 1000 years; however, the highest temperature increases have occurred since 1970. The last two decades (2001-2020) were warmer than those during the past century, with the decade of 2011–2020 being the warmest on record (IPCC 2021). As noted by Robert Kopp, a climate scientist in the Department of Earth and Planetary Sciences at Rutgers University, more than 90% of the warmest years on record have occurred since 2000. Among the warmest years recorded by NASA and NOAA (i.e., 2015–2020), 2016 and 2020 were the warmest, reaching 1.02°C above the baseline 1951–1980 mean. Between 1880 and 2012, the globally averaged combined land and ocean surface temperature increased by 0.85°C (0.65 to 1.06°C) (Wong et al. 2014). The highest combined temperatures were observed after 2000, with 2010 tying with 2005 as the warmest combined global land and ocean annual surface temperature increase of 0.62 ± 0.07°C (IPCC 2014). Global climate models predict that anthropogenic and natural forcing factors will lead to significant

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Climate Change and Estuaries

FIGURE 1.1  Conceptual diagram, adapted from Paerl (2014), of interactive physical, chemical, and biological controls on eutrophication, harmful algal bloom formation, and hypoxia along the freshwater-to-marine continuum.

planetary temperature increases during the 21st century, likely exceeding 1–2°C unless significant measures are taken to mitigate greenhouse gas emissions (IPCC 2021). Water temperatures in the estuaries and oceans of the world are increasing as well (Kennish 2019; Bindoff et al. 2019; Fox-Kemper et al. 2021). However, the temperature increases have been significantly less than over land; for example, for the period from 1850–1900 to 2011–2020, the GMST increased by 0.88°C on the ocean surface compared to 1.59°C over the land (Gulev et al. 2021). Most of the warming in the oceans (0.60°C) has occurred since 1980, with the rate of ocean warming more than doubling since 1993 (Pörtner et al. 2019; Gulev et al. 2021). Marine heatwaves (i.e., sustained periods of anomalously high near-surface temperatures that can lead to severe and persistent impacts on marine ecosystems) are becoming more frequent, doubling in number since the 1980s (Fox-Kemper et al. 2021). Ocean surface temperatures are projected to increase further during the 21st century (on average by 0.86°C for the period from 1995–2014 to 2081–2100 in the SSP1-2.6 emission scenario and by 2.89°C in the SSP5-8.5 emission scenario) (Fox-Kemper et al. 2021). The oceans are the primary sink for the increase in energy storage in the earth’s climate system, accounting for more than 90% of the heat energy accumulated from 1971 to 2000. Much of the added global heat is concentrated in the upper ocean (Levitus et al. 2012). Between 1955 and 1995, water temperature for the combined Atlantic, Pacific, and Indian Oceans increased by 0.06°C in the upper 3000 m and 0.31°C in the upper 300 m. Similarly, water

temperature in the Southern Ocean between the 1950s and 1980s increased by 0.17°C at mid-depths (700–1100 m) (Moran 2011). The mortality rates of many marine organisms will increase with rising ocean temperatures, as will range shifts in organism geographic distributions and altered trophic interactions (Robins et al. 2016; HoeghGulberg et al. 2018; Bindoff et al. 2019; Fox-Kemper et al. 2021). These changes will significantly affect ecosystem services (IPCC 2021). More than 70% of the world’s coastlines have experienced significant increases in sea surface temperature over the past three decades at a rate of 0.18 ± 0.16°C per decade (Lima and Wethey 2012). The mean rate of increase in sea surface temperature along coastlines has been greater than that of the open ocean (Wong et al. 2014). Rising temperatures will significantly change the physical state of highlatitude systems (e.g., reduction in glaciated fjords and loss of permafrost in watersheds) (IPCC 2021). Rising global temperatures are greatly affecting estuarine organisms by altering physiological and reproductive processes, as well as growth and survival rates (Robins et al. 2016; Frid and Caswell 2017; Kennish 2019). Warmer estuarine waters promote algal blooms, and greater precipitation and runoff driven by climate change increase bacterial and viral pathogens (e.g., Vibrio spp., adenovirus, and norovirus), facilitating disease transmission via waterborne and foodborne sources (Paerl et al. 2014, 2017; Kennish 2019). Such transmission increases in estuaries receiving heavy microbial loads in agricultural runoff and sewage inputs from combined sewer overflows (CSOs), particularly

Introduction to Climate Change and Estuaries

after storm events that deliver large amounts of precipitation (Kennish 1997, 2002, 2016; Robins et al. 2016). These effects can significantly degrade water quality (Kennish 2019). Increasing atmospheric and ocean temperatures are coupled with escalating greenhouse gas emissions, most notably carbon dioxide (CO2), largely driven by human-induced fossil-fuel combustion. Deforestation also has contributed to the increase in atmospheric CO2 concentrations observed over the past century (IPCC 2021). CO2 concentrations in the atmosphere now exceed 410 ppm, up from 285.5 ppm in 1850 (Gulev et al. 2021), and likely reaching their highest level in the last 15 million years in 2020. In addition, watershed land-use changes, urbanization, as well as agricultural and industrial expansion are responsible for increased emissions of other potent greenhouse gases, including methane (CH4) and nitrous oxides (N2O). For example, CH4 and N2O levels reached levels of 1866.3 ppb and 332.1 ppb, respectively, in 2019 far exceeding pre-industrial levels (Gulev et al. 2021). A long-term goal of the 2015 Paris Climate Agreement is to limit the increase in global mean temperature to well below 2°C above pre-industrial levels and to pursue efforts to limit the temperature increase to 1.5°C above pre-industrial levels. Limiting temperature increases below these levels will substantially reduce the environmental risk and effects of climate change on estuarine and marine organisms and habitats. However, the current rate of CO2 and other greenhouse gas emissions will lead to higher levels of temperature increases and more intense and persistent droughts. Unless these emissions are substantially curtailed or eliminated, the global mean surface temperature increase will exceed 2°C or even 3°C by the end of the 21st century. Currently, less than 20 of the 195 countries in the Paris Climate Agreement are on track to meet their CO2 emission goals.

1.3.2 Sea-Level Rise The mean ocean heat content has steadily increased over the past 50 years. Thermal expansion of ocean waters and the melting of polar ice sheets and continental glaciers are the primary drivers of rising sea-levels worldwide (Bindoff et al. 2019; Kennish 2019; Pörtner et al. 2019; Fox-Kemper et al 2021). Global warming projections indicate significant increases in oceanic mass and higher sea-levels during the 21st century as well. The Greenland and Antarctic (western Antarctic Peninsula) ice sheets, which have been losing mass over the past two decades because of increasing temperatures and meltwater generation, contribute substantially to the observed increases in sea-level. The volume of Arctic sea ice has also been rapidly declining over the past four decades, and projections indicate that extensive areas in the Arctic will be ice-free during the 21st century. NASA has reported that the extent of Arctic sea ice is decreasing at a rate of 12.9% relative to the 1981 to 2010 average. Overall, Arctic sea-ice cover has declined by ~50% since 1980.

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It is important to differentiate global mean sea-level (GMSL) rise and relative sea-level (RSL) rise. Horton et al. (2018) defined GMSL as the areal mean of either RSL or sea surface height over the global ocean. They defined RSL as the difference in elevation between the sea surface and the land. RSL rise determinations must consider climateinduced GMSL rise, regional sea-level variations, and local non-climate-related sea-level changes (Wong et al. 2014). The rate and magnitude of RSL vary geographically with the occurrence of various drivers of change including: (1) atmosphere/ocean dynamics (e.g., winds and currents); (2) static-equilibrium effects (i.e., gravitational, rotational, and elastic deformational effects) of ocean/cryosphere/hydrosphere mass distribution on the height of the geoid and the earth’s surface; (3) glacio-isostatic adjustment (GIA); (4) coastal subsidence due to sediment compaction, groundwater removal, and oil and gas withdrawals; (5) tectonics (seismic movement); and (6) mantle dynamic topography (Horton et al. 2018). According to Pörtner et al. (2019), there is a high degree of confidence that the rates of GMSL increases were 1.5 mm yr–1 for the period 1902–2010, 2.2 mm yr–1 for 1970– 2015, 3.2 mm year–1 for 1993–2015, and 3.6 mm yr–1 for 2006–2015. More recently, Gulev et al. (2021) reported that GMSL increased at a rate of 1.35 mm yr–1 for the period 1901–1990 to 3.25 mm yr–1 for 1993–2018. Fox-Kemper et al. (2021) noted that sea-level rise has accelerated since the late 1960s at an average rate of 2.3 mm yr–1 for the period 1971–2018 to 3.7 mm yr–1 for 2006–2018. The dominant cause of GMSL rise since 1970 has been attributed to anthropogenic forcing (Dangendorf et al. 2015, 2017; Slangen et al. 2016; Horton et al. 2018; Fox-Kemper et al. 2021). GMSL increased by 0.20 (0.15–0.25) m over the period 1901 to 2018 (Fox-Kemper et al. 2021). It is projected to increase through the 21st century and beyond (IPCC 2021). GMSL rise projections have been estimated for 2050 and 2100 based on physical modeling and statistical techniques. Best estimates of GMSL rise for 2050 range up to 0.5 m. Fox-Kemper et al. (2021) proposed that, relative to the period 1995–2014, GMSL will increase between 0.18 m and 0.23 m by 2050. GMSL rise estimates for 2100 range up to 1.0 m under 1.5°C stabilization and up to 1.1 m under 2.0°C stabilization (Bittermann et al. 2017; Rasmussen et al. 2018). Gulev et al. (2021) also reported that GMSL could increase up to 1.1 m by 2100. Arias et al. (2021) conveyed that GMSL is projected to increase between a median of 0.38 m (0.28–0.55 m likely range) under a lower temperature increase scenario (SSP11.9) and 0.77 m (0.63–1.02 m likely range) under a higher temperature increase scenario (SSP5-8.5) by 2100. Further, a study by Bamber et al. (2019) showed that under a hightemperature scenario, sea-level rise could even exceed 2 m by 2100. A long-term increasing trend of sea-level rise is projected for most coastlines of the world, which will have a marked impact on the structure and function of estuarine and coastal marine ecosystems.

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Sea-level changes are not spatially or temporally uniform but vary over a wide range in large part because anthropogenic climatic and non-climatic drivers are highly variable. As a result, the rate of sea-level rise in many areas can deviate considerably from the global mean value as is evident by RSL rise measurements at the local scale, which often differ from GMSL rise measurements. Geographically variable, anthropogenic non-climatic drivers of change, such as coastal development, land-use changes, altered sediment delivery due to dam building and other structural elements, and human-induced coastal subsidence, cause large variations in local RSL rise that may exceed GMSL rise by an order of magnitude or more. Among the areas experiencing the most rapid changes of RSL are large cities on coastal plains and deltas that are subsiding because of heavy building loads, groundwater withdrawal, oil and gas extraction, sediment accumulation, and compaction (Mazzotti et al. 2009). Examples include the Po delta, eastern Tokyo, and Shanghai. In the USA, the Mississippi River delta is subsiding at a much greater rate than other coastal areas, leading to a high rate of inundation and coastal wetlands habitat loss. White and Tremblay (1995) noted that sea-level rise and subsidence were largely responsible for nearly 60% of the wetland loss along the northern Gulf of Mexico. Jankowski et al. (2017) reported a total land-surface subsidence rate of ~10 mm yr–1 over decadal timescales along the Louisiana coastline. GMSL rise, land-surface subsidence, and sediment deficits are the primary drivers of the rapid coastal land loss in the Mississippi River delta region (Frederick et al. 2019). Vertical crustal movements along active tectonic plate margins can also account for substantial RSL change. For example, the Great Alaskan Earthquake of March 27, 1964, the largest ever recorded in North America (magnitude 9.2), caused abrupt subsidence of extensive areas near Anchorage amounting to ~2.5 m. By contrast, other coastal areas affected by this subduction zone earthquake were raised nearly 10 m (Hansen 1967). Similarly, the Tōhoku Earthquake of March 11, 2011, the largest ever recorded in Japan (magnitude 9.0), resulted in ~0.6 m of coastal subsidence along a 400-km stretch of coastline, with some areas subsiding up to 1.2 m (Norio et al. 2011). Crustal tectonic movements in other regions of the world are not rapid, but sustained, generating considerable RSL change over time. The rate of coastal sea-level rise varies regionally as well due to changing climate and ocean dynamic processes (e.g., air-sea heat and freshwater fluxes, as well as variable winds, air pressure, and ocean currents) (Wong et al. 2014). Regional climate variability occurs on interannual, decadal, and interdecadal timescales driven by large-scale climate phenomena that also affect sea-level (Zhang and Church 2012). For example, the El Niño–Southern Oscillation (ENSO), the dominant source of interannual variability, induces significant sea-level changes in the tropical Pacific Ocean, varying from GMSL by up to 40 cm (Landerer et al. 2008; Walsh et al. 2012). Similarly, the Pacific Decadal Oscillation (PDO) and the Interdecadal Pacific Oscillation

Climate Change and Estuaries

(IPO) are ENSO-like climate variability patterns in the Pacific Ocean occurring on decadal and interdecadal timescales. Zhang and Church (2012), using near-global TOPEX/ Poseidon and Jason altimetry data, ascribed the rapid rates of sea-level rise in the western tropical Pacific Ocean in part to basin-scale decadal climate variability. They also attributed the negligible sea-level rise (or even falling sealevel) in the eastern tropical Pacific Ocean and west coast of the USA to a combination of decreasing sea-levels associated with decadal climate variability and a positive sealevel trend. Glacial isostatic rebound is another driver of RSL rise variation. Coastlines near melting glaciers and ice sheets exhibit falling sea-levels (Milne et al. 2009). This is because of decreasing gravitational attraction of the melting ice sheet on ocean waters as well as rising lands in response to the loss of ice mass, changing shape of the seafloor under the reduced load of ice sheets, and the altering of the earth’s rotation in response to a change in mass distribution (Gomez et al. 2010; Wong et al. 2014).

1.3.3 Other Climatic Drivers Increasing temperature and sea-level rise are accompanied by other climate change effects. Changing wind patterns coupled with rising regional and global temperatures can significantly affect coastal and estuarine circulation patterns, altering upwelling and downwelling, wave heights, and other water movements in the sea. High wind stress driven by extreme weather events, such as tropical and extratropical cyclones and other major coastal storms raises sea-levels, albeit ephemerally, as do astronomical tides. Falling atmospheric pressure during major storms also raises sea-levels. For example, a storm surge along the New Jersey coast (USA) during Hurricane Sandy in October 2012 raised sea-levels by ~3–4 m, impacting estuarine basins. The highest sea-levels observed in coastal regions have occurred when tropical cyclones made landfall concurrently at the time of spring tides, such as in New Jersey during Hurricane Sandy. While there has been an increase in the frequency and intensity of the strongest tropical cyclones since the 1970s, future trends are unclear, although it is very likely that the global mean tropical cyclone precipitation and maximum wind speed will increase (Wong et al. 2014). Future projections of storm surges are uncertain as well because of the vagaries of tropical cyclones and coastal storm characteristics. However, it is likely that storm surges will continue to be an important factor in the genesis of extreme sea-levels locally or regionally on an ephemeral basis. GMSL rise will exacerbate these effects by raising the sea-level platform and causing greater storm surges and inundation impacts on estuaries, coastal landforms, and bayshore communities. Shifts in coastal ocean currents near estuaries could result in significant physico–chemical and biotic changes in these systems. Phytoplankton productivity often significantly

Introduction to Climate Change and Estuaries

increases in areas of greater coastal upwelling and declines in areas where upwelling decreases. Phytoplankton blooms occurring in some wind-driven coastal upwelling areas enhance organic carbon production and elevate BOD, leading to a reduction in dissolved oxygen levels, which has been the recurring case in nearshore ocean waters of New Jersey (USA). Seafloor topographically controlled coastal upwelling centers along the New Jersey coast are co-located with historical regions of low dissolved oxygen (Glenn et al. 2004). Increasing wind stress linked to ongoing climate change will be a potentially significant driver of coastal upwelling events during the 21st century. Predictive models indicate greater global precipitation (up to ~15–20%) in the future; however, for some locations (Southwest USA) more droughts are predicted, which will negatively impact freshwater delivery to estuaries. A significant effect of ocean warming is the increasing frequency and intensity of high rainfall tropical cyclones, which have major impacts on coastal flooding, nutrient/organic matter loading, biogeochemical changes along the estuarine aquatic continuum, and salinity regimes (Figure 1.2) (Paerl et al. 2019). However, there will be important regional differences in precipitation, and significant geographic variation in freshwater runoff as well, with up to a 50% increase in precipitation above 50°N latitude and a 20% decrease between 20 and 50°N and °S latitudes (Frid and Caswell 2017). In contrast, there will be a local reduction

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in freshwater runoff and sediment delivery to estuaries and coastal waters in many areas due to the location of dams, dikes, levees, reservoirs, and other upland anthropogenic structures together with freshwater diversions for agricultural and domestic water use. Although large regional variability exists in freshwater delivery to coastal waters, there is a net global declining trend in freshwater input (Wong et al. 2014). Decreased freshwater delivery reduces the flushing rate, increases salinity, and protracts water residence time in estuaries, affecting the distribution and abundance of organisms. Variability of water residence time strongly influences phytoplankton biomass and community composition, which have important consequences for energy flow through food webs in estuarine ecosystems (Paerl et al. 2006, 2014). For estuaries receiving high freshwater runoff from coastal watersheds in response to hurricanes and other storms, greater sediment and nutrient delivery is characteristic, as is salinity reduction (Paerl et al. 2019). Changes in vertical stratification and mixing affect biotic production and food web structure (Scavia et al. 2002; Kennish 2016). Increasing nutrient inputs promote eutrophication with a greater incidence of algal blooms, reduced dissolved oxygen levels, and loss of seagrass and other essential habitats, particularly in the more susceptible coastal lagoons (Figure  1.2) (Kennish et al. 2007; Howarth et al. 2011; Kennish and Paerl 2010; Paerl and Paul 2012; Paerl 2017;

FIGURE 1.2  Conceptual diagram, adapted from Crosswell (2013), showing the biogeochemical response of a lagoonal estuary (right) to the passage of Hurricane Irene (2011) along the US Atlantic coast (left). Salinity profiles are constructed from direct measurements in the Neuse River Estuary, North Carolina, before and after the storm. Data shown are from August 15, 2011 (A), hypothetical salinity values under well-mixed conditions (B), and August 30, 2011 (C).

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Paerl et al. 2018; Kennish 2019). Biogeochemical changes will affect organisms in bottom sediments as well as in the water column. Deoxygenated conditions in many coastal water bodies have been on the rise for decades concomitant with greater nutrient loading and ongoing climate change (Diaz and Rosenberg 2008; Diaz 2015; Kennish 2017; Breitburg et al. 2018). The structure and function of biotic communities in estuaries are affected by these changes, most notably those in the benthos. For example, altered nutrient recycling in bottom sediments can significantly affect primary production, particularly in shallow estuaries. There are increasing challenges to determining changes in local and regional conditions affecting climate change forcing factors, and the physical, chemical, and biological changes that occur in response to these factors. These challenges require more effective monitoring and modeling of climate change factors, as well as greater resolution of climate and impact predictions (Robins et al. 2016). Local changes in precipitation, occurrences of drought, and resulting shifts in freshwater flows in particular are difficult to predict and directly measure because of the spatial and temporal variability of forcing factors.

1.4 ANTHROPOGENIC NON-CLIMATIC DRIVERS There are numerous anthropogenic non-climatic drivers of change in estuaries that interact with climate drivers to create more deleterious conditions than those attributed to climatic change alone. Many of these drivers are linked to human development and activities in coastal watersheds. Anthropogenic non-climatic drivers can be organized into 12 major categories including: (1) habitat loss and alteration (lagoon construction, shoreline hardening, and land reclamation); (2) watershed impervious cover; (3) enrichment (nutrients, organic carbon, and thermal loading); (4) sewage and pathogenic inputs; (5) chemical contaminants; (6) human-induced sediment/particulate inputs; (7) dredging and dredged-material disposal; (8) human-altered hydrological regimes; (9) invasive/introduced species; (10) overfishing and intensive aquaculture; (11) coastal subsidence; and (12) floatables/plastics/debris (Table 1.1) (see Chapter 5). These 12 major categories can be organized further into three broad classes. Class 1 drivers are those that degrade water quality and are primarily chemical and biological in nature (e.g., nutrient enrichment, organic and thermal loading, chemical contaminants, sediment/ particulate inputs, and pathogens). Class 2 drivers are those that impact habitat and are mainly physical factors (e.g., watershed impervious cover, shoreline hardening, lagoon construction, and land reclamation). Class 3 drivers are those that alter biotic communities; they are linked to multiple drivers (e.g., overfishing, aquaculture, invasive/introduced species, and human-altered hydrological regimes). Chapter 5 provides comprehensive coverage of these classes of drivers.

Climate Change and Estuaries

1.5 INTERACTIVE DRIVERS OF CHANGE Estuaries are highly susceptible to human impacts because of dense human populations inhabiting upstream and proximate coastal watersheds and the multiple drivers originating in both terrestrial and coastal marine environments (Monbaliu et al. 2014). Notably, direct anthropogenic non-climatic factors – land-, estuarine-, and ocean-based impacts – rather than climate-related factors are the primary drivers of change in most estuaries often linked to overdevelopment, overexploitation of resources, freshwater diversions and upstream reservoir construction, chemical pollution, shoreline hardening, and habitat fragmentation or destruction (Kennish 2002, 2016, 2017, 2021, 2022). For example, the input of pollutants from industrial and domestic sources (e.g., sewage and pathogens, fertilizers, metals, persistent organic pollutants, and plastics), freshwater diversions, and hardening of shorelines have adversely affected estuaries, most greatly in urbanized areas. The introduction of invasive species is increasing in estuaries, and these species often outcompete and replace many endemic forms (Austin et al. 2010). Storm surges and rising sea-levels linked to climate change are significant coastal drivers, causing salinization of estuarine waters and saline intrusion of coastal watersheds. These factors can significantly compromise reproductive success, abundance, and distribution of estuarine organisms while concurrently impacting intertidal and subtidal habitats via inundation and erosion. Based on a study of 96 estuaries, Prandle and Lane (2015) found that a sealevel rise of 1 m increases saline intrusion length by >7% in deep estuaries and by >25% in estuaries 6 m). Lagoon-type, bar-built estuaries typically have microtidal ranges. They are characterized by sandy sediments in tidal and river deltas, washover fans, and wave-built structures (i.e., aligned beaches, spits, and baymouth bars). Finer-grained silts and clays accumulate in deeper, more quiescent areas away from high-energy inlet areas. Barnegat Bay–Little Egg Harbor estuary, New Jersey (USA), is a microtidal system, as is Galveston Bay, Texas (USA). Mesotidal estuaries are typified by moderate to strong tidal currents, well-developed tidal deltas, and meandering tidal channels. Sands accumulate in tidal deltas and sand bars, with silts and clays found in tidal flats and wetlands. An example is the Wadden Sea (the Netherlands). Macrotidal estuaries usually have funnel-shaped embayments with linear sand bars and broad marginal tidal flats. Sandy sediments accumulate in sand bars and high-energy central basin areas, whereas finer silts and clays occur in tidal flats and adjoining wetlands. Mangrove swamps may replace tidal flats in tide-dominated systems. Morecambe Bay (England) is an example of a macrotidal estuary. Hypertidal estuaries also typically have a funnel shape. They are characterized by a tidal bore that injects high energy into the system, resuspending and mobilizing bottom sediments, increasing water column turbidity, and dispersing biotic components (e.g., eggs and larvae of organisms). Examples are the Dee estuary in Liverpool Bay (England) and Hangzhou Bay (China) (Tu and Fan 2017).

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3.4.8 Classification Based on Sediment Infilling A sedimentological approach to the classification of estuaries incorporates components of sediment supply and modes of basin infilling. Three categories of estuaries comprise this type of classification: positive-filled, negative-filled, and neutral-filled estuaries (Rusnak 1967). An important determining factor is if an estuary is river-dominated, wavedominated, or tide-dominated (Semeniuk and Semeniuk 2016). A positive-filled estuary is one in which sediment transported by rivers fills the estuarine basin in a manner similar to a prograding delta. The Gironde estuary (France) is an example of this type of estuary. An inverse-filled estuary, in contrast, is one in which the sediments from beaches and the nearshore ocean fill estuarine basins via overwash events and tidal currents. An example is Wash (England) (Evans 1965). Rising sea-levels, coastal storms, and storm surges driven by climate change are increasing the overwash of barrier beaches and infilling of lagoonal estuarine areas. Human-induced global warming and associated sea-level rise (Bindoff et al. 2019) are facilitating landward transgression of lagoon-type, bar-built estuaries (Kennedy 2011). Non-filling, or neutral-filled, estuaries are those in a steady state of equilibrium with little or no change in basin volume due to sedimentation. Sediments may be redistributed by currents in the basin, but little sediment enters the estuary via riverine and marine inputs. Deep fjords in Norway and British Columbia are examples of this type of estuary. The long-term evolutionary trend of an estuary is to progressively infill, unless the estuary is dominated by high fluvial discharges (Kennedy 2011). Each estuary has a characteristic sedimentary character and infill pattern. However, the pattern of infill can be significantly modified by tectonic or isostatic movements. As stressed by Kennedy (2011), the evolutionary trend of an estuary, including its infill pattern, also depends on the response to humaninduced climate change. He presented models of estuarine infilling affected by sea-level rise. Semeniuk (2016) examined estuarine sub-environments that have specific and distinctive suites of sediments along and across an estuarine basin and, with accretion, in its stratigraphy. He also discussed the main sedimentary environments, or facies, occurring in a generalized estuary from landward to seaward reaches. These included the following:

1. Major estuarine-head delta(s) 2. Subsidiary deltas 3. Marginal estuary alluvial fans 4. Spits and their lagoons 5. Marginal estuary environments (platforms, tidal flats, mangrove- and/or salt-marsh-vegetated tidal flats, and biostromes) 6. Central estuary basin, comprising deep-water environments 7. Central estuary channels

Climate Change and Estuaries

8. Tide channels 9. Tide channels fringed by biostromes 10. Central estuary shoals, some vegetated by mangrove and/or salt marsh 11. Margins on the leeward side of a barrier 12. Flood-tidal delta 13. Washover fans Variation occurs among estuarine systems, and therefore one or more of the aforementioned sedimentary environments may be absent in a particular estuary.

3.4.9 Classification Based on Multidisciplinary Criteria Whitfield (1992) employed a multidisciplinary approach to classifying estuaries. He differentiated five types of estuaries based on hydrodynamics, salinity regimes, morphometrics, and mouth dynamics (i.e., estuarine bays, estuarine lakes, river mouths, temporarily opened/closed estuaries, and permanently open estuaries). Roy et al. (2001), working in southeast Australia, also applied a multidisciplinary approach to classifying estuaries but used different criteria, notably a combination of water quality attributes, geological properties, and environmental factors. This classification scheme identified four main types of estuaries, specifically drowned river valleys, mature river estuaries, barrier estuaries, and saline coastal lagoons. Elliott and McLusky (2002) favored a classification scheme that couples water circulation and sediment dynamics to biogeochemistry and ecological features. These processes and features have significant implications for the ecology of estuarine systems. Hume et al. (2007) developed a classification of estuaries based on different scales of multidisciplinary controls on physical and biotic components. Termed the “estuary environment classification,” Hume et al. (2007) identified several key elements that control the physical and biological characteristics of estuaries, namely climatic, oceanic, riverine, and catchment factors. Level 1 in the classification scheme differentiated variation among estuaries due to global scale factors such as latitude, ocean basins, and large landmasses. Level 2 differentiated variation among estuaries based on differences in hydrodynamic processes driven by a river and oceanic forcing as well as basin morphometry. Level 3 differentiated variation among estuaries due to catchment processes controlled by land cover and geology (Whitfield and Elliott 2011).

3.4.10 Classification Based on Anthropogenic Effects Edgar et al. (2000) described a classification approach based on the particular types and levels of anthropogenic impact on different types of estuaries. They investigated 111 estuaries in Tasmania (Australia) initially organizing them into nine groups according to similarities in physical attributes. Differences in river runoff, salinity, tidal range,

Estuaries

seaward barrier, and estuary size were important factors in categorizing the estuarine groups. The level of anthropogenic disturbance to each estuary was determined based on land development and human population data. The highest level of conservation significance was assigned to the least disturbance found in each estuarine group (Whitfield and Elliott 2011). Borja et al. (2011) examined the ecological quality and integrity of estuaries as factors useful for classification development. They focused on methods for classifying estuaries based on anthropogenic stress to organisms. Data on primary community structural variables (i.e., abundance, species richness, and biomass) and derived community structural variables (i.e., diversity indices, biomass ratios, and evenness indices) were vital in determining ecological quality and system integrity. The assessment also included functional analyses (e.g., feeding guilds) and biotic indices (e.g., AZTI’s marine biotic index, AMBI; and the benthic quality index, BQI). The aim was to detect anthropogenic stress on benthic communities and to generate a single evaluation of estuarine ecosystem conditions, while also considering measures of ecosystem structure and function. One significant challenge is that some assessment methods detect both anthropogenically and naturally stressed areas, thereby reducing the capability of detecting and maximizing the signal-to-noise (anthropogenic change to natural change) ratio. Thus, as noted by Borja et al. (2011), “the ecological integrity of an estuary or a lagoon should be evaluated using all information available, including as many biological ecosystem elements as is reasonable, and using an ecosystem-based assessment approach.”

3.5 CONCLUSIONS Numerous definitions and classifications of estuaries have been published in the scientific literature, reflecting in large part the inherent complexity of the spatial and temporal variability of physical, geological, chemical, and biological factors characterizing these important coastal ecosystems. A key feature is the spatial and temporal continua of environmental variables in estuaries (e.g., turbidity, salinity, biotic community structure). There is utility in defining the constituent parts of an estuary as a means to better understand and manage the ecosystem. Because of the complexity and variability of estuarine components, the classification schemes have been largely arbitrary (Elliott and McLusky 2002). Estuaries differ considerably in attributes (e.g., origin, physiography, morphometrics, hydrodynamics, etc.), which is the basis of the diverse classification schemes. In addition, investigators formulating estuarine classifications have been motivated by different perspectives in their work (e.g., landform geomorphology, evolutionary origins, and formative processes), purposes (e.g., understanding structure, variability and dynamics, functions and values, and interactions with adjoining fluvial and coastal ecosystems), and applications (e.g., categorizing,

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mapping, and management) over diverse spatial and temporal scales (Dyer 1990; Simenstad and Yagani 2011). Valle-Levinson (2010, 2011) conveyed that the traditional approaches of classification have focused on: (1) estuarine origin or geomorphology, (2) water balance, (3) competition between tidal flow and river discharge, and (4) stratification and water circulation. Anthropogenic factors have taken on greater significance due to an expanding world population in the coastal zone (Kennish and Elliott 2011; Kennish 2016). Many estuarine classifications have been developed for local case study systems, focusing on certain disciplinary elements. Fewer interdisciplinary classifications have been developed. Various drivers of change in estuaries require greater quantification and assessment useful for classification development. Detection and attribution of environmental change in estuaries can be particularly problematic because it is difficult to identify and differentiate the effects of anthropogenic forcing factors (e.g., the input of land-based pollutants, shoreline hardening, dredging, overfishing, and anthropogenic climate change) from effects of natural environmental factors that vary significantly in space and time (Kennish 2021, 2022; Elliott et al. 2022). Effects of multiple driver interactions can be synergistic, additive, or antagonistic on estuarine biotic communities, resulting in beneficial, neutral, or detrimental organism responses (Crain et al. 2008; Boyd et al. 2018). This adds complexity to system analyses. Trajectories of climate change and human activities will increase environmental variation in estuaries during the 21st century. Joint influences of climatic drivers and direct anthropogenic drivers are evident in altered freshwater flows, sediment budgets, biogeochemical changes (e.g., nutrient and dissolved oxygen flux), and shifts in species composition and distribution (Kimmerer and Weaver 2013). Extreme weather events, storminess, precipitation, and higher temperatures predicted with ongoing climate change will accelerate surface runoff, sediment erosion, turbidity, and pathogens in combined sewer overflows, as well as the mobilization of persistent organic compounds, metals, and other pollutants from coastal watersheds to estuarine basins. Higher inputs of nutrients in freshwater discharges will stimulate organic carbon production and harmful algal blooms (HABs), leading to greater biological oxygen demand and the reduction of dissolved oxygen (hypoxic or anoxic conditions) in estuarine basins (Kennish and Elliott 2011; Kennish 2016; see Chapters 8, 17, and 27). Eutrophication of estuarine waters will escalate with climate change (see Chapter 17). The toxicity, mobility, and fate of contaminants will also change with increasing temperature and salinity and decreasing dissolved oxygen and pH, posing a potential threat to ecologically sensitive habitats (Sheahan et al. 2013). Acidification problems will worsen (see Chapter 9). Intense rainfall and storm surge events will cause greater inundation and flooding of coastal watersheds and communities, adversely affecting ecosystem services and community infrastructure.

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Classifications are an important tool for improving the understanding of estuarine structure, function, and processes. However, the development of estuarine classifications faces significant challenges, most notably: (1) integrating classifications of the structure of estuary/coastal landscapes at different scales and relating their organization to ecosystem processes; (2) categorizing the magnitude and direction of dynamic change in estuary/coastal class and landscape structure, and the significance to ecosystem processes; and (3) classifying systems that relate physical and biogeochemical attributes of estuaries to their social, cultural, and economic attributes (Edgar et al. 2000; Simenstad and Yagani 2011). Valle-Levinson (2011) noted that attempting to fit an estuary into a specific type of classification is challenging because of significant fluxes in its properties both temporally and spatially. Shifting conditions can place estuaries into different categories of classification even over relatively short time intervals. Of particular note, climate change is a major forcing factor modulating estuarine conditions and, hence, will be of increasing importance in the classification of estuaries.

ACKNOWLEDGMENTS This is Publication Number 4619 of the Department of Marine and Coastal Sciences, Rutgers University, New Brunswick, New Jersey, USA. Special thanks to Michael Elliott of the University of Hull, Hull, UK for providing data sources on estuarine classifications.

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53 Pritchard, D. W. 1967. Observations of circulation in coastal plain estuaries. Pp. 37–44. In: G. H. Lauff (ed.), Estuaries, Vol. 83. Washington, DC: American Association for the Advancement of Science. Psuty, N. P. and D. D. Ofiara. 2002. Coastal Hazard Management: Lessons and Future Directions from New Jersey. New Brunswick: Rutgers University Press. Romao, C. 1996. Interpretation manual of European Union habitats, Version EUR15. European Commission, DGXI (Environment, Nuclear Security and Civil Protection), Brussels. Roy, P. S., R. J. Williams, A. R. Jones, I. Yassini, P. J. Gibbs, B. Coates, R. J. West, P. R. Scanes, J. P. Hudson, and S. Nichol. 2001. Structure and function of south-east Australian estuaries. Est. Coastal Shelf Sci. 53(3): 351–384. Rusnak, G. A. 1967. Rates of sediment accumulation in modern estuaries. Pp. 180–194. In: G. H. Lauff (ed.), Estuaries, Vol. 83. Washington, DC: American Association for the Advancement of Science. Schubel, J. R. and D. J. Hirchberg. 1978. Estuarine graveyard, climatic change, and the importance of the estuarine environment. Pp. 285–297. In: M. L. Wiley (ed.), Estuarine Interactions. New York: Academic Press. Semeniuk, V. 2016. Stratigraphy of estuaries. Pp. 177–187. In: M. J. Kennish (ed.), Encyclopedia of Estuaries. Dordrecht: Springer. Semeniuk, V. and C. Semeniuk. 2016. Deltas. Pp. 623–648. In: M. J. Kennish (ed.), Encyclopedia of Estuaries. Dordrecht: Springer. Sheahan, D., J. Maud, A. Wither, C. Moffat, and C. Engelke. 2013. Impacts of climate change on pollution (estuarine and coastal). MCCIP Sci. Rev. 2013: 244–251. Simenstad, C. and T. Yanagi. 2011. Introduction to classification of estuarine and nearshore coastal ecosystems. Pp. 1–6. In: C. Simenstad and T. Yanagi (eds.), Treatise on Estuarine and Coastal Science, Vol. 1. Classification of Estuarine and Nearshore Coastal Ecosystems, Treatise on Estuarine and Coastal Science. Oxford: Elsevier. Stommel, H. 1951. Recent developments in the study of tidal estuaries. Technical Report No. 51-33, Woods Hole Oceanographic Institution, Woods Hole, Massachusetts. Stommel, H. 1953. Computation of pollution in a vertically mixed estuary. Sew. Ind. Wastes 25: 1065–1071. Tagliapietra, D., M. Ssigovini, and A. V. Ghirardini. 2009. A review of terms and definitions to categorise estuaries, lagoons, and associated environments. Mar. Freshw. Res. 60(6): 497–509. Tu, J. and D. Fan. 2017. Flow and turbulence structure in a hypertidal estuary with the world’s biggest tidal bore. J. Geophys. Res. Oceans 122(4): 3417–3433. Tomczak, M. 1996. The shelf and coastal zone. Lecture notes. http://www​.cmima​.csic​.es​/ mirror​/ mattom​/ ShelfCoast​/ chapter11​.html. Valle-Levinson, A. 2010. Definition and classification of estuaries. Pp. 1–11. In: A. Valle-Levinson (ed.), Contemporary Issues in Estuarine Physics. Cambridge: Cambridge University Press. https://doi​.org​/10​.1017​/CBO9780511676567​.002. Valle-Levinson, A. 2011. Classification of estuarine circulation. Pp. 75–86. In: C. Simenstad and T. Yanagi (eds.), Treatise on Estuarine and Coastal Science, Vol. 1. Classification of Estuarine and Nearshore Coastal Ecosystems, Treatise on Estuarine and Coastal Science. Oxford: Elsevier. Venice System. 1958. Symposium on the classification of brackish waters, Venice April 8-14, 1958. Arch. Oceanogr. Limnol. 11 (Suppl.): 1–248.

54 Whitfield, A. K. 1992. A characterization of southern African estuarine systems. J. Aquat. Sci. 18: 89–103. Whitfield, A. and M. Elliott. 2011. Ecosystem and biotic classifications of estuaries and coasts. Pp. 99–124. In: C. Simenstad and T. Yanagi (eds.), Treatise on Estuarine and Coastal Science, Vol. 1. Classification of Estuarine and Nearshore Coastal Ecosystems, Treatise on Estuarine and Coastal Science. Oxford: Elsevier. Williams, S. 2009. Past, present, and future sea-level rise and effects on coasts under changing global climate. Pp. 37–46. In: D. Lavoie (ed.), Sand Resources, Regional Geology and

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4

Sea-Level Rise and Estuaries John A. Church and Xuebin Zhang

CONTENTS Abstract................................................................................................................................................................................. 55 4.1 Introduction................................................................................................................................................................. 56 4.2 Pre-historical Sea-Levels............................................................................................................................................. 56 4.3 Historical Changes in Sea-Level Rise and Extreme Events........................................................................................ 57 4.4 Understanding Trends in Sea-Level Change...............................................................................................................60 4.5 Sea-Level Projections for the 21st Century................................................................................................................. 61 4.5.1 Global Projections to 2100............................................................................................................................... 63 4.5.2 Regional Projections to 2100........................................................................................................................... 65 4.5.3 Evaluation of Models for Projection of Sea-Level Change............................................................................. 65 4.5.4 Other Causes of Regional Sea-Level Change and Variability.........................................................................66 4.5.5 Extreme Events................................................................................................................................................ 67 4.6 Longer-Term Commitments and Uncertainties........................................................................................................... 67 4.7 Impacts......................................................................................................................................................................... 68 4.8 Conclusions.................................................................................................................................................................. 69 Acknowledgments................................................................................................................................................................. 69 References............................................................................................................................................................................. 70

ABSTRACT Over a billion people, many of the world’s mega cities, and much of the world’s agriculture, industry, infrastructure, fisheries, and tourism are located around estuaries. Our modern society has developed in a period of stable sealevels and with sea-levels actually falling along much of the world’s coastline. The long timescales of the oceans and ice sheets mean they are now out of balance with recent anthropogenic increases in greenhouse gas concentrations, and, as a result, sea-levels are rising at an accelerating rate. From 1900 to 2018, the global mean sea-level rose at a rate of 1.7 mm yr–1, an order of magnitude faster than in previous centuries, increasing the frequency and severity of extreme sealevel events. The two largest contributions to the observed rise were from ocean thermal expansion and the loss of mass from glaciers, with smaller but accelerating contributions from the ice sheets of Greenland and Antarctica. There is now a clear indication that anthropogenic climate change is the main reason for the observed rise. The projected likely sea-level rise for 2100 ranges from 0.28–0.55 m, if greenhouse gas emissions are consistent with the Paris Agreement target of limiting global warming to 1.5°C, to 0.63–1.02 m if emissions continue to rise rapidly. For this latter scenario, the rate of rise during the last two decades of the 21st century will be equivalent to that during the last major deglaciation of the earth when sea-level rose at rates of more than 1 m per century for many millennia. Including uncertain ice sheet instabilities could result in a rise of 0.63 to 1.61 m (or more) by 2100. High emissions would commit DOI: 10.1201/9781003126096-5

the world to ongoing, and likely essentially irreversible, sea-level rise amounting to meters over the coming centuries. The regional projections are mostly within about 20% of the global mean for the majority of the world’s non-polar coastline. The changes in mean sea-level are rapidly emerging above the background of natural climate variability. Even for 1.5°C of global surface warming, many locations will experience the current 1-in-100-years extreme sea-level event at least once a year by 2100. Large-scale sea-level variability can be skillfully predicted up to several seasons in advance in parts of the Pacific and Indian oceans, mainly associated with dynamics of El Niño–Southern Oscillation. The reduction of sediment supply and sediment compaction, particularly following water or petroleum extraction, have had negligible impacts on the average global coastal rate of rise. However, it can be substantial and larger than the climate change sea-level rise in some regions of large population densities. Sea-level rise will have a major impact on the natural environment as well as on human populations. We cannot stop all sea-level rise, and we will need to adapt to the sea-level rise we can no longer prevent. Adaptation options include: (1) retreat from the coastline abandoning parts of the land to the ocean; (2) defend valuable parts of the coastline, likely to lead to estuarine ecosystems being “squeezed out”; and (3) adapt to rising sea-levels with appropriate local and regional planning. The amount of adaptation necessary is strongly related to future greenhouse gas emissions. There will be difficult trade-offs between expense, safety, conservation, and economic development. Addressing 55

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these challenges requires open and informed discussion. Avoiding meters of sea-level rise, which would significantly impact coastal and estuarine ecosystems and greatly challenge our ability to adapt, requires significant, urgent, and sustained reduction of our greenhouse gas emissions. Key Words:  Sea-level rise, Global and regional projections, Extremes, Commitments, Impacts, Adaptation.

4.1 INTRODUCTION Sea-level change has been a major control on estuaries throughout the earth’s history, from before the ancient continent of Gondwanaland through to the present configuration of continents (which was established more than 20 Ma), and will continue into the future. Over the last 1 million years, sea-levels have risen and fallen by more than 100 m during ice-age cycles (Rohling et al. 2009). During ice ages, large ice sheets formed over North America, northern Europe and northern Asia, and the Antarctic and Greenland Ice Sheets were also larger. As a result of this large transfer of mass from the oceans to the land, during ice ages, sea-levels were more than 100 m below current day levels. These sea-level changes also facilitated and hindered the migration of modern humans out of Africa and around the world, with many stories of sea-level change still surviving in our oldest cultures (Nunn 2014). Today, much of the world’s population, many of the world’s megacities, and much of the world’s agriculture, industrial developments and infrastructure, fisheries, and tourist attractions are located near the coast, and often around presentday estuaries, many of which were formed less than 10,000 years ago during the late Holocene epoch. Recent estimates are that over a billion people live in the low elevation coastal zone (LECZ) defined as the area within 10 m of current-day sea-levels, with 770 million of these within 5 m of sea-level (Kulp and Strauss 2019), and more than 200 million people live within the current 1-in 100-years coastal flood plain (Nicholls et al. 2021). These numbers (and the percentage of the world’s population living in the LECZ) are continuing to grow as people continue to move to the coasts in both the developed and developing nations despite ongoing rising sealevels (Neumann et al. 2015). Our modern industrial society has mostly developed in a period of relatively stable and low rates of sea-level change, and with sea-levels actually falling along much of the world’s coastline from ongoing vertical land motion (Glacial Isostatic Adjustment (GIA), Peltier 1998, 2004) as a result of changes in surface loading of the earth as major ice sheets of the last glacial maximum (~20,000 years ago) flowed into the ocean. However, in some areas (e.g., northeastern USA), GIA causes ongoing land subsidence resulting in sea-level rise relative to the land prior to the recent climate change-related rise. The long timescales of the oceans and ice sheets mean they are out of balance with recent anthropogenic increases in greenhouse gas concentrations and the resulting climate

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change. Sea-levels are now rising at an accelerating rate around the vast majority of the world’s coastlines (e.g., IPCC 2013, 2019, 2021; Wang et al. 2021a). These past emissions guarantee ongoing sea-level rise (Church et al. 2013b; FoxKemper et al. 2021) and will result in dramatic increases in coastal flooding events. Even for a global surface warming of 1.5°C (the preferred target of the Paris Agreement), many locations will experience the current 1-in-100 years extreme sea-level event at least once a year before 2100, with the tropics more sensitive than northern high latitudes (Tebaldi et al. 2021). Sea-level change is a multidisciplinary science involving the study of the oceans, ice sheets, glaciers, land water changes, and the atmosphere and has a wide range of economic, environmental, and social impacts. Here, we give a brief introduction to the main factors underlying current and projected sea-level change and its impacts on estuaries and coasts. For a more complete discussion of the issues, the reader is referred to the comprehensive compilation of Church et al. (2010) and recent Intergovernmental Panel on Climate Change (IPCC) assessments (Church et al. 2013b; Wong et al. 2014; Oppenheimer et al. 2019; Fox-Kemper et al. 2021).

4.2 PRE-HISTORICAL SEA-LEVELS While it is well known that sea-levels were lower during past cold (glacial) periods, paleo evidence also indicates sealevels were higher during past warm (interglacial) periods (Dutton et al. 2015). During the Mid Pliocene Warm Period (3.3 million years ago), sea-levels are estimated to have been between 5 and 25 m (or more) above the present-day level, at carbon dioxide concentrations similar to today and temperatures between 2.5–4°C warmer than during 1850–1900 (Oppenheimer et al. 2019; Dutton et al. 2015). During the last interglacial period (LIG; 129,000–116,000 years ago), sealevel is estimated to have been 5–10 m above the present level at temperatures 0.5–1.5°C warmer than 1850–1900 (Gulev et al. 2021; Masson-Delmotte et al. 2013; Church et al. 2013b; Fox-Kemper et al. 2021). The higher values indicate that there must have been a contribution to LIG sea-level from the loss of mass of the Antarctic Ice Sheet as well as the Greenland Ice Sheet. These paleo data, indicating that sea-levels were higher during warm interglacial periods, are important in evaluating models of ice sheet evolution and thus projecting future sea-level change. However, they are not a direct analog for future sea-levels, and there is recent work suggesting the current sea-level estimates for the last interglaciation may be too high (Dyer et al. 2021). At the peak of the last ice age, about 20,000 years ago, sea-levels were ~130 m below current day sea-levels (Lambeck et al. 2014). As the earth warmed coming out of the last ice age (23,000 to 19,000 years ago), global sea-level rose at a rate of more than a meter per century for many thousands of years, and with a peak rate of over 4 m per century (Figure 4.1; Lambeck et al. 2014). The rate of rise slowed around 7,000 years ago. For the last two millennia,

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FIGURE 4.1  Ice-volume and equivalent sea-level (esl) change. (A) Individual esl estimates (blue) and the objective estimate of the denoised time series (red line). The inset gives an expanded scale for the last 9,000 yr. (B) The same esl estimate and the major climate events in the interval [the last glacial maximum (LGM), Heinrich events H1 to H3, the Bölling–Allerod warm period (B–A), and the Younger Dryas cold period (Y–D)] as well as the timing of Melt Water Pulse-1A, 1B, and the 8.2 ka BP cooling event. (C) The 95% probability range of the esl estimates. (D) Estimates of sea-level rate of change. (From Lambeck et al. 2014)

prior to the industrial revolution, sea-level change was small, with no indication of oscillations of global mean sealevel greater than 15–20 cm over periods greater than 200 years (Masson-Delmotte et al. 2013; Lambeck et al. 2014). Recent sea-level reconstructions of proxy sea-level data (Kopp et al. 2016) indicate that global mean sea-level rose by 0.1±0.1 mm yr–1 over 0–700 Common Era (CE) and then fell at 0.2±0.2 mm yr–1 over 1000–1400 CE, with the latter associated with a global mean cooling of about 0.2°C. The rapid rise in sea-level after the last glacial maximum means that many of today’s estuaries are drowned river valleys. As well as determining the present configuration of our coasts and estuaries, the large transfers of mass between the oceans and the land as the ice sheets waxed and waned resulted in deformation of the earth’s surface. During ice ages, the earth’s surface becomes depressed under the weight of the ice sheets due to the visco-elastic response of the mantle. Part of the mantle below the ice sheets flowed sideways to form peripheral forebulges around the extremities of the ice sheets. As the ice sheets decayed, the lithosphere below the former ice sheets rebounded and the peripheral forebulges began subsiding. These multi-century trends of vertical land motion (due to GIA) as a result of slow adjustments to past surface load changes on the earth are still ongoing today as the stresses in the mantle gradually relax (Peltier 1998). As a result, on millennial time scales, there have been clear differences in the rate of relative sea-level change (ocean surface relative to the land). In some regions of former ice sheets during the last glacial maximum (e.g., Hudson Bay, Baltic Sea), prior to the 20th century, sea-level (relative to the land) has been falling

despite the volume of water in the ocean remaining essentially constant (Figure  4.2). In contrast, in the peripheral forebulge regions immediately adjacent to former ice sheets (such as the northeast coast of the USA and Canada) relative sea-level has been rising (Figure 4.2) over the last several thousand years. In the far field distant from the location of present and former ice sheets, relative sea-level has been falling as a result of ongoing GIA (Figure 4.2). These ongoing land motions and little change in global mean sea-level (GMSL) means that our modern industrial society has mostly developed in a period of relatively stable and low rates of sea-level change, and with sealevel actually falling along much of the world’s coastline (Figure 4.2). We now face major challenges as sea-levels are rising at an accelerating rate around the vast majority of the world’s coastlines as a result of anthropogenic climate change.

4.3 HISTORICAL CHANGES IN SEALEVEL RISE AND EXTREME EVENTS The first continuous direct measurements of coastal sea-level began with the installation of tide gauges in some European ports in the 18th century, with the longest records indicating an acceleration in the rate of sea-level rise (Woodworth 1999) from the preindustrial rates of change of less than a few tenths of a millimeter per year. The number of gauges increased during the 19th and 20th centuries (mostly in support of shipping activities). These tide gauge records (Holgate et al. 2013) indicated that the rate of global mean sea-level rise began accelerating in the second half of the

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FIGURE 4.2  Changing rates in sea-level (mm yr–1) from ongoing glacial isostatic adjustment (GIA) (Peltier 2004). (a) Changes in sea-level relative to the land as would be measured by tide gauges. (b) Change in geocentric sea-level (relative to the center of the earth) as would be measured by satellite altimeters.

Climate Change and Estuaries

19th century (Church and White 2006, 2011; Hay et al. 2015; Kopp et al. 2016; Dangendorf et al. 2017). As a result, sealevels rose an order of magnitude faster during the 20th century than over previous millennia. From 1900 to 2018, the most recent assessment (Gulev et al. 2021; Fox-Kemper et al. 2021) is that global mean sea-level rose at an estimated rate of 1.73 mm yr–1, with a very likely (5–95%) range of 1.28 to 2.17 mm yr–1 (Figure 4.3). The tide gauge data also indicates an acceleration in the rate of rise of 0.0053 mm yr–2, with a very likely range of 0.0042 to 0.0073 mm yr–2 from 1900 to 2010. From 1970 to the near present, this acceleration has increased to 0.06 mm yr–2 (Dangendorf et al. 2019; Wang et al. 2021a). The 20th-century sea-level rise is extremely likely (probability >0.95) to have been faster than during any century of the last three millennia (Kopp et al. 2016). The GIA trends described above directly impact present and future sea-level change, with some parts of the northeast coast of the USA experiencing a 20th-century rise about double the global average rise (Piecuch et al. 2018). A recent analysis of the paleo sea-level records indicates that in the North Atlantic region, the acceleration of sea-level first became apparent in the Mid-Atlantic Bight and later farther south and north, and also in northern Europe (Walker et al. 2022). The first (nearly) global measurements of sea-level came with the launch of satellite altimeters in the 1970s and 1980s, with continuing direct observations of nearglobal (rather than just coastal sea-level from tide gauges,) sea-level data available since 1993 following the launch of the Topex-Poseidon satellite in late 1992. The data from

FIGURE 4.3  Global mean sea-level from 1900 to 2018 (or the end of the respective record) for five different estimates of global mean sea-level using coastal tide gauges, and the ensemble average (black) of these five estimates. The inset shows the same records from 1993 compared to the satellite altimeter record (red).

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the Topex-Poseidon and follow-on satellite altimeter missions indicate that global mean sea-level rise had increased to 3.25 mm yr–1 (2.88 to 3.61 mm yr–1) from 1993 to 2018 and is continuing to accelerate (0.094 mm yr–2, with a very likely range of 0.082 to 0.115 mm yr–2) (Gulev et al. 2021; Fox-Kemper et al. 2021; Wang et al. 2021a). These data indicate the rate of global mean sea-level rise is now larger than 4 mm yr–1 over the last decade (WCRP Global SeaLevel Budget Group 2018; Wang et al. 2021a). Analysis of coastal tide gauges also shows a similar increase in the rate over the satellite period (Wang et al. 2021a). The satellite altimeter data has revealed large-scale patterns of interannual to decadal variability in geocentric sea-level (i.e., sea-level measured relative to the center of the earth), particularly in the low-latitude Pacific Ocean (e.g., Zhang and Church 2012). These patterns are primarily related to variations in the zonal winds (Merrifield et al. 2012; Lyu et al. 2017) and the related variations in ocean steric sea-level rise (Wu et al. 2017). Natural variability in sea-level is large on interannual to decadal time scales and tends to obscure long-term trends. However, averaged over the full satellite altimeter record to date, the trends indicate geocentric sea-level is rising virtually everywhere, but with significant regional variations mostly related to the natural climate variability, such as El Niño–Southern Oscillation (ENSO) (Figure 4.4a). However, zonally averaged observed sea-level changes (Figure 4.4b) are consistent with climate-change simulations within uncertainties and cannot be explained by internal variability alone, indicating an anthropogenically forced contribution is detectable at global scales (Richter et al. 2020).

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The frequency and severity of extreme sea-level events have increased over recent decades, both regionally (e.g., Church et al. 2006; Marcos et al. 2009; Grinsted et al. 2012; Sweet and Park 2014; Haigh et al. 2015; Paprotny et al. 2018) and globally (e.g., Woodworth and Blackman 2004; Menéndez and Woodworth 2010; Marcos et al. 2019). The main driver of the resultant increase in flooding has been the changes in (local) mean sea-level, including the influence of seasonal, climate variability, and ocean eddy variability (Woodworth and Blackman 2004; Menéndez and Woodworth 2010, Woodworth and Menéndez 2015; Barnard et al. 2015), with much more limited evidence of changes in storminess (e.g., in the western North Pacific, Oey and Chou 2016; along the US east coast, Grinsted et al. 2012). There is only low confidence that anthropogenic changes in tropical cyclones (TCs) have been detected and attributed during the historical record (Knutson et al. 2019). Climate change impact on TCs from 1980 to 2018 was found to be more evident in the spatial pattern of TC occurrence, rather than the number of TCs (Murakami et al. 2020). Flooding from the combination of extreme events and sea-level rise is not restricted to the coastline but also occurs far from the ocean along estuaries and tidal lakes (e.g., Hanslow et al. 2018; Lopes et al. 2022). As well as inundation, rising sea-levels are associated with salt-water intrusion and an increase in erosion, with a substantial proportion of the world’s sandy beaches already eroding (Vousdoukas et al. 2020b) and other impacts on estuarine and coastal ecosystems. (See other chapters in this volume for a more complete discussion of these impacts.)

FIGURE 4.4  Geocentric sea-level trend over 1993–2020 (mm yr–1) as measured by satellite altimetry: (a) spatial map (b) zonal average. Note effects of inverse barometer and GIA have been applied before calculating the trend. The dots are the rate of relative sea-level change at coastal tide gauges for the same period. Vertical land motion is a major (but not the only) contributor to the differences in relative and geocentric sea-levels.

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4.4 UNDERSTANDING TRENDS IN SEA-LEVEL CHANGE For many years, the observed sea-level rise during the 20th century was an enigma with no adequate explanation for the observed rise, especially in light of the constraint imposed by changes in the observed earth rotation (the earth’s rotation was expected to change in response to significant loss of mass from high latitude ice; Munk 2002). However, Mitrovica et al. (2006) showed this constraint is weaker than previous estimates because of responses deeper within the earth. Also, there are now improved estimates of the observed rise and the many components contributing to it. Our understanding has developed considerably following the commencement of the satellite altimeter record in 1993, satellite measurements of changes in the mass of the ocean in 2002, and high-quality comprehensive near-global measurements of ocean warming in 2006. As a result, we now have a reasonably good understanding of sea-level change (i.e., closing the sea-level budget) over the 20th century and particularly in recent decades. Note the impacts of sea-level change are caused by relative sea-level change (sea-level relative to the moving land, Figure 4.5) as measured by coastal tide gauges. In contrast, satellite altimeters measure geocentric sea-level (i.e., sea-level with respect to the center of mass of the earth, Figure 4.5). The differences between these two sealevels (Figure 4.4a) are often small, but can be critical to understanding and projecting sea-level change. Particular progress was made in the studies explaining the rise in global mean sea-level since the 1960s (Church et al. 2011b; Moore et al. 2011), for the complete 20th century (Gregory et al. 2013; Frederikse et al. 2020, although significant uncertainties remain for the first half of the century), for the satellite altimeter period since 1993 (Chen et al. 2017; WCRP sea Level Budget Group 2018), for regional

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sea-level change since the 1950s (Frederikse et al. 2020), and local sea-level change since the 1950s (globally, Wang et al. 2021b; in the USA, Harvey et al. 2021). The two largest contributions to sea-level rise since 1900 are ocean thermal expansion (ocean water expands as it warms and as a result sea-level rises) and the loss of mass from glaciers, including the peripheral glaciers on Greenland (Figure 4.5). It is important to note that over 90% of the increase in heat content in the earth’s climate system is stored in the ocean, implying ocean thermal expansion (and hence sea-level rise) is a primary indicator of climate change (Rhein et al. 2013). There were also smaller contributions from the ice sheets of Greenland and Antarctica, but these have been accelerating significantly over the last three decades (Chen et al. 2017; Shepherd et al. 2018, 2020; Fox-Kemper et al. 2021). There are also changes in the mass of water stored on land as a result of natural climate variability, and human interference in the water cycle through the extraction of water from aquifers (much of which makes its way to the ocean) and the storage of water in terrestrial reservoirs (see the discussion in Church et al. 2013b and Fox-Kemper et al. 2021). See FoxKemper et al. (2021) and Wang et al. (2021b; Figure 4.6a) for the latest global mean sea-level budget since 1960. The land water storages are important in understanding the reasons for the 20th-century change, but are a relatively small contribution to projections for the 21st century (Church et al. 2013b; Fox-Kemper et al. 2021). Each of these contributions has a regional pattern (Figure 4.5 and 4.6c–h) as well as a global mean contribution. The ocean sterodynamic sea-level change (SDSL, defined as the global mean thermal expansion and its regional distribution due to ocean dynamics (see Gregory et al. 2019 for a definition of sea-level terms), is directly related to changes in surface ocean fluxes of heat, freshwater, and wind stress. The barystatic sea-level changes (that is resulting from changes

FIGURE 4.5  Climate-sensitive processes and components that can influence global and regional sea-level. Changes in any one of the components or processes shown will result in a sea-level change. The term “ocean properties” refers to ocean temperature, salinity, and density, which influence and are also dependent on ocean circulation. Both relative and geocentric sea-level vary with position. Note that the geocenter is not shown. (From IPCC AR5 Chapter 13, Church et al. 2013b)

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in the mass of the oceans) also have a regional distribution with a greater than the global average sea-level rise in the far fields from the location of ice/water loss (on land) and a smaller or even a fall in sea-level closer to the regions of mass loss. This regional pattern is a result of changes in the earth’s gravity, the earth’s rotation, and the viscoelastic solidearth deformation (GRD) as the surface loading of the earth changes (Gregory et al. 2019). This GRD is the contemporary equivalent of the ongoing GIA sea-level changes from past mass redistributions and is slightly different for relative and geocentric sea-levels (Peltier 1998; Frederikse et al. 2017). As well as these global features, relative sea-level change is affected by local factors such as the compaction of sediments and the extraction of water and petroleum products (see Section 4.5.4). The sum of each of the contributions (depicted in Figure 4.6), together with an estimate of vertical land motion from other sources (e.g., caused by sediment compaction, particularly in deltaic regions and/or where water or petroleum products are extracted), agree with the observed sealevel rise at 272 globally distributed tide gauges to within 90% confidence limits for the 1958 to 2015 period (Wang et al. 2021b). Sterodynamic sea-levels and glaciers were the dominant contributions to both global and coastal sea-levels with smaller contributions from ice sheets (although these are increasing) and changes in land water storage. Over this period, the sum of contributions over the total ocean area and along the world’s coastline is 1.5 and 1.1 mm yr–1, respectively. The smaller coastal rise is principally a result of an offsetting contribution from ongoing GIA, particularly in regions of former ice sheets (such as in the Baltic Sea) and in the far field. Globally, GIA is the dominant cause of regional differences in relative sea-level trends at 272 tide gauges over the last 50 years. However, away from the regions of former ice sheets, the sterodynamic sea-level is the dominant cause of regional differences in trends, with GIA being a secondary contributor (see Wang et al. 2021b for a more detailed discussion). The warming of the oceans and the related ocean thermal expansion, the loss of mass from glaciers and the Greenland Ice Sheet (increased melting is larger than increased snowfall) are all directly related to surface temperature increases (e.g., Slangen et al. 2012; Church et al. 2013b). For Antarctica, analysis of data since the early 1990s (Shepherd et al. 2018) indicates that accelerated ice discharge (resulting from the warming of ocean waters beneath ice shelves; mainly in west Antarctica) is larger than the increased snowfall (also related to increased atmospheric temperatures), resulting in a net positive contribution to sea-level rise. There is now a clear indication that anthropogenic climate change is the main reason for the observed warming and sea-level rise. While natural regional and local variability in sea-level is large on interannual to decadal time scales and tends to obscure the long-term trends, models of the various contributions forced by anthropogenic factors, together with observational estimates of the ice sheet

and land water contributions can now reasonably represent global mean sea-level change over the 20th century (Church et al. 2013a; Slangen et al. 2017) and its regional distribution (Meyssignac et al. 2017). Different forcing factors (greenhouse gases, aerosols, ozone changes, and natural variability) result in different temporal and spatial patterns of ocean change (Slangen et al. 2015; Bilbao et al. 2015; Fasullo et al. 2020). These patterns have been used to detect the anthropogenic influence in ocean temperatures (Marcos and Amores 2014; Slangen et al. 2014; Weller et al. 2016) and in zonal means of sea-level change over 1993–2015 (Richter et al. 2020). The anthropogenic signal has also been detected in the glacier contribution to sea-level change for the latter 20th century (Marzeion et al. 2014) with a more recent analysis indicating the anthropogenic signal is present throughout the 20th century (Roe et al. 2021). Since 1970, at least 70% of the global mean sea-level rise is a result of a warming climate from anthropogenic climate change (Slangen et al. 2016).

4.5 SEA-LEVEL PROJECTIONS FOR THE 21ST CENTURY There are many components to bring together to make global-mean and regional sea-level projections for the 21st century and beyond in response to changing greenhouse gas concentrations, volcanic activity, and solar variability (Slangen et al. 2012; Church et al. 2011a; Church et al. 2013a; Fox-Kemper et al. 2021; Figure 4.5). Climate models (coupled atmosphere–ocean–land surface models) are used to evaluate sea-level changes resulting from the warming of the oceans, and changes in ocean circulation and density. These sterodynamic sea-level changes have a global mean component (global averaged ocean thermal expansion) and a regional distribution, as sea-level change is not uniform around the globe (dynamic sealevel change). These regional changes are associated with changing ocean density and circulation resulting from changing interactions with the cryosphere and atmosphere (for example changing surface winds). Additional models are used to evaluate changes in the mass of glaciers, ice sheets, and land water changes and thus the mass of the ocean. These mass changes result in both global mean sea-level changes and regional GRD fingerprints in sealevel change as a result of the changing the shape of the earth and the earth’s gravitational and rotational fields in response to these large mass transfers (see Figure 4.6 for the regional distributions from historical mass transfers). There are also large-scale ongoing changes to the earth following the loss of ice sheets at the end of the last glacial maximum (ongoing GIA) and regional/local tectonic changes from earthquakes, compaction of sediments, etc. Many uncertainties remain about the uncertainties of Antarctic Ice Sheet's contributions to sea-level change (Church et al. 2013b; Oppenheimer et al. 2019; Fox-Kemper et al. 2021) and the regional distribution of sterodynamic sea-level change (Gregory et al. 2016).

FIGURE 4.6  (a) GMSL time series (mm) from the ensemble mean of different reconstructions (orange; shading area indicating 90% CL), the sum of all contributions (purple), SDSL (red), glaciers (blue), ice sheets (yellow), and TWS (green). (b) The trend (mm yr–1) of GMSL and individual contributions over 1958–2015, with the same colour defined in (a). The Error bars indicate 90% CL. Regional trend (mm yr–1) maps of RSL over 1958–2015 from individual sea-level components, including (c) SDSL with inverse barometer effect corrected, (d) the total contemporary barystatic-GRD fingerprint, (e) GIA, (f) glaciers, (g) ice sheets (Greenland and Antarctic), (h) TWS, and the sum of all contributions (i). Colored circles in (i) denote trends from TG records with other VLM corrections. (From Wang et al. 2021b)

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4.5.1 Global Projections to 2100 Global Climate Projections have been evaluated for a range of possible future greenhouse gas emission scenarios. The Intergovernmental Panel on Climate Change (IPCC) Fifth Assessment Report (AR5; IPCC 2013) and the IPCC Special Report on the Ocean and Cryosphere in a Changing Climate (SROCC; IPCC 2019) used four different scenarios, called Representative Concentration Pathways (RCPs) in an attempt to span potential future possibilities. The lowest scenario (RCP2.6) requires strong mitigation of greenhouse gas emissions and negative emissions after about 2070, whereas the highest scenario (RCP8.5) represents the continued growth of emissions beyond 2100. Two intermediate scenarios are RCP4.6 and RCP6.0. The IPCC Sixth Assessment Report (AR6; IPCC 2021) used five Shared Socioeconomic Pathways (SSPs), ranging from SSP1-1.9 (designed to be consistent with limiting global warming to the preferred Paris Target of 1.5°C), SSP1-2.6 (similar to RCP2.6), SSP2-4.5 (similar to RCP4.5), SSP3-7.0, and SSP5-8.5 (similar to RCP8.5). The long timescales of the oceans and ice sheets mean that they are out of balance with current greenhouse gas concentrations. As a result, past greenhouse gas emissions have not only resulted in sea-level rise during the 20th and early 21st centuries but also guarantee ongoing sea-level rise. Projections of the amount of sea-level rise before 2050 are a result of past and ongoing greenhouse gas emissions but are only weakly dependent on greenhouse gas emissions between now and 2050 (Church et al. 2013b; Fox-Kemper

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et al. 2021). However, these near-term emissions impact the rate of acceleration of sea-level rise (Church et al. 2013b; Fox-Kemper et al. 2021; Wang et al. 2021a) and thus lock in a commitment to a greater sea-level rise after 2050, by 2100 and beyond (Figure 4.7). Under all scenarios, the projections over the 21st century are larger than the observed sea-level rise over the 20th century (Church et al. 2013b; Fox-Kemper et al. 2021). These projections include growing contributions from ocean thermal expansion and loss of mass from glaciers and ice sheets, with additional contributions from changes in terrestrial storage. After 2050, projected sea-level rise is strongly dependent on future emission scenarios and continues to accelerate under high-emission scenarios. For strong mitigation of greenhouse gas emissions (RCP2.6) consistent with the central value of projected warming relative to pre-industrial of less than 2°C, the projection by the IPCC AR5 (Church et al. 2013b) is a 0.44 m rise with a range of 0.28 to 0.61 m relative to 1986–2005 (Table 4.1). For this scenario, the rate of rise from 2081–2100 is about 4.4 mm yr–1, having stabilized mid-century and then decreasing slowly. For continuing emission consistent with a 3°C (RCP6.0) and more than 4°C (RCP8.5) temperature rise above 1850–1900 by 2100, the projected sea-level rise in 2100 is larger at 0.55 m (0.38–0.73 m) and 0.74 m (0.52–0.98 m). Importantly, for these scenarios, the rate of rise from 2081–2100 is much larger at 7.4 mm yr–1 (4.7–10.3 mm yr–1) and 11.2 mm yr–1 (7.5–14.7 mm yr–1), and is continuing to increase, implying a commitment to rapidly rising sea-levels over subsequent

FIGURE 4.7  Global mean sea-level change in meters, relative to 1900. The historical changes are observed (from tide gauges before 1992 and altimeters afterwards), and the future changes are assessed consistently with observational constraints based on emulation of CMIP, ice-sheet, and glacier models. Likely ranges are shown for SSP1-2.6 and SSP3-7.0. Only likely ranges are assessed for sea-level changes due to difficulties in estimating the distribution of deeply uncertain processes. The dashed curve indicates the potential impact of these deeply uncertain processes. It shows the 83rd percentile of SSP5-8.5 projections that include low likelihood, high-impact ice sheet processes that cannot be ruled out because of low confidence in projections of these processes. This curve does not constitute part of a likely range. Changes relative to 1900 are calculated by adding 0.158 m (observed global mean sea-level rise from 1900 to 1995–2014) to simulated and observed changes relative to 1995–2014. (From IPCC 2021 AR6 Summary for Policy Makers)

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decades and centuries. These latter values are comparable to the rate of sea-level rise during the last major deglaciation of the earth when sea-levels rose at over a meter per century for many thousands of years. The IPCC AR5 recognized the potential for marine ice sheet instabilities in Antarctica to lead to a substantially larger rise, particularly for high emission scenarios, but was unable to make robust projections. Based on limited evidence, it suggested this “additional contribution would not exceed several tenths of a meter of sea-level rise during the 21st century.” The likely ranges of the IPCC AR6 global sea-level projections (Table 4.2, Figure 4.7) are similar to those of the AR5. The AR6 projections are relative to 1995–2014 (different from the reference period of 1986–2005 used by AR5) and also include a lower emission greenhouse gas scenario (SSP1-1.9) consistent with a global mean surface warming by 2081–2100 of 1.4°C (1.0–1.8°C). These projections benefited from an improved constraint on the knowledge of the climate sensitivity to greenhouse gases and the first internationally coordinated ice sheet and glacier model intercomparison projects (Nowicki et al. 2016; Hock et al. 2019). Most importantly, this resulted in a slightly larger and more robust estimate for the projected Antarctic contribution than those of the AR5. For SSP1-1.9, the projected global-mean sea-level rise by 2100 is 0.38 m (0.28–0.55

m), whereas for SSP5-8.5 the rise is essentially double this value at 0.77 m (0.63–1.02 m) (Table 4.2, Figure 4.7). The AR6 did not complete projections equivalent to RCP6.0 but included SSP3-7.0 (equivalent to a warming of about 3.6°C (2.8–4.6°C) by 2081–2100 (Fox-Kemper et al. 2021). Including ice sheet instabilities, in which there is only low confidence, the IPCC AR6 projection under SSP5-8.5 is larger at 0.88 m (0.63 to 1.61 m) by 2100, and with a rapidly rising rate (15.9 (8.8–30.2) mm yr–1) at that time. Note that the uncertainty ranges in both the amount and rate of sea-level rise from Antarctica are not symmetric, with the possibility of substantially larger rates and amounts of rise for high emission scenarios. In the AR6, a rise approaching 2 m by 2100 and 5 m by 2150 under the high SSP5-8.5 could not be ruled out because of the deep uncertainty in ice sheet processes (IPCC 2021). Note that pre-October 2021 national policies from nations around the world imply warming above preindustrial temperatures (approximated by average 1850–1900 temperatures) of 2.9–3.2°C, with current pledges implying 2.4–2.9°C of warming. Recent promises of many nations to have net zero emissions (covering 72% of global emissions) imply 2.0–2.4°C of global warming, potentially bringing the Paris Agreement goal of well below 2°C within reach, if they are fully and swiftly implemented (Höhne et al. 2021), and followed by further actions.

TABLE 4.1 IPCC AR5 projections of global mean sea-level in 2100 compared to 1986–2005

Global Mean rise (m) Rate of rise (mm/yr)

Additional potential Antarctic contribution for high emissions

RCP2.6

RCP6.0

RCP8.5

0.44 [0.28 to 0.61] 4.4 [2.0 to 6.8]

0.55 [0.38-0.73] 7.4 [4.7-10.3]

0.74 [0.52 to 0.98] 11.2 [7.5 to 15.7]

Several tenths of a metre

Source: Church et al. (2013b). Note: The numbers in each box are the central estimate of sea-level rise, with the likely range (17–83%) given in brackets and the estimated rate of rise over 2081 to 2100 given on the second line. The AR5 noted the potential of an additional several tenths of a meter sea-level contribution from possible but uncertain instabilities of the Antarctic Ice Sheet, particularly for the high emission scenarios. (Note, for comparison to the AR6 projections, about 0.03 m needs to be subtracted from the AR5 projected rise in 2100 because the AR6 projections are relative to a later base period of 1995 to 2014.)

TABLE 4.2 IPCC AR6 projections of global mean sea-level in 2100 compared to 1995–2014

Global Mean rise (m) Rate of rise (mm/yr)

SSP1-1.9

SSP1-2.6

SSP3-7.0

SSP5_8.5

SSP5-8.5 Low confidence

0.38 (0.28-0.55) 4.3 (2.5-6.6)

0.44 (0.33-0.61) 5.3 (3.3-8.1)

0.68 (0.55-0.90) 10.4 (7.5-14.9)

0.77 (0.63-1.02) 12.2 (8.8-17.7)

0.88 (0.63-1.61) 15.9 (8.8-30.2)

Source: Fox-Kemper et al. (2021). Note: The numbers in each box are the central estimate of sea-level rise, with the likely range (17–83%) given in brackets, and the estimated rate of rise over 2080–2100 given on the second line. The AR6 noted the potential of an additional low confidence contribution from possible but uncertain instabilities of the Antarctic Ice Sheet for the high emission scenarios.

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4.5.2 Regional Projections to 2100 The regional projections of sea-level change flow directly from the global projections and have been developed for the globe and a number of regions (e.g., McInnes et al. 2015). The regional pattern associated with ocean dynamics comes directly from global climate models, with a greater than average rise near the poleward boundaries of the subtropical gyres, in the northwest Atlantic associated with the weakening of the Atlantic meridional overturning circulation (AMOC), in the Arctic Ocean and lower rise in the high latitude Southern Ocean (Yin et al. 2009; Church et al. 2013b; Bilboa et al. 2015; Lyu et al. 2020; Fox-Kemper et al. 2021). The regional pattern associated with mass changes comes from the evaluation of changes in the earth’s gravity, earth’s rotation, and viscoelastic solid-earth deformation (Church et al. 2013b; Fox-Kemper et al. 2021). Particularly for high emission scenarios, the sea-level fingerprint from ice-sheet mass loss results in greater than global averaged sea-level rise far from the ice sheets. However, the regional differences in the projections of sea-level change (including land motion from GIA, but not other vertical land motion) in the IPCC AR5 are mostly within about 20% of the global mean projections for the majority of the world’s non-polar coastline (Figure 4.8; Church et al. 2013b). Vertical land motion results from tectonics, sediment compaction, petroleum, and water extraction as well as GIA. The regional sea-level projections from the AR6, including estimates of vertical land motions (based on Kopp et al. 2014), are available for the global ocean and at coastal tide gauges (https://sealevel​.nasa​.gov​/ipcc​-ar6 ​-sea​-level​ -projection​-tool). However, note that Fox-Kemper et al. (2021) caution that there is only “low to medium confidence

in the GIA and VLM projections employed in this report. In many regions, higher fidelity projections would require more detailed regional analysis.” The projected changes in mean sea-level are rapidly emerging above the background of natural climate variability (Lyu et al. 2014; Richter and Marzeion 2014). For example, Lyu et al. (2014) estimate that the Time of Emergence (ToE; the time when local sea-levels will be significantly larger and emerge from background variability) of regional sea-level rise with all contributing components considered will occur by 2020, with respect to 1986–2005 levels, for 50% of the global ocean. This ToE for sea-level rise is substantially earlier than that for surface air temperatures, which means coastal and estuarine regions will experience (and are already experiencing) the impacts of sea-level rise. IPCC AR5 or AR6 sea-level projections provide useful sea-level information on large scales (>100 km) mainly based on global climate model simulations, but do not resolve regional- or local-scale (5–10 km) processes.

4.5.3 Evaluation of Models for Projection of Sea-Level Change The ability to explain (Church et al. 2011b; Gregory et al. 2013; Wang et al. 2021b) and simulate (Church et al. 2013a,b, Slangen et al. 2017; Meyssignac et al. 2017) past global and regional sea-level changes gives confidence in the process-based projections of future sea-level change based on dynamical models. However, these historical simulations do not include independent ice sheet simulations, and thus are an incomplete evaluation of the models. Also, the sea-level projections are difficult to rigorously evaluate as the short overlapping period between observations

FIGURE 4.8  The IPCC AR5 regional projections (meters) of relative sea-level change between 1986–2005 and 2081–2100 for (a) RCP2.6, (b) RCP4.5, (c) RCP6.0, and (d) RCP8.5. These patterns do not contain an allowance for vertical land motion other than GIA.

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and projections and natural climate variability make the detection of trends and accelerations challenging. Wang et al. (2021a) attempted to critically evaluate the IPCC AR5 and SROCC sea-level projections with observed GMSL and coastal sea-level from a global network of tide gauges, and an allowance for local vertical land motion. They found that the observed trends from GMSL (Figure 4.9) and the regional weighted mean at tide-gauge stations confirm the projections under three RCPs within a 90% confidence level during 2007–2018. The central values of the observed GMSL (1993–2018) and regional weighted mean (1970–2018) accelerations are larger than projections for RCP2.6 and lie between (or even above) those for RCP4.5 and RCP8.5 over 2007–2032, but are not yet statistically different from any scenario. While the confirmation by Wang et al. (2021a) of the projection trends gives us confidence in our current understanding of near-future sealevel change, it leaves open questions concerning possible non-linear accelerations from ice-sheet contributions. There are a broad range of challenges in accurately projecting regional sea-level change. First, we do not yet know how rapidly our global society will limit greenhouse gas emissions in response to the 2015 Paris agreement. Perhaps the most important scientific issues for projecting regional sea-levels are: the sensitivity of the climate system to greenhouse gas levels (and thus projections of global mean sealevel), the regional distribution of sea-level change, and the future, particularly on the longer term (see Section 4.6), of the ice sheets of Greenland and particularly Antarctica. Regional to estuary scale projections cannot currently be done with fine enough resolution with global climate models and thus require the implementation of high-resolution regional hydrodynamical models or regional ocean general circulation models with boundary conditions coming from the global models and surface boundary conditions from

Climate Change and Estuaries

high-resolution atmospheric fields (often via dynamical downscaling; e.g., Zhang et al. 2017; Hermans et al. 2020; Jin et al. 2021; Chaigneau et al. 2022).

4.5.4 Other Causes of Regional SeaLevel Change and Variability After climate change, the largest scale sea-level events are naturally occurring climate variability such as the El Niño– Southern Oscillation (ENSO) phenomena in the Pacific and Indian oceans. The large scale and slowly evolving nature of this variability means that there is potential for the prediction of sea-levels months in advance. Over the last decade, real-time prediction of seasonal to interannual changes in mean sea-level, that strongly influence coastal erosion and flooding (Barnard et al. 2015), is becoming a useful tool in the tropical Pacific and Indian Oceans and along the west Australian and American coasts. These predictions initially used statistical techniques (Chowdhury et al. 2014) but have more recently relied on real-time dynamical models of climate variability allowing regional to local mean sea-levels to be skilfully predicted up to several seasons in advance in some regions (Miles et al. 2014; McIntosh et al. 2015; Widlansky et al. 2014; Long et al. 2021) in the Pacific and Indian Ocean regions. Other modes of ocean and climate variability, including on decadal time scales (Han et al. 2019), and on shorter time scales associated with the movement of ocean eddies (Firing and Merrifield 2004), also result in sea-level changes but the prediction of these sealevels is less well advanced. There are a number of regional/local causes of sea-level rise and its impacts. As rivers are dammed, there is a reduction of sediment supply to the estuaries and the coast (e.g., Syvitski et al. 2009) impacting the natural environment such as coastal wetlands. This results in increases in both

FIGURE 4.9  Monthly observed GMSL from satellite altimeter (1993–2018) and GMSL reconstructions (1970–2018) smoothed with a five-month running mean filter, and annual multi-model ensemble mean GMSL from the AR5 and SROCC projections (2007–2032) under three RCPs. GMSL trends including both linear and quadratic terms are also shown (offset). The blue shaded area indicates the overlapping period between observations and projections (2007–2018). (From Wang et al. 2021a)

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coastal flooding and erosion, as the loss of the sediment supply slows and in some cases prevents the ongoing building of coastal marshes and deltaic regions. In many coastal regions, particularly the heavily populated deltaic regions, sediments are also compacting over time, and this compaction is exacerbated by the extraction of ground water (e.g., Syvitski et al. 2009). Similarly, the extraction of petroleum products leads to a sinking of the land thus further exacerbating climate change–related sea-level rise in some local regions such as Galveston in the Gulf of Mexico. A recent analysis (Nicholls et al. 2021) indicates that these non-climate related anthropogenic effects have had negligible impact on the average rate of sea-level rise rate along the global coast line over the past two decades. However, they are critically important for society as the coastal population density is much larger in regions where society is exacerbating climate change–related sea-level rise with regional and local anthropogenic activities. They estimate the average coastal population has experienced a relative sea-level rise of up to four times the global average, with some regions (notably Southeast Asia where coastal population densities are large) experiencing a populationweighted rise of as much as 20 mm yr–1. An example of the impact of a large local sea-level rise is that Indonesia is currently planning to relocate the capital from Djakarta to an area less prone to local sea-level rise. Population-weighted rates are also high in south and east Asia and parts of northern Africa. Nicholls et al. (2021) also point out that by controlling ground water extraction (as is already practiced in many regions) we can reduce the number of people likely to experience a 1-in-100-years coastal flood by 2050 by tens of millions.

4.5.5 Extreme Events At the regional and estuary scale, changing sea-levels are felt most acutely through the impacts of extreme events. The height of these extreme events is impacted by phenomena over a wide range of space scales from global (tides, large-scale climate change, and variability) to local features (regional and local storm surges) and on time scales from centuries and decades (climate change and variability) to days and hours (tides, storm surges, and waves). Rising sea-levels can result in more frequent coastal flooding (see Section 4.3, sometimes referred to as nuisance flooding as roads are intermittently closed; Sweet and Park 2014) or exacerbate catastrophic events (such as Hurricane Katrina in New Orleans and Hurricane Sandy in New York and New Jersey (Strauss et al. 2021)). Hunter (2012) developed a simple method based on sealevel projections and historical extreme sea-level characteristics for estimating the height that protection measures need to be raised (termed allowances) to prevent the current risk of flooding of the coastal environment from increasing as sea-level rises. Because of the uncertainty of the projections, even for a given greenhouse gas scenario, this allowance is somewhat larger than the projected mean sea-level

rise. These allowances include changes in mean sea-level and are useful in evaluating potential impacts on estuarine ecosystems, but they do not include changes in the extreme sea-level characteristics (such as magnitude/frequency of storm surges). As the mean sea-level rises, it may interact with other extreme sea-level driving factors to cause nonlinear interactions up to tens of centimeters between sealevel rise, tide, storm surges, and waves, which can be either positive or negative depending on the environment (Idier et al. 2019). Projections of changes in future extreme events require high-resolution models to resolve regional sea-level change and to simulate the impact of tides and extreme wind and wave events (e.g., Muis et al. 2020). Projections of surface waves have begun to reveal changes that can be expected during the 21st century (Hemer et al. 2013; Morim et al. 2018). Accurate estimates of the impact of rising sea-levels on flooding from extreme events in estuarine environments are better estimated by numerical downscaling techniques rather than the much simpler “bath-tub” approach which can overestimate the impacts (Lopes et al. 2022). It is also important to include the non-linear interaction of tides and storm surges if local impacts are not to be overestimated (e.g., Idier et al. 2019; Arns et al. 2020; Xiao et al. 2021). A potential further complication is “compound events” resulting from the interaction of coastal and fluvial flooding, especially in estuaries with large catchment areas, which can be examined via hydrodynamic modeling considering both factors (Kumbier et al. 2018; Bevacqua et al. 2019). Recent work on estimating the impact of extreme events has also recognized the need to consider higher order (skewness and kurtosis) changes in the probability density function of sea-level, as well as changes in the mean and the standard deviation (Jin et al. 2022). Recent analysis has included the impact of tides and changes in storm surges and waves to determine the impact of rising sea-levels on a global scale (Vousdoukas et al. 2018). The changes in projected mean sea-levels have a dramatic effect on extreme levels, as is already observed. Tebaldi et al. (2021) estimate that even for 1.5°C of global surface warming many locations will experience the current 1-in-100 years extreme sea-level event at least once a year before 2100, with the tropics more sensitive than northern high latitudes. The high end of these projections has major implications for the cost of potential damages. They have recently been used to demonstrate that the costs of coastal damage can be significantly reduced along the European coastline by raising dykes (Vousdoukas et al. 2020a) and highlight the risk of major losses of the world’s sandy beaches (Vousdoukas et al. 2020b).

4.6 LONGER-TERM COMMITMENTS AND UNCERTAINTIES Under all climate change scenarios, sea-levels will continue to rise for centuries, even after greenhouse gas emissions cease, and at even faster rates if net emissions are

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not reduced to zero (Levermann et al. 2013b; Clark et al. 2016). The ocean is the principal storage of heat in the climate system and it takes many centuries for the ocean heat content to approach equilibrium with changed atmospheric temperatures, as the surface warming gradually penetrates to deeper levels (Wu et al. 2021). AMOC, which plays a significant role in redistributing heat and salt content in the global oceans and in particular producing deep waters, is projected to weaken in coming centuries (e.g., Caesar et al. 2018). Non-polar glaciers have a limited mass (of the order of 40 cm of sea-level equivalent) and therefore a limited long-term contribution to sea-level rise. The ice sheets of Greenland and Antarctica contain over 60 m of equivalent sea-level rise and thus have the greatest potential for longterm sea-level rise. For high emission scenarios, the question is not if there will be meters of sea-level rise but rather how quickly it will occur (See the discussions in Church et al. 2013b; Fox-Kemper et al. 2021). Global mean sea-level projections to 2300 in the IPCC AR6 under RCP2.6 are for a sea-level rise of 0.5 to 3 m and under RCP8.5 a rise of 2 to 7 m. The AR6 also projects that a rise of more than 15 m cannot be ruled out (IPCC 2021), but there is only low confidence in these estimates. Sea-level rises under high emission scenarios are effectively irreversible on centennial and potentially millennial time scales and could be significantly determined by warming thresholds occurring before 2100. The oceans will continue to warm for centuries even if greenhouse gas concentrations are stabilized. For Greenland (presently holding about 7.3 m of sea-level equivalent), the ice sheet decreases in volume with warming. For example, for sustained warming of 2.5°C (a warming we could attain during the 21st century without strong mitigation), Gregory et al. (2020) estimate the Greenland Ice Sheet alone might lose the equivalent of about 4 m of sea-level rise over millennia, and would not regrow to its present size even if late 20th– century climate was restored. For Antarctica, the contribution under high emissions is uncertain but could amount to many meters as a result of ice sheet instabilities in both West and East Antarctica and would be essentially irreversible (Golledge et al. 2015; Fox-Kemper et al. 2021). The regional pattern of sterodynamic sea-level rise approximately scales with ocean heat content (and global temperatures) during the 21st century (Bilbao et al. 2015; Wu et al. 2021) as the upper layers of the ocean warm. Combined with the patterns of ice sheet contributions, this results in the patterns of the relative sea-level rise being largely independent of climate scenario (Figure 4.8) until larger ice sheet contributions become dominant. However, once atmospheric greenhouse gas concentrations begin to stabilize (or even decrease), the sterodynamic pattern is globally more uniform as the deeper layers of the ocean continue to warm for centuries (Wu et al. 2021). The regional patterns of sea-level rise from ice sheet contributions are relatively well understood once the ice sheet contributions are determined and can be derived from

Climate Change and Estuaries

solving the sea-level equation (Mitrovica et al. 2011; Gomez et al. 2010). These patterns consist of a much lower rise or even a fall in sea-level near the locations of ice mass loss and a greater than global average rise (by up to 30%) in the far field. For example, for New York, the regional sea-level rise from a Greenland contribution is only about 40% of the global average, but for a west Antarctic contribution, it is about 20% larger than the global mean rise (Mitrovica et al. 2009). On the longer multi-century time scale and with larger ice-sheet contributions, these spatial fingerprints will begin to dominate.

4.7 IMPACTS Rising relative sea-levels from climate change and other anthropogenic activities and sea-level variability will directly impact many people living adjacent to the coast and estuaries, as well as estuarine and coastal ecosystems. Recent estimates are that currently, about 1 billion people live in an area less than 10 m above current high tide levels (the Low Elevation Coastal Zone; LECZ), including 250 million less than 1 m (Kulp and Strauss 2019). For a lowemission scenario, the number of people currently living below the projected high tide level in 2100 is about 190 million, up from 110 million today. For a high emission scenario and for a rapid Antarctic ice sheet contribution, the number of people subject to annual flooding (with today’s population distribution) is up to about 630 million, up from 380 million today. The largest numbers are for tropical regions, particularly Asia, where large populations in deltaic regions are at risk (Kulp and Strauss 2019; Hooijer and Vernimmen 2021). While the uncertainties are significant, the numbers are large and increasing as people move towards the coast in both developed and developing nations (Neumann et al. 2015). Even today, the costs of sea-level rise are substantial. For example, it is estimated that more than 12% of the $US60 billion costs of Hurricane Sandy are attributable to sea-level rise from anthropogenic climate change (Strauss et al. 2021). Significant cultural and historical sites are also at risk (Marzeion and Leveramann 2014). As well as permanent inundation and flooding, estuaries and coasts are subject to saline intrusion into ground waters and erosion. A substantial fraction of the world’s sandy beaches are currently eroding, with major beach loss, many in heavily populated regions, projected to occur by 2100 (Vousdoukas et al. 2020b). A thorough review of the impact of changes in sea-level and surface waves on saline intrusion and estuarine and coastal erosion is beyond the scope of the present chapter. However, we note that fine-grained mud (as is more common in estuaries) can be transported away resulting in greater long-term erosion than might occur for sandy beaches. A wide range of coastal ecosystems are already under stress from ocean warming, acidification, and sea-level rise, exacerbated by pressures from human activities. As well as anthropogenic sea-level rise, there are many other processes

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affecting relative sea-level change. Rising relative sea-levels can cause natural ecosystems such as mangroves and salt marshes to migrate landward. However, when combined with land-use changes, sea-level rise can squeeze out coastal and estuarine ecosystems caught between a retreating coast line and advancing human developments (Bindoff et al. 2019; Oppenheimer et al. 2019). These processes have already led to a loss of global wetland area by nearly 50% relative to preindustrial levels (Bindoff et al. 2019). The natural environment may be further impacted by increased salinity intrusion with higher sea-levels (Bindoff et al. 2019). Climate change impacts are not limited to relative sealevel change. Other potential impacts go beyond the present scope and include: increasing ocean and coastal temperatures and related changes in the frequency and intensity of coastal marine heatwaves (Collins et al. 2019; Marin et al. 2021); changes in the intensity, frequency, range, and increased precipitation of typhoons/cyclones; changes in coastal winds affecting coastal currents, local upwelling conditions, and thus the supply of oceanic nutrients; waves; ocean stratification and oxygen levels; and changes in coastal protection as coral reefs and other aspects of the natural environment are affected (see Bindoff et al. 2019 and Oppenheimer et al. 2019 for a more complete discussion). And of course, the impacts on society are dependent on how we manage our coasts and estuaries, and their development. There are three main coastal adaptation options (Nicholls 2006; and see Oppenheimer et al. 2019 for a more thorough discussion): • Retreat from the coastline, abandoning parts of the land to the ocean. This is already happening in some locations around the world, and it is simply too expensive to protect all the estuarine and coastal environments. This option also allows estuarine ecosystems to migrate landward. • Defend valuable parts of the coastline with either soft (nature-based) solutions or hard engineering solutions, as exemplified by the Thames Barrier, protecting many billions of pounds of infrastructure in London, and the dykes of the Netherlands and other northern European nations, or even build out into the ocean reclaiming land (e.g. the Netherlands and Singapore). There are important cultural and historical sites (such as Venice) to be considered. This approach is likely to lead to estuarine ecosystems being “squeezed out.” • Adapt to rising sea-levels with appropriate local and regional planning by designing infrastructure to cope with rising sea-levels and allowing for regular flooding. There are many social, cultural, and political challenges, and each of these options has difficult trade-offs between expense, safety, conservation, and economic development.

Generally, stakeholders have a wide range of potentially conflicting interests and have differing abilities to influence decision making. Making sound choices requires careful and long-term planning based on a solid understanding and projections of sea-level rise, education of all participants of the risks and benefits of each of the options, and legislation across multiple (national, state, and local) policy levels, and governance domains (Oppenheimer et al. 2019). Perhaps most important is open and informed discussion.

4.8 CONCLUSIONS We cannot stop all sea-level rise, and we will need to adapt to the sea-level rise we can no longer prevent. However, after about 2050, the amount of adaptation necessary and the ability to adapt are strongly related to the world’s future greenhouse gas emissions. Higher emissions, including the current trajectories that are well above the 2015 Paris Agreement targets, commit the world to larger and more rapid rises and the likelihood of crossing thresholds during the 21st century leading to ongoing, and likely essentially irreversible, sea-level rise amounting to meters over coming centuries. These larger rates of rise make adaptation much more difficult, whereas slower rates of rise provide greater flexibility and a wider range of adaptation options (Oppenheimer et al. 2019). Earlier mitigation of greenhouse gas emissions will result in a greater reduction in sea-level rise by 2100 than a delayed reduction in emissions. Global surface temperatures are directly related to the sum of greenhouse gas emissions (IPCC 2013). However, for sea-level rise by any given time-frame, the amount of sea-level rise is proportional to the integral of the time greenhouse gas emissions are in the atmosphere, such that an early reduction in emissions result in a smaller sea-level rise than later emission reductions of the same total amount (Bouttes et al. 2013; Church et al. 2013b; Mengel et al. 2018; Fox-Kemper et al. 2021). Thus, delayed mitigation leads to a larger sea-level rise in the same future time (e.g., 2100). There are already enough fossil fuels available for potentially crossing critical thresholds leading to meters of sea-level rise. Avoiding these scenarios and thus avoiding meters of sea-level rise, which would greatly challenge our ability to adapt and significantly impact coastal and estuarine ecosystems, requires a significant, urgent, and sustained reduction of present-day greenhouse gas emissions.

ACKNOWLEDGMENTS JC was supported by the Australian Research Council's Discovery Project funding scheme (project DP190101173) and the Australian Research Council Special Research Initiative, Australian Centre for Excellence in Antarctic Science (Project Number SR200100008). XZ was supported by Australian National Environmental Science Program (NESP) Climate Systems Hub.

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5

Anthropogenic Drivers of Estuarine Change Michael J. Kennish

CONTENTS Abstract................................................................................................................................................................................. 75 5.1 Introduction................................................................................................................................................................. 75 5.2 Estuarine Anthropogenic Impacts............................................................................................................................... 78 5.2.1 Habitat Loss and Alteration............................................................................................................................. 79 5.2.2 Dredging and Dredged-Material Disposal...................................................................................................... 79 5.2.3 Enrichment......................................................................................................................................................80 5.2.3.1 Nutrients and Eutrophication............................................................................................................80 5.2.3.2 Organic Matter..................................................................................................................................80 5.2.3.3 Thermal Loading.............................................................................................................................. 81 5.2.4 Sewage Wastes and Pathogens......................................................................................................................... 83 5.2.5 Chemical Contaminants.................................................................................................................................. 83 5.2.6 Human-Altered Hydrological Regimes........................................................................................................... 83 5.2.7 Human-Induced Sediment/Particulate Inputs................................................................................................. 85 5.2.8 Introduced/Invasive Species............................................................................................................................ 85 5.2.9 Overfishing and Intensive Aquaculture........................................................................................................... 87 5.2.10 Coastal Subsidence.......................................................................................................................................... 88 5.2.11 Floatables/Plastics/Debris............................................................................................................................... 88 5.2.12 Climate Change............................................................................................................................................... 89 5.3 Environmental Impact Factors.................................................................................................................................... 91 5.4 Impact Remediation..................................................................................................................................................... 91 5.5 Conclusions.................................................................................................................................................................. 93 Acknowledgments.................................................................................................................................................................94 References.............................................................................................................................................................................94

ABSTRACT Estuaries rank among the most biologically productive and heavily impacted coastal ecosystems on earth. Numerous direct anthropogenic impacts occur in estuarine environments largely attributed to pollution inputs from land-based sources, physical alteration and loss of habitats, and overexploitation of resources. The interactions of direct anthropogenic drivers of change and escalating climatic drivers cause both acute and insidious adverse effects in estuaries. Degraded water quality and disrupted habitats lead to decreases in organism abundance and diversity as well as shifts in the trophic organization of biotic communities that are detrimental. The altered structure and function of estuarine ecosystems result in declining human use. The most extensive anthropogenic activities, such as in highly urbanized and industrialized regions, can greatly compromise the ecological integrity and sustainability of these vital coastal ecosystems. Climate change during the 21st century poses increasing challenges for policymakers and coastal managers to remediate environmental impacts. Sound policies and DOI: 10.1201/9781003126096-6

management strategies are essential to mitigate and control these human-induced impacts, concurrently with field applications to improve estuarine conditions via the implementation of effective restoration, protection, and conservation programs that enhance ecosystem sustainability. Key Words:  estuaries, climatic and non-climatic drivers of change, water quality, pollution, habitat loss, ecosystem structure and function, restoration, conservation, resilience, sustainability

5.1 INTRODUCTION Estuarine environments are susceptible to a wide range of anthropogenic impacts mainly because of increasing human population growth, development, and activities in coastal regions of the world. Approximately 40% of the more than 8 billion people in the world live within 100 km of the coast (Kennish 2019), with many of them dependent on the substantial ecologic and economic provisions of these valuable environments (Barbier et al. 2011; Costanza 75

76

et al. 2014). Estuaries have exceptional recreational and commercial value; provisioning services of estuarine environments support many sectors of society that are fundamentally important to human well-being (Costanza et al. 2014; Kubiszewski et al. 2017; Wolanski et al. 2019; Elliott et al. 2022). Costanza et al. (2007) estimated the value of estuarine ecosystem services as high as $22,000 ha–1 yr–1. Coastal ecosystem human services are valued in the trillions of dollars and include recreational and commercial fisheries, tourism, mariculture, electric power generation, oil and gas operations, transportation, shipping, as well as natural substances used in the production of specialty chemicals and foods, medicines, and pharmaceuticals. The natural capital asset value of the Mississippi River Delta (USA) alone is estimated to be $330 million to $1.4 trillion (Batker et al. 2010). Similar values are estimated for other coastal ecosystems worldwide (Costanza et al. 2007, 2014, 2017). Many of the largest cities in the world are located near estuarine and coastal marine waters, including New York, London, Calcutta, Shanghai, and Tokyo. Land areas surrounding the New York Bight Apex and Tokyo Bay rank among the most populated and industrialized in the world (Day et al. 2012). Urbanized estuarine environments, particularly near metropolitan centers, are most susceptible to anthropogenic impacts which can significantly reduce societal goods and services. Estuaries are among the most temporally and spatially variable ecosystems and one of the most biologically productive on earth. Many estuarine and marine organisms utilize an array of habitats (open water, bottom sediments, submerged aquatic vegetation, mangroves, tidal flats and creeks, fringing wetlands, etc.) that serve as important spawning, nursery, foraging, and refuge areas. Included here are numerous species of recreational and commercial importance that occupy estuaries during their lifetime and support vital fisheries. Population growth, development, and anthropogenic activities in coastal watersheds surrounding estuaries, as well as impacts in the waterbodies themselves and those originating in ocean waters, are responsible for substantial ecosystem degradation (Kennish 2002, 2016, 2019; Day et al. 2012; Robb 2014). Thus, there are few estuaries exhibiting little or no stresses due to anthropogenic activities. For example, Merrifield et al. (2011) reported that only 16% of the estuaries along the west coast of the USA exhibited minimal or no stresses from anthropogenic activities. They also noted that some estuaries along the west coast have lost 90% of their bordering marshes and 99% of their estuarine shellfish species as coastal land use/land cover changes have increased over time. By comparison, they indicated that 500 MW). The calefaction of receiving waters stresses organisms and results in behavioral modifications, physiological dysfunction, disease promotion, and death (Kennish 1992). The heated water reduces the metabolic rate of exposed fauna, and it also interferes with enzyme activity, feeding, reproduction, respiration, and photosynthesis, often leading to changes in species composition, diversity, and density of biotic communities. The effect is most evident in outfall canals and other near-field regions of power plant sites where the highest thermal discharge temperatures occur. The most hazardous conditions arise from excessively high or rapidly fluctuating water temperatures during power plant operations culminating in heat- and cold-shock mortality of

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FIGURE 5.3  Map of the Barnegat Bay–Little Egg Harbor estuarine system (New Jersey, USA) showing surrounding watershed areas. Nitrogen and carbon inputs from the coastal watershed, atmospheric deposition, and the nearshore ocean have resulted in a highly eutrophic estuary, particularly in the northern segment bordered by the most developed area of the watershed. (From Kennish, M. J., Haag, S. M., and Sakowicz, G. P., Seagrass decline in New Jersey coastal lagoons: a response to increasing eutrophication, in Coastal Lagoons: Critical Habitats of Environmental Change, Kennish, M. J. and Paerl, H. W., Eds., CRC Press/Taylor & Francis group, Boca Raton, Florida, 2010, 171. With permission.)

organisms, most notably fish populations attracted to thermal plumes. Organisms concentrating in power plant outfalls occasionally experience mass thermal mortality due to the failure of smooth muscle peristalsis, denaturation of proteins in the cells, increased lactic acid in the blood, and oxygen deficit related to the increased respiratory activity (Kennish 1992, 2019). The release of heated effluent from electric generating stations and other industrial facilities can also deplete dissolved oxygen levels since warmer waters retain less dissolved oxygen than colder waters. Aside from increased mortality due to reduced dissolved oxygen concentrations, heat-shock and cold-shock mortality at electric generating stations has been responsible for the mass mortality of finfish and invertebrate populations that cannot adapt to the

rapid changes in water temperature in outfalls associated with abrupt changes in power plant operations. Organisms exposed to power plant thermal discharges may exhibit attraction or avoidance reaction to the heated effluent, and their altered behavior can displace them from their preferred natural habitat. Organisms inhabiting waters in close proximity to power plants may also experience higher mortality due to biocidal releases as well as impingement and entrainment effects that add to thermal impacts. Biocides are used to control the biofouling of power plant condensers to maintain operating efficiencies (Taylor 2006). However, biocides such as chlorine released to outfall canals have caused significant mortality of non-target organisms, especially phytoplankton and zooplankton (Kennish 1992). Impingement and

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entrainment account for the loss of millions of estuarine and marine organisms at larger power plants. Impingement mortality occurs when intake water flow is used to cool power plant condensers, and it traps fish and invertebrates on intake screens fronting condenser systems. Unable to free themselves from the screens due to the pressure of the water flow, many of the organisms drown or die from other effects. Entrainment occurs when plankton, microinvertebrates, and small juvenile fish are drawn through intake screens into plant condenser systems, entrained with the cooling water, and are then subjected to thermal stresses, hydraulic shocks, pressure changes, biocides, and mechanical damage during the in-plant passage. Eggs and larvae of fish and invertebrates are particularly susceptible. Most of these organisms are killed by the stresses, often exceeding a billion individuals annually at large power plants. Efforts to mitigate environmental impacts at coastal power plants usually entail the application of the best available technology such as modifying, retrofitting, and upgrading intake and cooling water system components to reduce impacts to organisms from the in-plant passage and detrimental effects in receiving waters.

5.2.4 Sewage Wastes and Pathogens Many estuarine waters worldwide receive treated or untreated sewage wastes that impact organisms and habitats. Dissolved and particulate organic carbon levels in estuarine and nearshore ocean waters receiving large amounts of sewage wastes commonly exceed 100 mg l–1 compared to organic carbon levels of ~1–5 mg l–1 in waters not affected by sewage wastes. Sewage waste loading to bottom sediments can dramatically alter the structure of benthic communities as bacterial abundance and respiration increase, dissolved oxygen levels decline, and anaerobic zones expand in the sediment column (Robb 2014; Kennish 2016, 2019). In addition, shellfish beds are typically closed to harvesting in these areas due to water quality degradation and the occurrence of pathogenic microorganisms (bacteria, viruses, protozoa, and helminths). Sewage wastes entering estuarine waters also pose a threat to human health by causing acute disorders such as cholera, dysentery, hepatitis, and typhoid. Water quality monitoring programs are conducted in estuaries because of the potential threat to coastal environments and human health from inputs of stormwater runoff, industrial and municipal wastewater, and sewage contamination. A component of these monitoring programs involves measurements of Enterococci sp. and fecal coliform bacteria levels in shellfish growing waters and at bathing beaches.

5.2.5 Chemical Contaminants Estuarine and coastal marine environments have been repositories for numerous toxins, most originating from land-based anthropogenic sources. Over the past half-century, for example, more than 70,000 synthetic chemicals have been found in these environments, with many toxic

to marine life (Kennish 1992, 1997, 2019). Halogenated hydrocarbons, petroleum hydrocarbons, polycyclic aromatic hydrocarbons (PAHs), heavy metals, radioactive substances, and volatile organics are some groups of chemical contaminants detrimental to marine life and habitats. Serious adverse effects of chemical contaminants, such as petroleum hydrocarbons, have been reported in estuarine and marine organisms (Table 5.4). The main pathways of their entry into estuarine and coastal marine waters include land runoff, groundwater, municipal and industrial wastewaters, sewage wastes, atmospheric deposition, fixed installations (coastal refineries, oil drilling platforms), marinas, tanker spills, shipping, and trash disposal. Less is known about the toxic effects of emerging chemical compounds accumulating in these environments such as pharmaceuticals, cleaning agents, and hygiene products (Frid and Caswell 2017). Chemical contaminants occur in the water column, bottom sediments, and biota of estuarine and coastal marine environments. Bottom sediments are sinks for the accumulation of these substances, often sorbed to sediment particles. At high concentrations in the water column and bottom sediments, these toxins are hazardous to biotic communities. Pathological responses of estuarine and marine organisms to toxic levels of these contaminants may include neurological, digestive, reproductive, and respiratory disorders, tissue inflammation and degeneration, neoplasia, and developmental abnormalities. Aberrant feeding behavior and growth inhibition also commonly occur (Kennish 1992, 1997, 2019). A major concern is a potential for some chemical contaminants to bioaccumulate or biomagnify in food chains. For example, heavy metals tend to bioaccumulate in estuarine and marine organisms, and some are very toxic (Figure 5.4). Methylmercury is a bioaccumulative organometallic toxicant that biomagnifies in estuarine and marine food chains from bacteria to top predators (fish and mammals) (Figure 5.5). Humans accumulate methylmercury by consuming contaminated seafood products, and it can be fatal. Methylmercury was responsible for the deadly Minimata disease disaster in Japan in the 1950s and 1960s, when nearly 2000 people died after consuming mercurycontaminated shellfish and other seafood products, the mercury derived from a petrochemical plant. Arsenic, lead, cadmium, mercury, and chromium are other particularly toxic metals posing a potential threat to estuarine and coastal marine organisms as well as humans. Thus, the concentrations of metals are often monitored in recreationally and commercially caught fish and shellfish in estuarine and coastal marine waters.

5.2.6 Human-Altered Hydrological Regimes Significant shifts in estuarine salinity, temperature, nutrients, turbidity, and circulation accompany human-altered hydrological regimes in many coastal watersheds and upland areas, which act as major drivers of change in

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TABLE 5.4 Commonly Reported Effects of Petroleum and Individual PAHs on Living Organisms Effects

Plant

Invertebrate

Fish

Reptile and Amphibian

Bird

Mammal

X X X

X X X

X X X

X

Individual Organisms Death Impaired reproduction Reduced growth and development Impaired immune system Altered endocrine function Altered rate of photosynthesis Malformations Tumors and lesions Cancer Altered behavior Blood disorders Liver and kidney disorders Hypothermia Inflammation of epithelial tissue Altered respiration or heart rate Impaired salt gland function Gill hyperplasia Fin erosion

X X X

Local population changes Altered community structure

X X

Biomass change

X

X X X

X X

X

X X X X

X

X X X Groups of Organismsa,b X X

X X X X X X

X X

X X X X X

X X X X X

X X X X X X X

X

X X

X

Source: Albers (1994). With permission. a Some effects have been observed in the wild and in the laboratory, whereas others have only been induced in laboratory experiments. b Populations of chlorophyllous (microalgae) and non-chlorophyllous plants (bacteria, filamentous fungi, yeast) can increase or decrease in the presence of petroleum, whereas animal populations decrease.

ecosystem structure and function. Marked changes in species composition, abundance, and distribution of estuarine and marine organisms typically occur in response to these shifting estuarine physical–chemical factors. Biotic productivity varies as well in response to the flux of nutrient inputs. Human population growth and development in coastal regions are placing greater demands on limited freshwater resources to meet domestic, industrial, and agricultural needs. Freshwater is important for electric power generation and maintaining infrastructure. As such, the increasing human use of freshwater is not sustainable in many coastal regions. There is excessive groundwater pumping locally that contributes to reduced freshwater base flow to the coast, saltwater intrusion, and, in some cases, land subsidence. To alleviate the strain on freshwater resources, hydrological systems are undergoing engineered modifications, such as across extensive areas of California (USA). More than 50% of the freshwater flow into San Francisco Bay, for example, has been diverted for human use, resulting in acute changes in the structure and function of biotic communities in the bay, including decreased phytoplankton biomass and zooplankton density, altered distribution and

diversity of benthic fauna, diminished reproductive success and variable abundance of some fish populations, and a depressed pelagic food web (Kennish 2000). Inland of estuaries, dams and water storage reservoirs constructed along rivers for flood control and to increase water supplies for domestic consumption, irrigation, livestock, and energy production have significantly reduced freshwater discharge to coastal systems. They have also caused river-bed and stream-bed modifications. There are an estimated 800,000 dams constructed worldwide with many significantly reducing water inflow in impounded estuarine tributaries. The development and urbanization of coastal watersheds, in contrast, results in greater impervious land cover with the removal of soils and the loss of natural vegetation that reduce infiltration, accelerate freshwater runoff, and increase stream and river discharges to estuaries. Upland deforestation, soil compaction, and wetlands reclamation worsen these effects and contribute to the loss of natural water storage capacity and greater downstream flows. Increasing precipitation in many regions predicted with accelerated climate change during the 21st century will exacerbate these effects. Greater freshwater inflow

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The residence time of estuarine waterbodies can also change substantially with the diversion of freshwater flows and other hydrological modifications. This change can have a great impact on the composition of biotic communities and the distribution of organisms from the head to the mouth of estuaries. Fisheries production may change greatly as well (Kennish et al. 2008).

5.2.7 Human-Induced Sediment/Particulate Inputs

FIGURE 5.4  Metal levels in recreationally and commercially caught marine fish and shellfish. Shown are means (vertical line), standard error (box), and range (horizontal line). (From Burger and Gochfeld (2016). With permission.)

decreases the water residence time in estuaries, while increasing the capacity to dilute, transform, and flush contaminants (Kennish 2000). Other anthropogenic modifications that can significantly alter water-flow regimes in estuaries and along coasts include channelization for flood control to protect infrastructure and marsh impoundment, which also affect natural water storage capacity and the filtering of pollutants (Kennish et al. 2008). Freshwater diversions have ecological applications as well, such as for coastal wetlands restoration. For example, river water has been diverted into Louisiana coastal wetlands to reduce the loss of salt marshes by reversing or slowing the rates of habitat degradation (Teal et al. 2012). However, secondary effects of wetlands restoration occur in estuarine waterbodies. For example, freshwater diversions into coastal salt marshes such as those in Louisiana can cause significant changes in estuarine water quality, salinity, nutrient availability for primary production, and sediment distribution.

Sediment loads of major rivers of the world are considerable (Table 5.5). Silviculture operations and the development of coastal watersheds disrupt natural landscapes, remove soils and vegetation, and increase erosion and sediment loads to estuaries and coastal marine waters. Timbering and logging in some countries of Southeast Asia and South America provide examples. Farming and other agricultural activity deliver high concentrations of sediments to some coastal ecosystems as well, such as along the Mississippi River where sediment inputs have also caused significant water quality problems (Mikhailov and Mikhailov 2010; Merten et al. 2016). Pollutants entrained in runoff and sorbed to sediments adversely affect water and sediment quality in estuarine and coastal marine environments. Water column turbidity increases the attenuation of light and shading of benthic algae and seagrass beds, which serve as important habitats for benthic communities and demersal fish. Moore et al. (2012, 2014) attributed the dieback of seagrasses over extensive areas of the Chesapeake Bay ecosystem to elevated turbidity in the water column. High sediment delivery and deposition during major coastal storms can smother shellfish beds and degrade benthic communities. Stormwater and sediment management strategies exist to reduce runoff and transport of sediments, nutrients, and chemical pollutants in coastal watersheds. Implementation of best management practices includes regulatory erosion controls and sediment load limits (e.g., at farmlands, construction sites, and urban centers), use of structural controls (e.g., constructed wetlands, detention basins, infiltration facilities, and water quality inlets), and restoration of protective habitat (e.g., stream and river banks). Locally, other engineered structures are effective in mitigating sediment inputs to receiving waters, such as sediment traps, filter fences, and check dams. Since storms can cause significant pulses of precipitation that erode and transport sediments to streams, rivers, and coastal systems, regional stormwater management programs are often in place to abate flooding and its effects (e.g., flood control networks) (Kennish 2019).

5.2.8 Introduced/Invasive Species The introduction or invasion of species not endemic to estuarine ecosystems can cause considerable ecological disruption because they often have no natural controls in their adopted estuaries, and thus can alter biotic community structure by outcompeting native species for food sources and space, rapidly dominating floral and faunal

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FIGURE 5.5  Schematic food web diagram illustrating the bioamplification of methylmercury from sediment to methylating bacteria to plankton, to small, medium, and large fish and their predators. (From Burger, J. and Gochfeld, M., Habitat, Population, Dynamics, and Metal Levels in Colonial Waterbirds: A Food Chain Approach, CRC Press/Taylor & Francis, Boca Raton, Florida, 2016, 257. With permission.)

TABLE 5.5 Runoff and Sediment Load in Major Rivers of the World Drainage Area (106 km2)

Runoff (km3/year)

Sediment Load (106 tons/year)

Amazon

6.2

6300

900

Danube Ganges Irrawaddy Magdalena Mekong Mississippi Niger Ob Orinoco Yangtze Yellow Yenisei

0.8 1.5 0.4 0.2 0.8 3.3 1.2 2.5 1.0 1.9 0.8 2.58

206 971 428 237 470 580 192 385 1100 900 49 560

67 1670 265 220 160 210 40 16 210 478 1080 13

Zaire

3.8

1250

43

River

Source: Lal (2000). With permission.

communities, and reducing species diversity (Ruiz and Carton 2003). In addition, they commonly alter habitats. Predation and competition by introduced and invasive species also frequently displace native species, resulting in changes in the trophic organization of estuaries (Burros 2016; Weis 2016). Marked changes in species composition and distribution typically occur in the invaded waters (Burros 2016; Weis 2016).

San Francisco Bay (USA) is an example of an estuarine ecosystem significantly altered by introduced and invasive species. For example, the Asian clam (Potamocorbula amurensis) introduced in Suisun Bay (USA) outcompeted native shellfish species (Macoma balthica and Mya arenaria) and concurrently decimated the phytoplankton community. More than 250 non-indigenous species now inhabit the San Francisco Bay ecosystem, and the dominant species in the bay are mainly introduced forms (Kennish 2000, 2019). Most of the benthic organisms in the inner shallows of the bay are non-indigenous species, as are the fish species in the delta region (Kennish 2017). Some species (e.g., striped bass, Morone saxatilis) have been introduced for commercial or recreational purposes. Other introduced species (e.g., Spartina alterniflora) now have an expanded distribution, significantly altering the marsh habitat surrounding the bay (Weis 2016). There are four main mechanisms of species introductions or invasions in an estuary: (1) deliberate introductions; (2) accidental introductions alongside the deliberate movement of species; (3) shipping-related relocations; and (4) Lessepsian migration (i.e., introductions through the Suez Canal and other artificial, human-constructed waterways) (Frid and Caswell 2017). A major vector for species invasions in many estuaries is vessel-mediated, with exotic species and other forms entrained in ship-ballast water tanks being the sources of entry (Bax et al. 2003). As shipping activity increases worldwide, more species invasions are likely to occur. In addition, more mariculture activity in coastal waters worldwide is introducing new species in some waterbodies. Moreover, increasing numbers

Anthropogenic Drivers of Estuarine Change

of invasive marine species are evident worldwide due to climate change, and these species typically have greater tolerance to a broader thermal range than native species (Canning et al. 2011). Introduced and invasive species can pose an economic threat to the viability of recreational and commercial fisheries. For example, the parasitic protozoan Haplosporidium nelsoni, which entered the Delaware Estuary (USA) in the 1950s presumably as spores transferred from the West Coast of the USA or from ballast water of Asian ships, caused extensive mortality of the American oyster (Crassostrea virginica) population. Its infestation severely decimated the population and had catastrophic consequences for decades on the multimillion-dollar commercial oyster shellfishery (Kennish 2019).

5.2.9 Overfishing and Intensive Aquaculture As the global human population has grown to more than 8 billion (in January 2023), the demand on fisheries harvests is peaking. The United Nations Food and Agricultural Organization (FAO) has estimated that nearly 70% of the world’s fisheries are overfished, depleted, or recovering from overfishing; in 2018, global fish production was estimated to be 179 million mt, with aquaculture accounting for 46% of the total (FAO 2020). The maximum fishery biomass that can be removed annually while maintaining a stable standing stock (i.e., the maximum sustainable yield) is estimated to be 100–120 million mt/year (Sverdrup and Kudela 2014). It is evident, then, that the global fisheries yield is approaching the maximum sustainable yield, meaning that the world’s fish stocks are being fully exploited or overexploited. Overfishing is a serious concern because it leads to the collapse of fisheries (Le Pape et al. 2017). However, overfishing occurs in some nations but not others (Sissenwine et al. 2014; Colloca et al. 2017). For example, overfishing has significantly depleted fisheries in Mediterranean waters, but has been avoided in Iceland, Norway, the Philippines, and elsewhere (Colloca et al. 2017). The implementation of effective fishery management frameworks has been a primary reason why nations have averted the collapse of fisheries (Sissenwine et al. 2014). Overfishing and overharvesting of target species not only threaten sustainable seafood resources, but also cause ecological impacts as well. Excessive fishing pressure on top predators, such as bluefish, weakfish, and striped bass, can create imbalances in biotic community structure and function, leading to reductions in biodiversity and shifting trophic interactions. Depleted predatory species results in greater fishing pressure on lower-trophic-level target species to increase catch sizes, a practice termed “fishing down the food web” (Pauly et al. 1998). The net effect is lower catches of large piscivorous fishes and greater catches of omnivorous and planktivorous forms. These changes in the abundance of various fish species have effects on other biotic groups which depend on fish for nutrition and

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survival (e.g., seabirds and shorebirds, marine turtles, dolphins, seals, and whales). Recreational and commercial fisheries in estuaries have been exceptionally productive and an important source of seafood protein for people in the world, although there is increasing concern about them as well. For example, a significant decline in San Francisco Bay fisheries (i.e., chinook salmon, Oncorhynchus tshawytscha; delta smelt, Hypomesus transpacificus; and striped bass, Morone saxatilis) has occurred over the past 50 years due to overfishing. Similar fisheries decline due to overfishing have been reported in the past for Albemarle–Pamlico Sounds, North Carolina, and Sarasota Bay, Florida (USA) (Kennish 2000). Overharvesting, together with disease and predation, has been responsible for the marked declines in hard clam and oyster populations in Mid-Atlantic estuaries (Kennish 2019). Science-based fishery management programs are necessary to maintain sustainable and viable recreational and commercial fisheries in estuarine and coastal marine waters. Aquaculture, which has been expanding in many countries around the world, has compensated in part for increasing overfished stocks. It now represents the main source of seafood for human consumption, exceeding the harvest of wild-capture fisheries (FAO 2016; Guillen, et al. 2019). Nations in Asia (China, Vietnam, the Philippines) are the primary producers of aquaculture products, notably fish, shellfish, and algae in coastal waters. China produces more than one-third of the global fish supply and accounts for more than 60% of the global aquaculture production (Cao et al. 2015; Wartenberg et al. 2017). While aquaculture produces large amounts of seafood products for human consumption and other uses, it also causes environmental disturbances that impact habitat and water quality in estuarine and coastal marine ecosystems. Intensive aquaculture can significantly degrade sediment quality, convert natural habitat, and act as a vector for diseases and parasite transmission via the accumulation of feces, pseudofeces, and products from feedlot operations (New 2002; Martinez-Porchas and Martinez-Cordova 2012). Organisms grown in net cages, pen cultures, on ropes, in mesh bags, or other suspended structures release nutrients, particulate organic carbon, and other products that promote eutrophication of surrounding estuarine and coastal marine waters, with algal blooms, increased biochemical oxygen demand, and decreased dissolved oxygen in bottom sediments and the water column detrimental to biotic communities. Antibiotics used in fish feed to treat or prevent diseases accumulate in bottom sediments with food products and can be problematic as well, causing bacterial resistance, food hazard concerns, and environmental degradation. Finally, the use of non-indigenous exotic species in aquaculture has the potential for increasing introduced/ invasive species that are potentially damaging to their adopted waters. In summary, aquaculture impacts in estuarine and coastal marine environments can be grouped into several

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categories, namely chemical impacts (nutrient inputs, toxic compounds, pharmaceuticals, and metals), physical impacts (habitat modification and alteration to hydrodynamics), ecological impacts (algal blooms, pathogens, and invasive species), and trophic-level impacts (algae, shellfish, and fish impacts). All of these impacts require additional study. In addition, various management efforts are necessary to mitigate adverse effects, such as the enforcement of appropriate stocking densities and improvement in policies and regulations. The implementation of integrated multi-trophic aquaculture will provide biomitigative services to the ecosystem and improve economic farm output as well (Chopin et al. 2012).

5.2.10 Coastal Subsidence Subsiding coastal areas increase relative sea-level rise, inundation, shoreline retreat, and impacts on fringing wetlands and other coastal watershed habitats. Relative sea-level change is defined as the sum of geocentric sea-level change plus vertical land motion (Shirzaei et al. 2021). Several natural and anthropogenic factors drive coastal subsidence via flux in vertical land motion. Natural processes that drive vertical land motion include glacial isostatic adjustment (GIA), tectonics and earthquakes, sediment compaction due to sediment loading, and sinkhole formation by karst processes. Anthropogenic factors driving vertical land motion include compaction of aquifer systems and hydrocarbon reservoirs accompanying withdrawal of water, oil, and gas, as well as peat oxidation following drainage (Shirzaei et al. 2021). Coastal subsidence due to anthropogenic fluid withdrawal can be rapid (~30 cm yr–1); where subsurface fluid extraction is heavy, anthropogenic subsidence can be the main contributor to relative sea-level rise in coastal environments (Ingebritsen and Gallaway 2014). Subsiding coastal areas have an average relative sea-level rise of up to 7.8 to 9.9 mm yr–1, which is four times faster than the global-mean relative sea-level rise of 2.6 mm yr–1 (Nicholls et al. 2021). Subsiding coastal areas pose a threat to coastal communities due to the increasing submergence of protective wetlands and other low-lying habitats and exposure of coastal-built communities to serious flooding. In addition, estuarine basin morphometry and bathymetry within subsiding coastal areas can change considerably together with circulation, salinity, and other physical–chemical factors that affect biotic communities in the waterbody. Coastal subsidence due to oil and gas extraction occurs in the area of Galveston Bay, Texas (USA) (Shipley and Kiesling 1994). Excessive groundwater withdrawal is responsible for coastal subsidence in the Chesapeake Bay area (USA); Boon et al. (2010) reported subsidence rates of –1.3 to –4.0 mm yr–1. Similar drivers of change are evident in other countries as well. For example, excessive groundwater pumping has resulted in significant coastal subsidence in the Mekong Delta, Vietnam and Cambodia (Ingebritsen and Gallaway 2014). Substantial coastal subsidence has also been documented in the Po Delta, Italy,

Climate Change and Estuaries

and Tokyo, Japan (Kennish et al. 2008). Anthropogenic coastal subsidence and consequent coastal flood exposure can be mitigated with appropriate science-based environmental policies that effectively limit groundwater pumping and hydrocarbon extraction. These management efforts are necessary because land subsidence and sea-level rise will continue to pose threats to coastal communities and ecosystems with ongoing climate change through the 21st century and beyond.

5.2.11 Floatables/Plastics/Debris The United Nations Environment Programme (UNEP) has defined marine litter as “any persistent, manufactured or processed solid material discarded, disposed of, or abandoned in the marine and coastal environment.” The International Coastal Clean-up of UNEP collected 1.02 million kg of marine litter from 20,183 km of coastline in 2005–2007, and 18.06 million kg of marine litter from 25,189 km of coastline in 2015 (Frid and Caswell 2017). A wide range of materials comprised the marine litter collected, including packaging, glass and plastic bottles, bottle caps, plastic and metal containers, paper bags, cardboard, fishing line, and other items. Most debris found in estuarine and marine environments (~80%) derives from land-based sources (e.g., coastal urban centers, towns, landfills, beaches, etc.), with maritime sources of debris (e.g., commercial and recreational vessels, military ships, and offshore drilling platforms) accounting for most of the remainder (Garrison 2007). Most estuarine and marine debris consists of plastics – synthetic organic polymers that constitute globally ubiquitous and persistent pollutants. There are four size classes of plastic debris in these environments: microplastics (20 mm), and megaplastics (>100 mm). Microplastics are extremely abundant in coastal and marine ecosystems, ranging from 0.001–140 particles m–3 in the water column and 0.2–8,766 particles m–3 in bottom sediments (Thushari and Senevirathna 2020). Estuaries represent one of the most heavily impacted coastal ecosystems with regard to plastic pollution (Li et al. 2016). Frid and Caswell (2017) report that plastics account for 70% of the litter on shorelines, 68–99% of the litter in the water column, and 23–89% of the litter on the seafloor of the continental shelf. Critchell et al. (2019) note that plastics comprise up to 80% of all debris in the world’s oceans. An estimated 8 million mt yr–1 of plastic waste enters the oceans annually, and this waste is increasing at a rate of 7% a year (Jambeck et al. 2015). An estimated 5.25 trillion plastic particles occur in the world’s oceans (Erickson et al. 2014). Plastics are particularly troublesome because they resist degradation and can persist in estuarine and marine environments for centuries, posing a long-term threat to biotic communities (Li et al. 2016). Plastics are hazardous to marine life, causing debilitation and death of organisms via ingestion, entanglement, and

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suffocation (Barbosa et al. 2019). The threat is particularly problematic for birds, turtles, fish, and mammals, which often ingest the materials or become entangled in nets, fishing lines, and packing bands. When ingested by organisms, the larger plastics can obstruct airways and digestive systems frequently causing starvation, intestinal blockage, and death. Microplastics easily consumed by marine organisms can bioaccumulate and biomagnify in marine food webs. In addition, marine plastic debris is also a vector for the transport of toxic chemicals such as persistent organic pollutants (e.g., PAHs, PCBs, DDT, dioxins, furans) that are hazardous to estuarine and marine organisms (Li et al. 2016). Flame retardants and phthalate plasticizers associated with plastics can also mobilize through food webs to humans. Further, plastics can act as platforms in the sea that transport invasive species and other debris to new environments where ecosystem function and habitats are impacted. Microplastics have become a primary target of environmental remediation programs because they are components in frequently used commercial products that often pollute estuarine and coastal marine ecosystems. Wastewaters containing cosmetics and hygiene products (e.g., facial cleaners and toothpaste), medicines, and textiles (polyester or nylon fibers) as well as breakdown substances of larger plastic wastes deliver very large amounts of microplastics to coastal waters. There is now an international effort underway to reduce the input of microplastics as well as larger plastics in marine environments. GESAMP, the Joint Group of Experts on the Scientific Aspects of Marine Pollution, is a major force in this regard. A holistic management approach is necessary to control and prevent the accumulation of plastic debris in estuarine and marine environments, including regional, national, and global institutional involvement. At the global scale, actions of the United Nations Food and Agricultural Organization, United Nations Environment Programme, and International Maritime Organization are essential for addressing debris/litter in coastal and open ocean waters (Thushari and Senevirathna 2020). Academic institutions, government agencies, non-governmental organizations (NGOs), and conservation groups are monitoring plastics and other debris in these waters as part of assessment and remediation programs. Studies are focusing on the sources, transport pathways, inputs, and the fate of plastics and other debris in these environments (Li et al. 2016; Thushari and Senevirathna 2020). Approaches to reduce plastic pollution include government bans on the use of single-use plastic bags, levies on their use, and the introduction of non-polluting reusable cloth bags and other non-polluting products at stores and other venues. The UK banned the use of microbeads in rinse-off cosmetics and personal care products in 2017. Canada and the USA are the only other countries to implement bans on microbeads, although many additional countries are considering the phasing out of microbeads in these products (Xanthos and Walker 2017). Market-based strategies and policies are effective control measures to reduce single-use

plastics at their source. Education and community outreach programs are also part of overall management efforts to mitigate plastic and other debris pollution in estuarine and marine environments (Chang 2015; Xanthos and Walker 2017). These programs stress the reuse, recycling, and reduction of plastics.

5.2.12 Climate Change The earth’s climate is a major factor modulating the dynamics of biotic communities in estuaries. Climate change is defined as regional or global changes in mean climate state or in patterns of climate variability over decades to millions of years often identified using statistical methods, and sometimes referred to as changes in long-term weather conditions (Cronin 2016; IPCC 2021). As noted by Easterling and Kunkel (see Chapter 2, this volume), the earth’s climate has changed considerably over the past 150 years, mainly due to increasing emissions of greenhouse gases and land use cover changes driven by anthropogenic activities. The mean global temperature over this period has increased more than 1°C, with the most significant increases occurring since the 1970s (Figure 5.6). Mean sea surface temperatures recorded for more than a century have been the highest during the past three decades, consistent with the highest air temperatures recorded as well for this period. Similarly, estuarine and coastal marine waters have exhibited increasing temperatures over the past several decades. The rate of increase in sea surface temperatures along coastlines has exceeded that of the open ocean, amounting to 0.18 ± 0.16°C per decade along more than 70% of the world’s coastlines (Lima and Wethey 2012; Wong et al. 2014). Marine heatwaves are increasing in frequency, creating stressful conditions for biotic communities (IPCC 2021). Climatic forcing has profound effects on physical–chemical processes, influencing estuarine stratification, mixing, circulation, and biogeochemistry that modulate ecosystem structure and function. Temperature and salinity regimes in estuaries are shifting with climate change. Warming estuarine and coastal marine waters, rising sea-levels, increasing storm intensity, variable precipitation, and altered hydrology affect organism growth, abundance, reproduction, distribution, diversity, phenology, and production, as well as species interactions and food web dynamics. Warming waters promote the stratification of estuaries and the increased occurrence of pathogenic bacteria, such as Vibrio species (Bindoff et al. 2019). Stronger stratification of the water column also reduces vertical mixing increasing the risk of hypoxia in deeper waters and higher mortality of benthic organisms. Estuarine biotic communities are changing with climateinduced temperature increases and other environmental forcing factors (Robins et al. 2016; Frid and Caswell 2017; Bindoff et al. 2019; Kennish 2019, 2021). Climatic and non-climatic drivers of change that interact synergistically can stress biotic communities (see Chapter 1, this volume).

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FIGURE 5.6  Global temperature time series showing the annual departure of temperatures from the 20th-century average. Note increasing temperatures since the 1970s. (From NOAA/NASA. http://www​.ncei​.noaa​.gov/)

Shifting geographic distribution and altered physiological tolerances of organisms to climate change are causing the significant decline of recreational and commercial fisheries in some regions. Poleward range shifts of organisms are evident between estuaries in response to the warming waters (Hallett et al. 2018). Temperature is the main limiting range factor in seagrass meadows, and mangroves are encroaching into subtropical salt marshes in response to rising temperature (Hoegh-Gulberg et al. 2018; Bindoff et al. 2019). Greater precipitation is increasing land runoff and nutrient inputs to many estuaries, driving greater primary production and their susceptibility to eutrophication and hypoxia (see Chapter 17, this volume). Biotic invasions and extirpation of species are increasing as well driven by regional and local changes in physical–chemical conditions (Frid and Caswell 2017; Bindoff et al. 2019). Global mean sea-level rise (GMSL) increased by 0.20 (0.15–0.25) m between 1901 and 2018 (Fox-Kemper et al. 2021), and the rate of sea-level rise is increasing (Figure 5.7). For example, Easterling and Kunkel (see Chapter 2, this volume) reported that the average rate of sea-level rise during the period 1880–1992 was 1.5 cm decade –1, but since 1993, it has increased to 2.9 cm decade –1. GMSL is projected to increase further through the 21st century and beyond (IPCC 2021). The projected sea-level rise for 2050 based on physical modeling and statistical techniques range up to 0.5 m. The future projections of sea-level rise will reduce the resilience and sustainability of many coastal wetlands and other essential estuarine and coastal marine habitats that provide a protective buffer for developed communities in

coastal watersheds from extreme weather events, inundation, and flooding (Bianchi et al. 2019; Bindoff et al. 2019; Kennish 2019; IPCC 2021). Rising sea-levels cause marine transgression and, together with higher intensity coastal storms, accelerate erosion, and lead to the loss of wetlands habitat as estuarine shorelines retreat, eliminating essential habitat for many recreational and commercial species, and altering the morphology of estuarine basins as they widen and deepen. This process also eliminates vital habitat for other estuarine and marine organisms of ecological importance that utilize wetlands habitat for spawning, nursery, foraging, and refuge areas. Management strategies to address climate change impacts on coastal environments include protective legislative actions and policies to mitigate climate change drivers, science-based management plans to increase ecosystem sustainability, and integration of national, regional, and international program concepts (Kennish 2022). Education programs designed to inform and instruct the public about the causes and consequences of anthropogenic impacts on these environments and the remediation steps necessary to improve environmental conditions are critical as well. A major goal is to increase the resilience of natural and built communities to rising sea-levels and coastal storms. In addition, improved monitoring, modeling, and adaptive management schemes are needed. Further, a concerted effort is necessary from multiple sectors of society to implement the adaptation and resilience measures necessary to improve conditions for natural and built communities impacted by climate change.

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Anthropogenic Drivers of Estuarine Change

FIGURE 5.7  Global sea-level measurements from satellite altimeter with sea-level rise scenarios provided for the 2014 National Climate Assessment. Data from satellite altimeters show that global sea-level has been rising at about 3 mm/yr over the last several decades. Global sea-level rise is currently tracking between the Intermediate High and Intermediate Low Scenarios. (From NOAA, Tides and Currents. https://tidesandcurrents​.noaa​.gov​/slregional/.)

5.3 ENVIRONMENTAL IMPACT FACTORS There are multiple anthropogenic drivers of change that impact estuarine ecosystems largely due to the high human population growth, development, and altered landscapes in coastal watersheds. Land-based sources of pollution and other land-based anthropogenic drivers of change, together with overfishing, intensive aquaculture, dredging, and in-basin habitat impacts, are often detrimental to biotic communities. These drivers of change may interact synergistically, antagonistically, or additively to cause significant ecological change, although the ecological outcomes are often complex and difficult to delineate because of the combined impacts of multiple drivers that can vary considerably with the type of factors involved (Kennish 2021). The combined effect can be greater or less than expected (i.e., synergistic or antagonistic), rather than additive or multiplicative (Gunderson et al. 2016; Kroeker et al. 2017). When synergistic, the effect of a combination of drivers is more severe than the sum of each individual driver; when antagonistic, the effect of a combination of stressors is less severe than the sum of each individual driver; when additive, the effect of a combination of drivers is equal to the effect of each individual driver (see Chapter 23, this volume). The ecological impact of one driver frequently depends on the magnitude of another. The interaction of multiple climate change drivers and local direct anthropogenic drivers, therefore, can be particularly complex, accounting for significant non-linear effects of cumulative environmental impacts on biotic

communities and ecosystem function (Griffen et al. 2016; Gunderson et al. 2016). Measuring the effects of multiple interactive environmental drivers in highly variable estuarine and coastal marine waters remains a challenge in part because of confounding factors. A more accurate quantitative analysis of climate change impacts on estuarine ecosystems is necessary to improve understanding of the effects of interactive factors occurring in these important coastal environments and to develop effective management strategies to remediate adverse conditions. As noted by Robins et al. (2016), many climate impact predictions lack resolution, and more research is needed along the watershed–river–estuary–coastal ocean continuum to understand the impacts and to formulate mitigatory strategies. To document multiple driver impacts and to predict the combined impacts of the drivers under variable conditions, it is important to determine the frequency and intensity of the drivers as well as their spatial and temporal patterns of occurrence. Predicting the ecological outcomes of multiple driver interactions would greatly aid conservation programs in protecting estuarine biotic communities and habitats, as well as advancing estuarine sustainability.

5.4 IMPACT REMEDIATION Remedial measures are necessary to address the multitude of climatic and non-climatic drivers of change impacting estuaries. The long-term viability and sustainability of estuarine environments depend on the application of sound

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management strategies to control the effects of anthropogenic activities. Several approaches are proposed, including the following: (1) the enactment, coordination, and enforcement of land-use plans by municipal, county, and state governments that promote smart development; (2) the promotion and encouragement of municipal, county, state, and federal governments to control and limit non-point source pollution runoff; (3) the agreement of land owners, farmers, and municipal governments to minimize inputs of pesticides, herbicides, and fertilizers; (4) the participation of the scientific community to demonstrate in an understandable way the importance of coastal wetlands habitat to fisheries and other services; and (5) the education of the public that everyone contributes to the contamination problems affecting the environment and that each person is environmentally responsible (Hoss and Engel 1996; Kennish 2019). Greater pollution controls are needed in coastal watersheds (Table 5.6). The main goal of estuarine environmental management is to maintain and protect the ecological structure and functioning of ecosystems while concurrently ensuring the maintenance of ecosystem services beneficial to society (Elliott 2011, 2013; Elliott et al. 2022; Kennish 2022). To effectively assess the causes, consequences, and responses of change in estuarine and coastal marine environments impacted by anthropogenic activities, ecosystem-based management strategies should consider the application of a holistic problem-structuring framework, such as a Drivers– Activities–Pressures–State Change–Impacts–Responses approach (Elliott et al. 2017; Kennish 2022). An ecosystem-based management approach for estuarine and coastal marine environments is a holistic strategy that considers the entire ecosystem, including humans. The goal of this approach is to maintain coastal ecosystems in a healthy, productive, and resilient condition, while providing goods and services for human use (Elliott et al. 2017). Restoration is an integral part of conservation management that uses science-based techniques and methodologies to enhance the structure, function, and health of degraded estuarine and coastal marine ecosystems (Abelson et al. 2015, 2020; Elliott et al. 2016). A preferred management approach to effectively address interactive anthropogenic climatic and non-climatic driver impacts in estuaries is Integrated Coastal Zone Management (ICZM), a resource planning system which deals holistically with coastal problems employing ecological risk assessment and ecosystem-based management decision-making (Thia-Eng 1993). These are often placebased efforts, which include climate adaptation and protection strategies, such as establishing Marine Protected Areas (MPAs), installing living shorelines, and restoring impacted habitats (e.g., altered coastal wetlands) (McLeod and Leslie 2009). The management approach should integrate multiple system components across the watershed–river–estuary– coastal ocean continuum. An important goal is achieving estuarine sustainability based on ecosystem integrity and functioning (Day et al. 2012).

Climate Change and Estuaries

TABLE 5.6 Options for Treatment of Pollution in Coastal Watersheds Source Controls Animal waste removal Catch basin cleaning Cross-connection identification and removal Proper construction activities Reduced fertilizer, pesticide, and herbicide use Reduced roadway sanding and salting Solid waste management Street sweeping Toxic and hazardous pollution prevention Regulatory Controls Land acquisition Land-use regulations Protection of natural resources Structural Controls Constructed wetlands Detention facilities Extended detention dry ponds Wet ponds Filtration practices Filtration basins Sand filters Water quality inlets Infiltration Facilities Dry wells Infiltration basins Infiltration trenches Porous pavement Vegetative practices Grassed swales Filter strips Source: Bingham, D. R., Wetlands for stormwater management, in Applied Wetlands Science and Technology, Kent, D. M., ed., Lewis Publishers, Boca Raton, FL, 1994, 245. With permission.

Actions of national, regional, and international organizations are important as well to effectuate solutions to anthropogenic climatic and non-climatic driver impacts in the coastal zone. For example, the United Nations Convention on the Law of the Sea provides an international legal framework for controlling marine pollutants that often interact synergistically with climate change drivers in estuarine and coastal marine environments to create serious impacts on biotic communities and habitats. Other organizations dealing with global-scale issues of climate change and other environmental impacts include the United Nations Environment Programme (UNEP), United Nations Framework Convention on Climate Change, Joint Group of Experts on the Scientific Aspects of Marine Environmental Protection (GESAMP), International Maritime Organization, International Convention for the Prevention of Marine Pollution (MARPOL), International

Anthropogenic Drivers of Estuarine Change

Oceanographic Commission, and the Intergovernmental Panel on Climate Change. GESAMP, as an example, functions under the auspices of 10 United Nations organizations with a policy approach based on a scientific understanding of marine ecosystems and anthropogenic activities that affect them. The regular work of GESAMP revolves around five specific functions (www​.gesamp​.org​/work​/programme): • Integration and synthesis of results of regional and thematic assessments and scientific studies to support global assessments of the marine environment; • Provision of scientific and technical guidance on the design and execution of marine environmental assessments; • Provision of scientific reviews, analyses, and advice on specific topics relevant to the condition of the marine environment, its investigation, protection, and/or management; • Provision of an overview of the marine environmental monitoring, assessment, and related activities of UN agencies, and advising on how these activities might be improved and better integrated and coordinated; • Identification of new and emerging issues regarding the degradation of the marine environment that are relevant to governments and sponsoring organizations. There have been greater efforts over the past two decades to increase the coordination and integration of major national and international environmental programs to address broader anthropogenic climate change and other impacts. However, a comprehensive, unified global approach is necessary to successfully address climate change effects and other anthropogenic environmental impacts to maintain estuarine and marine environmental quality. While progress has been made on program coordination, the present international framework dealing with ocean pollution remains a patchwork of agreements, protocols, and regional and international conventions, as well as national laws, regulations, and guidelines.

5.5 CONCLUSIONS Anthropogenic activities in coastal watersheds and coastal waters affect the water quality, habitats, and biotic communities in estuaries. Land-derived pollution inputs, physical alteration of habitats, and overexploitation of resources cause significant impacts on estuaries that provide important ecosystem services, including significant societal goods and benefits. Rapid population growth and development in coastal watersheds are responsible for many of the persistent environmental problems affecting these waters. They are conspicuous at the land–sea margin where anthropogenic activities are most intense. For

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example, deforestation and construction, landscape partitioning and paving, impervious surfacing, marsh filling, diking, lagoon formation, dam installation, dredging and dredged-material disposal, mariculture, and hardening of shorelines not only damage habitats but also degrade biotic communities. Activities in ports, harbors, marinas, and coastal installations generally produce hot spots of contamination. The release of agricultural, industrial, and municipal wastes often compromises the ecological integrity of receiving waters. The most acute environmental degradation occurs in highly urbanized and industrialized estuaries. There are 12 types of notable anthropogenic drivers in estuaries, including: (1) habitat loss and alteration (e.g., lagoon construction, shoreline development and hardening, impervious land cover, and land reclamation); (2) dredging and dredged-material disposal; (3) enrichment (e.g., nutrients, organic matter, and thermal loading); (4) sewage wastes and pathogens; (5) chemical contaminants (toxins), also including emerging chemical compounds of concern (e.g., pharmaceuticals, cleaning agents, and hygiene products); (6) human-altered hydrological regimes; (7) human-induced sediment/particulate inputs; (8) introduced/invasive species; (9) overfishing and intensive aquaculture; (10) coastal subsidence; (11) floatables/plastics/ debris; and (12) climate change. They can be assembled into three broad categories: (1) those that impact habitat and are mainly driven by physical factors (e.g., coastal construction, shoreline hardening/development, dredging, and dredgedmaterial disposal); (2) those that affect water quality and are mainly chemical and biological in nature (e.g., nutrient enrichment, chemical contaminants, and pathogens); and (3) those that alter biotic communities and are driven by multiple stressors (e.g., human-altered hydrological regimes, introduced/invasive species, overfishing, and climate change and sea-level rise). The environmental condition of estuaries can be improved by formulating effective development plans that address population growth and land use in coastal watersheds. The most serious anthropogenic impacts in the marine realm occur in estuarine and coastal marine waters in close proximity to highly developed land areas, and anthropogenic activities in coastal watersheds drive a significant number of these impacts. Educational programs that inform school administrators, students, local communities, and political leaders regarding the detrimental effects of anthropogenic activities on coastal environments are vital for the conservation and sustainability of estuarine and coastal marine ecosystems. Management programs should target the restoration and protection of estuarine habitats impacted by anthropogenic activities. The restoration of damaged habitats to sustain ecosystems requires labor-intensive efforts and considerable funding. Restoration projects typically require many months or years to complete, but they are necessary to achieve habitat recovery on a broad scale while concurrently increasing ecosystem resilience.

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ACKNOWLEDGMENTS This is Publication Number 2468 of the Department of Marine and Coastal Sciences, Rutgers University, New Brunswick, New Jersey. Research components reported in this chapter were supported in part by grants from NOAA, the U.S. Environmental Protection Agency, the New Jersey Department of Environmental Protection, and the New Jersey Sea Grant Consortium.

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Climate Change and Estuaries McLusky, D. S. and M. Elliott 2004. The Estuarine Ecosystem: Ecology, Threats, and Management. Oxford: Oxford University Press. Merrifield, M. S., E. Hines, X. Liu, and M. W. Beck. 2011. Building regional threat-based networks for estuaries in the western United States. PLOS ONE 6(2): e17407. http://doi​ .org​/10​.1371​/journal​.pone​.0017407. Merten, C. H., H. L. Welch, and M. D. Tomerl. 2016. Effects of hydrology, watershed size, and agricultural practices on sediment yields in two river basins in Iowa and Mississippi. J. Soil Water Conserv. 71(3): 267–278. Mikhailov, V. N. and M. V. Mikhailov. 2010. Delta formation processes at the Mississippi River mouth. Water Resour. 37(5): 595–610. Mitsch, W. J. and S. E. Jørgensen 2004. Ecological Engineering and Ecosystem Restoration. New York: John Wiley and Sons. Mitsch, W. J. 2012. What is ecological engineering? Ecol. Eng. 45: 5–12. Moore, K. A., E. C. Shields, D. B. Parish, and R. J. Orth. 2012. Eelgrass survival in two contrasting systems: Role of turbidity and summer water temperatures. Mar. Ecol. Prog. Ser. 448: 247–258. Moore, K. A., E. C. Shields, and D. B. Parrish. 2014. Impacts of varying estuarine temperature and light conditions on Zostera marina (eelgrass) and its interactions with Ruppia maritima (widgeon grass). Estuar. Coasts 37 (Suppl. 1): S20–S30. New, M. 2002. Trends in freshwater and marine production systems. Pp. 21–27. In: D. Pauly and M. L. Palomares (eds.), Production Systems in Fishery Management. Vancouver, BC: University of British Columbia, Fisheries Center Research Report No. 10 (8). Nicholls, R. J., D. Linchke, J. Hinkel, S. Brown, A. T. Vafeidis, B. Meyssignac, et al. 2021. A global analysis of subsidence, relative sea level change and coastal flood exposure. Nat. Clim. Change 11(4): 338–342. Nixon, S. W. 1995. Coastal eutrophication: A definition, social causes, and future concerns. Ophelia 41(1): 199–220. Nixon, S. W. 2009. Eutrophication and the macroscope. Hydrobiologia 629(1): 5–19. NOAA. 2015. National Centers for Environmental Information. State of the Climate: Global Climate Report for December 2014. https://www​.ncdc​.noaa​.gov​/sotc​/global​/201412. Pauly, D., V. Christensen, J. Dalsgaard, R. Froese, and F. Torres, Jr. 1998. Fishing down marine food webs. Science 279(5352): 860–863. Perillo, G., E. Wolanski, D. Cahoon, and M. Brinson (eds.). 2009. Coastal Wetlands: An Ecosystem Approach. Amsterdam: Elsevier. Robb, C. K. 2014. Assessing the impact of human activities on British Columbia’s estuaries. PLOS ONE 9(6): e99578. http://doi​.org​/10​.1371​/journal​.pone​.0099578. Robins, P. E., M. W. Skov, M. J. Lewis, L. Giménez, A. G. Davies, S. K. Malham, et al. 2016. Impact of climate change on UK estuaries: A review of past trends and potential projections. Estuar. Coastal Shelf Sci 169: 119–135. Ruiz, G. M. and J. T. Carton 2003. Invasive Species: Vectors and Management Strategies. Washington, DC: Island Press. Schlüter, A., K. Van Assche, A.-K. Hornidge, and N. Văidianu. 2020. Land-sea interactions and coastal development: An evolutionary governance perspective. Mar. Policy 112: 103801. https://doi​.org​/101016​/j​.marpol​.2019​.10381.

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Climate Change and Saltwater Intrusion in Estuaries Arnoldo Valle-Levinson and Ming Li

CONTENTS Abstract.................................................................................................................................................................................99 6.1 Introduction.................................................................................................................................................................99 6.2 Factors Influencing Saltwater Intrusion Length in Surface Estuaries and Subterranean Estuaries........................ 100 6.2.1 Human-Related Factors................................................................................................................................. 100 6.2.2 Natural Factors.............................................................................................................................................. 101 6.2.2.1 Freshwater Discharge...................................................................................................................... 101 6.2.2.2 Evaporation..................................................................................................................................... 102 6.2.2.3 Wind................................................................................................................................................ 102 6.2.2.4 Tides................................................................................................................................................ 102 6.2.2.5 Ocean Currents............................................................................................................................... 103 6.2.2.6 Sea-Level Rise................................................................................................................................ 103 6.3 Climate Change Influence on Factors Affecting Saltwater Intrusion....................................................................... 103 6.4 Quantitative Approaches to Study Saltwater Intrusion in Surface Estuaries............................................................ 104 6.4.1 Exponential Function..................................................................................................................................... 105 6.4.2 Artificial Intelligence..................................................................................................................................... 106 6.4.3 Analytical/Theoretical Approaches............................................................................................................... 106 6.4.4 Three-Dimensional Numerical Simulations.................................................................................................. 107 6.5 Quantitative Approaches to Study Saltwater Intrusion in Subterranean Estuaries................................................... 108 6.5.1 Hydrostatic Approach – Ghijben–Herzberg Principle.................................................................................. 108 6.5.2 Dynamics Approach: Darcy’s Law............................................................................................................... 108 6.5.3 Mechanical Energy Approach: Bernoulli-Type Energetics........................................................................... 109 6.5.4 Numerical Simulations.................................................................................................................................. 110 6.6 Conclusion................................................................................................................................................................. 110 Acknowledgments............................................................................................................................................................... 110 References........................................................................................................................................................................... 110

ABSTRACT Saltwater intrusion in surface and subterranean estuaries (coastal aquifers) is affected by anthropogenic, or humanrelated effects, and natural factors. Human-related effects on saltwater intrusion are directly related to freshwater consumption and alterations to estuarine morphology. Moreover, the human influence on saltwater intrusion is indirectly linked to greenhouse-gas emissions. Natural factors include tides, winds, evaporation, river discharge, ocean currents, climate variability, and sea-level changes. However, human activities can alter these natural factors. The complex interplay between anthropogenic and natural agents in affecting saltwater intrusion is still poorly understood. Despite an incomplete understanding, artificial intelligence/statistical tools can be used, together with physics-based tools (analytical and numerical models), to guide the management of freshwater resources tied to surface and subterranean estuarine systems while DOI: 10.1201/9781003126096-7

attempting to mitigate the deleterious effect of saltwater intrusion. Key Words:  saltwater intrusion, surface estuaries, subterranean estuaries, climate change, tides, wind, ocean currents, sea-level rise, management

6.1 INTRODUCTION Clean water is indispensable for human subsistence and must remain free of salt contamination because of its deleterious health effects (e.g., Khan et al. 2014; Shammi et al. 2019; Sherin et al. 2020). In this chapter, saltwater intrusion is considered as the incursion of water of oceanic origin into freshwaters within coastal or near-coastal semienclosed basins. Such an incursion is identified in the basin of interest by absolute salinity values SA in true mass fraction, or g kg–1, in the absolute salinity scale of TEOS-10 (IOC et al. 2010). The key in this approach is to define a minimum value of SA that explains contamination by sea salts. For instance, for 99

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drinking water, values of 0.2 g kg–1 represent an approximate guide for salt contamination. These, and other low-salinity waters, however, can be used for other purposes such as powerplant cooling and irrigation. There is still ample work needed to determine what this minimum value should be. In this chapter, we arbitrarily consider values of interest in the range 1 < SA < 2. Of great relevance in the study of saltwater intrusion is determining the length (or position) of saltwater intrusion in a semienclosed basin, which will be denoted as ΛS. The position of ΛS defines the seaward extent of freshwater ecosystems, i.e., a habitat indicator, and governs a variety of chemical transformations that affect the carbon, nitrogen, and phosphorous cycles. The following section describes factors that influence ΛS, which ultimately plays a key role in determining the quality of freshwater resources. Saltwater intrusion varies over horizontal scales of less than one to tens of kilometers and at temporal scales ranging from tidal to seasonal to interannual to decadal. Saltwater intrusion affects the water quality and ecology of rivers (e.g., Jassby et al. 1995; Kaplan et al. 2010; Hutton et al. 2016), subterranean estuaries or coastal aquifers (e.g., Young et al. 2018; Pain et al. 2019), and soils (Bhuiyan and Dutta, 2012). Saltwater intrusion can also impact built infrastructure such as roads, electric grids, municipal drainage systems, water treatment plants, and buildings. This chapter centers on saltwater intrusion into rivers and coastal aquifers as it represents the main threat to life sustenance in coastal communities. A changing climate may bring reduced rain, droughts, decreases in river discharge, and rising sea-levels, which exacerbate these threats. From here on, we refer to semienclosed basins where riverine waters interact (or mix) with ocean waters as “surface estuaries.” Those environments where aquifer waters interact with waters of ocean origin will be referred to as “subterranean estuaries” (Moore 1999).

6.2 FACTORS INFLUENCING SALTWATER INTRUSION LENGTH IN SURFACE ESTUARIES AND SUBTERRANEAN ESTUARIES The factors that affect saltwater intrusion length Λ S in surface and subterranean estuaries may be distinguished

Climate Change and Estuaries

as anthropogenic and natural. In essence, anthropogenic agents are linked to water consumption or use, and morphologic and hydrologic modifications by humans. Natural factors are associated with environmental forcing and variability. Most factors described below, affect surface and subterranean estuaries. In order to manage water quality, in general, and the intrusion of salt, specifically, of any coastal freshwater resource, it is imperative to determine the relative importance of human-related agents versus natural factors that modulate intrusion.

6.2.1 Human-Related Factors Human-related factors typically involve the consumption or removal of freshwater from rivers or groundwater reservoirs and modifications to basin morphology (Figure 6.1). Such water removal should be quantified carefully as it most likely enhances saltwater intrusion. Factors that represent water removal by humans may be grouped into three major categories: (1) drinking and household use; (2) agriculture, including animal rearing and crop growth; and (3) industry and power generation. Basin morphology alterations also allow saltwater intrusion, typically by dredging or shoreline armoring (Figure 6.1). Undeniably, the most basic and widespread, but not most abundant, use of water is drinking. Household consumption is another part of daily life as demanded by bathrooms in sinks, toilets, showers, and bathtubs. Household use also involves laundering, kitchen activities related to cooking and dishwashing, cleaning activities in general, plus lawn and garden watering and irrigation. Per capita consumption of water in the USA is between 3,000 and 4,000 L day–1 (Dieter et al. 2018). With a per capita water use of 3 to 4 m3 day–1, a metropolitan area of 1 million people (the size of many metropolitan areas) in the USA would consume 35 to 46 m3 s–1. As a result, the largest world metropolis would withdraw non-negligible proportions if the rivers where the water plant intakes are located had mean discharges lower than ≈500 m3 s–1. Agricultural activities encompass the production of fruits and vegetables, as well as crops such as grains, seeds, herbs, sugarcane, cotton, tobacco, and rubber, among others. These activities also include the manufacturing of dairy

FIGURE 6.1  Anthropogenic and natural factors affecting saltwater intrusion in estuaries. Greenhouse-gas emissions, which are mostly human-related, may influence natural factors.

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products such as meat, fur, leather, and wool, as well as eggs. Much of the food that we consume, therefore, depends on water use, an excess of which contributes to saltwater intrusion. It is crucial to determine at what point there is an excess of water consumption for agricultural activities that promote saltwater intrusion. For reference, in the USA alone water withdrawals for agriculture amount to at least ≈8,800 m3 s–1 (Dieter et al. 2018), eliminating the combined average discharge of the rivers that enter the Mid-Atlantic Bight region. Industrial activities also represent a major contributor to water withdrawals. Among the most notorious extractors is the bottled water industry, with annual sales of >$6 billion in the USA (Ridder 2021). Hydraulic power production usually requires dams that hinder or reduce freshwater discharge toward estuaries. Dams also alter natural flow cycles and sediment budgets that configure estuaries. Paper and pulp manufacturing involves water at different stages of the production process. In general, many industrial engines and power plants are frequently cooled by the water that is withdrawn from rivers. Moreover, water withdrawals fill swimming pools, most notably in warm regions. In many cases, these agricultural and industrial activities contaminate water supplies by discharging effluents with dissolved and suspended chemicals. Such industrial activities also contribute to climate change as they require fossil fuels in their operations, thus increasing greenhousegas emissions. These emissions influence the climate and climate change affects saltwater intrusion in estuaries as discussed in Section 6.3.

The main human activity that allows for increased saltwater intrusion without freshwater extraction is bottom dredging. Bottom-dredging activities deepen an estuary to allow the navigation of increasingly larger vessels and to maintain those depths. Dredging essentially causes two physical responses in surface estuaries: modification of tides and increased baroclinicity, that is, propensity for ocean water intrusion into the estuary. The other activity that modifies estuarine morphology, shoreline armoring, is typically effected to mitigate flooding from tides or storms. However, armoring impedes the natural expansion of estuarine waters and most likely allows the re-direction and transport of salty waters further upstream into rivers. These human effects are explored in Section 6.3 together with sea-level rise, which also modifies tides and increases baroclinicity.

6.2.2 Natural Factors 6.2.2.1 Freshwater Discharge Freshwater discharge, in general, tends to counter saltwater intrusion, that is, increased discharge hinders intrusion in both surface and subterranean estuaries (e.g., Gong and Shen 2011; Parra et al. 2016; Bellafiore et al. 2021). For example, high river discharge hinders the intrusion of high salinity water in the Chesapeake and Delaware Bays (Figure 6.2). In Chesapeake Bay, however, the response to river discharge is more depth-dependent (Figures 6.2a–c) than in Delaware Bay (Figures 6.2d–f). There are instances, however, as in relatively deep surface estuaries where freshwater discharge

FIGURE 6.2  Along-channel distributions of salinity in Chesapeake Bay (left, a–c) and Delaware Bay (right, d–f) under different river discharge, obtained from an unstructured-grid model of the two estuaries and Mid-Atlantic Bight. (a) and (d) illustrate 2.5 times mean river discharge conditions; (b) and (e) display results under 1.5 times the mean river discharge; and (c) and (f) portray mean river discharge conditions.

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remains detached from the bottom, where the saltwater intrusion of near-bottom waters can actually increase with the enhanced river (or freshwater) discharge. This results from limited mixing between buoyant outflow and salty inflow. Fjords could exemplify this behavior of enhanced saltwater intrusion with increased discharge. Subterranean estuaries will generally show enhanced saltwater intrusion with reduced rain that recharges the aquifer (equivalent to freshwater discharge). The question of which surface estuaries will display enhanced or reduced saltwater intrusion, or increased vertical stratification, with increased river discharge, still lacks a universal answer (Valle-Levinson 2022). Each semienclosed basin will likely respond differently. Those that display increased saltwater intrusion with augmented river input will likely exhibit a threshold value of discharge, beyond which the intrusion will begin to decrease as the freshwater pulse will push the salty water seaward (e.g., Guerra-Chanis et al. 2019; Guerra-Chanis et al. 2022). These interactive processes remain to be explored regarding surface estuaries. Artificial Intelligence should allow the drawing of generalities through the production of hundreds to thousands of numerical simulations with different river discharge scenarios and distinct estuarine geometries. 6.2.2.2 Evaporation Evaporation represents a natural process of water extraction. The obvious consequence of evaporation in a semienclosed basin is for its salinity to increase, attaining values greater than ocean water (that is, becoming hypersaline). Under hypersaline conditions, typical density-driven estuarine circulation will reverse if the water density of the semienclosed basin is dominated by salinity and not by water temperature (Largier 2010). Such a reversal will consist of hypersaline water flowing seaward along the bottom, and ocean water entering the basin at the surface. Despite this saltwater extrusion toward the ocean being the opposite of saltwater intrusion into rivers, the mere contact of hypersaline waters with the basin’s edges will salinize soils. Soil salinization, in turn, can damage crops and contaminate aquifers. In subtropical and tropical regions with marked seasonality in the precipitation-evaporation cycle, dry seasons may threaten water and soil resources through the hypersalinization of semienclosed basins. 6.2.2.3 Wind Wind velocity can act to enhance or hinder saltwater intrusion (Valle-Levinson 2022). When winds blow down a stratified estuary, or seaward, it is likely that salt water in the bottom layer will intrude farther up the estuary, that is, against the wind. This response develops as wind drives surface (buoyant) waters seaward, which sets up a waterlevel slope that tilts upward toward the estuary’s mouth. Near-bottom waters will respond to this water-level slope, which reinforces the along-channel baroclinic pressure (density) gradient associated with the salinity field that increases seaward. In the same way, landward winds will

Climate Change and Estuaries

set up a water-level slope that tilts up toward the estuary’s head. This slope will oppose the baroclinic pressure gradient and delay saltwater intrusion. The response of saltwater intrusion to wind forcing will be modified by lateral variations in bathymetry (Guo and Valle-Levinson 2008; Chen and Sanford 2009). This is because wind-driven flows will develop most favorably over shallow portions of a cross-section, compared to the deepest portions. For example, wind-driven circulation in Albemarle Sound and northern Pamlico Sound, NC showed laterally sheared flows, with upwind currents in the deep portions and downwind currents over the shoals (Jia and Li 2012a). Under stratified conditions, the response will depend on the strength of vertical stratification, which determines whether the wind-driven flows ‘feel’ the bottom morphology. In addition to these local wind effects, remote winds blowing along the adjacent shelf can drive water into or out of an estuary (Wong 1994, Jia and Li 2012b). For instance, northerly winds during winter cold-air outbreaks result in enhanced saltwater intrusion into the North Branch of the Changjiang (Yangtze) River. This intrusion subsequently spills over the South Branch in spring tides and threatens freshwater supplies during winter (Zhang et al. 2019). Additional work is clearly needed to find generalities on the response of saltwater intrusion to wind forcing over laterally varying bathymetry. Furthermore, increased scrutiny is needed regarding responses to other atmospheric forcing agents such as barometric pressure, pluvial precipitation, evaporation, heat fluxes, and cloudiness. Under the influence of tropical storms, a coastal body of water may be subject to extreme storm surges (e.g., ValleLevinson et al. 2002). In these cases, wind forcing affects the ocean outside the semienclosed basin and drives an increase in water level that lasts from one to several days. Storm surges represent pulses of ocean water that intrude into estuaries and transport salty water. Flooding of lowlying areas may be compounded by excess precipitation related to the storm. When the ocean surge competes with a river surge, the saltwater intrusion length will be determined by the dominant agent or the phase lag between the incoming ocean surge and the outgoing river surge (ValleLevinson et al. 2020). Similar to the effect of river discharge on saltwater intrusion, Artificial Intelligence should be used to simulate hundreds to thousands of storms with reliable surges and pluvial precipitation in order to provide generalities on the challenging problem of compound flooding and its associated saltwater intrusion. 6.2.2.4 Tides Tidal forcing can influence saltwater intrusion at intratidal timescales (from hour to hour) and at subtidal timescales (from spring – or tropic – tides to neap – or equatorial – tides). An essential aspect of the influence of tides on saltwater intrusion is to determine the mechanism that dominates salt transport (e.g., Banas et al. 2004; Lerczak et al. 2006): (1) the covariance between tidal flow and salinity variations, also known as tidal pumping; or (2) the transport

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Climate Change and Saltwater Intrusion in Estuaries

by the subtidal or residual or exchange flow. In estuaries where tidal pumping dominates over exchange flow, such as in salt-wedge estuaries, maximum intrusion occurs with the strongest tidal currents as a consequence of maximum tidal advection. In these systems, such as the Mississippi River in the Gulf of Mexico or the Ebro or Po Rivers in the Mediterranean, saltwater intrusion is restricted to a few hours per day. On the other hand, in estuaries where exchange flows dominate over tidal pumping, as in most partially mixed or partially stratified estuaries, maximum intrusion occurs with the weakest tidal forcing (neap or equatorial tides). Decreased vertical mixing during weak tidal periods allows the development of enhanced stratification, exchange flows, and saltwater intrusion (Lerczak et al. 2006). Another way of determining whether tidal pumping or exchange flow will influence saltwater intrusion is to characterize the competition between tidal forcing and density gradients in determining the residual flows. Such competition can be assessed with the internal tidal Froude number, which compares tidal stresses to baroclinicity as forcing agents of the residual flow. The predominance of tidal stresses indicates that maximum saltwater intrusion should be expected in periods of greatest tidal forcing (ValleLevinson and Schettini 2016). Prevalence of baroclinicity implies that maximum intrusion shall appear during the weakest tidal forcing. As with river and atmospheric forcing, Artificial Intelligence shall help elucidate hundreds of simulations that explore the competition between tides and density gradients. 6.2.2.5 Ocean Currents Ocean currents can impact saltwater intrusion in estuaries by affecting exchange processes at the estuary’s entrance. The effect of ocean currents on exchange flows at an estuary occurs as follows. An ocean current, such as a western boundary current flowing nearly parallel to the coast (as the Gulf Stream in the Southeastern USA), is in approximate geostrophic balance. This means that the offshore water slope associated with the current (water elevation higher offshore than nearshore) is roughly in balance with the Coriolis accelerations. If the current relaxes, Coriolis accelerations decrease and to maintain the dynamic balance, the flow will move down the offshore slope driving water into the estuary’s entrance, favoring a subtidal transport that should enhance saltwater intrusion (e.g., Parra et al. 2016). Similarly, the acceleration of an ocean current will cause a water level set-down at the estuary’s entrance, driving subtidal outflows that hinder saltwater intrusion. Important to consider here is that a seemingly irrelevant factor, such as the strength of a remote current, may affect the length of saltwater intrusion in an estuary. Decadal and longer-term variations in ocean current strength likely influence the length of saltwater intrusion in estuaries. These decadal and longer variations can be caused by the ocean–atmosphere internal coupling, such as the El Niño–Southern Oscillation, the North Atlantic

Oscillation, the Pacific Decadal Oscillation, the Atlantic Multidecadal Oscillation, or the Atlantic Meridional Overturning Circulation. Some of this long-term variability may also have astronomical (e.g., Valle-Levinson and Martin 2020; Valle-Levinson et al. 2021) and anthropogenic influence. The latter arises from greenhouse-gas content in the atmosphere, which may vary from year to year and cause redistributions of water mass on earth and increased pressure of oceans on estuaries. As with the other natural factors that affect saltwater intrusion, examining long-term (>100 years) records, at least of water level and ideally of salinity, will help elucidate the influence of these long-term variations of ocean currents in saltwater intrusion. 6.2.2.6 Sea-Level Rise Sea-level rise is affecting saltwater intrusion in a similar way as weakening ocean currents. This process exacerbates saltwater intrusion into estuaries by driving increased volumes of ocean water into estuaries and aquifers (e.g., Hong et al. 2020; Bhuiyan and Dutta 2012; Bricheno et al. 2021). The process counteracts the pressure gradient associated with rivers and aquifers that drive freshwater seaward. As with ocean waters, human influence on sea-level rise and saltwater intrusion is evident. Increased greenhouse gases in the atmosphere contribute to ocean warming and thermal expansion of seawater, as well as the melting of land-based ice and land-anchored ice that increases the volume of the ocean. Moreover, human activities related to freshwater use compound the effects of sea-level rise in promoting saltwater intrusion into estuaries and aquifers. All “natural” factors described in this section, which affect saltwater intrusion, are modified by climatic change.

6.3 CLIMATE CHANGE INFLUENCE ON FACTORS AFFECTING SALTWATER INTRUSION The following descriptions related to the influence of climate on saltwater intrusion represent fertile ground for research. Freshwater discharge may increase or decrease with climate change. A changing climate may bring extended periods of drought, alternating with occasional severe freshwater pulses, leading to long-term trends of increased saltwater intrusion. For example, climate downscaling projections for the Sacramento/San Joaquin watersheds and the downstream San Francisco Bay for 2090 suggested ~20% decreases in the spring river runoff and spring/summer salinity increases of up to 9 g kg–1 (Knowles and Cayan 2002). In some instances, however, river discharge may be augmented by flooding rains and cause saltwater intrusion to decrease. This will obviously depend on the response of an estuary to increased freshwater inputs, which ultimately depends on the estuary’s morphology (Dai et al. 2011; Gong and Shen 2011; MacCready et al. 2018). In addition, changes in continental shelf salinity may affect estuarine salinity through inflowing waters.

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Evaporation rates will likely increase in some regions of the planet, thus exacerbating soil salinization and aquifer contamination with salt. In some instances, evaporation may even become a dominant driving force for estuarine circulation in places where it is not yet a factor. This switch in forcing dominance may cause a reversal in the circulation regime, from typical to inverse (Largier 2010). In other cases, evaporation may increase the seasonality in estuarine hydrodynamic behavior, switching the direction of saltwater transport from wet to dry seasons. These types of estuaries will become more vulnerable to water quality threats because of a reduced flushing time during the evaporative period (Largier 2010). Winds may become a more prominent agent in modifying saltwater intrusion than presently. This could happen if wind speed increases with atmospheric pressure gradients through enhanced warming on land. In some cases, saltwater intrusion will be exacerbated; in others, it will be delayed depending on predominant wind directions, as in Section 6.2.2.3. Moreover, as storms become more powerful with a changing climate, the response of saltwater intrusion to atmospheric forcing will be more pronounced. Tides could be affected by the deepening of estuaries in response to sea-level rise or dredging. Deepening a semienclosed basin could lead to increased tidal amplitude and currents resulting from decreased bottom frictional effects (e.g., Cai et al. 2012; DeJonge et al. 2014; Ralston et al. 2019). However, the tidal range could also decrease with sea-level rise (e.g., Holleman and Stacey 2014; Pickering et al. 2017; Lee et al. 2017; Du et al. 2018). Deepening of estuaries should also produce enhanced baroclinicity, as this forcing is hypersensitive to depth (baroclinicity has a cubic dependence on depth, e.g., MacCready 2007). In some cases, however, the mean depth of an estuary may decrease with sea-level rise as low-relief coastlines become inundated. For example, increased bottom salinities have been linked to increasing sea-levels in the Chesapeake and Delaware Bays (Figure 6.3; Hilton et al. 2008; Ross et al. 2015). Similarly, sea-level rise has increased salinity, stratification, and tidal range in the Pearl River Estuary, China (Hong et al. 2020). Channel deepening in the lower Hudson estuary, NY, has led to increased contamination risk to drinking water in the upper estuary (Hoagland et al. 2020). In addition, channel deepening and widening in Tampa Bay, FL, has resulted in larger tidal ranges and increased saltwater intrusion (Zhu et al. 2015). In comparison, salinity in the North Branch of China’s Changjiang River decreased as the North Branch became much shallower and narrower from 2007 to 2017 due to land reclamation on the tidal flat (Chen et al. 2019). Regardless of whether the dominant mechanism responsible for saltwater intrusion in an estuary is tides or baroclinicity, saltwater intrusion will most likely increase with a deepening of the basin. Ocean currents may weaken in response to climate change as the Atlantic Meridional Overturning Circulation slows down from warming and freshening of the North Atlantic Ocean around Iceland (e.g., Caesar et al. 2018;

Climate Change and Estuaries

FIGURE 6.3  (a) Time series of monthly average water level at Baltimore between 1950 and 2000. (b) Residual salinity time series at a grid cell in the surface water of the middle Chesapeake Bay (adapted from Hilton et al. 2008). (c) Residual salinity time series at a monitoring station in the upper Delaware Bay (adapted from Ross et al. 2015). The red lines are linear fits to the longterm trends.

Dima et al. 2021). In the Chesapeake Bay, for instance, bottom salinity was correlated to a Gulf Stream index (Lee and Lwiza 2008), and could also be explained by changes of salinity on the shelf of the southern Middle Atlantic Bight (Hilton et al. 2008). The weakening of ocean currents, together with the thermal expansion of a warming ocean and the melting of land-based ice, may also contribute to sea-level rise. In both instances of currents weakening and sea-level rise, saltwater intrusion into estuaries and aquifers is expected to increase. On balance, saltwater intrusion is already a pressing threat to freshwater resources in parts of the world. Together with the above agents, population growth and water consumption will worsen the problem of salinization of freshwater resources. Adequate management of these resources will require improved knowledge of all the factors outlined in previous paragraphs and their modifications by climate change.

6.4 QUANTITATIVE APPROACHES TO STUDY SALTWATER INTRUSION IN SURFACE ESTUARIES There are at least four general approaches to quantitatively examining the response of a surface estuary to saltwater intrusion: (1) use an ad hoc exponential function that approximates the saltwater intrusion length Λ S as an exclusive function of river discharge, or river discharge and tidal currents; (2) apply statistical relationships obtained with artificial intelligence such as Markov Chain Monte Carlo (Autoregressive) Simulations, Neural Networks,

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Climate Change and Saltwater Intrusion in Estuaries

Deep Learning, or Machine Learning; 3) devise analytical approximations or simplifications of the three-dimensional equations for salt and momentum conservation; and 4) implement numerical solutions of the complete threedimensional (or four-dimensional when considering time) equations of salt and momentum conservation (plus other equations needed to mathematically close the system).

6.4.1 Exponential Function This is the simplest, but perhaps, most incomplete approach to studying saltwater intrusion. It is the simplest because the only agent considered as limiting the length of saltwater intrusion Λ S is river discharge QR. As described in previous sections, there are several other factors that actually determine Λ S. Thus, it is the most incomplete approach. On the basis of observations in surface estuaries, Λ S and QR seem to follow an exponential decay relationship:

LS ~ QR -1/ m (6.1)

where m is a real number and 1/m denotes the rate of decay of Λ S as QR increases. Clearly, this relationship applies to estuaries where Λ S increases as QR decreases, which is typical of partially mixed estuaries but different from expectations in dynamically deep basins-like fjords. The challenge

with Equation (6.1) is that each estuary seems to have a different value of m that needs to be determined with a wide array of measurements of the salinity field and its corresponding river discharge. In particular, the approximation in Equation (6.1) can be specified as:

LS = g 1QR -1/ m + g 2 (6.2)

The three free parameters g 1, g 2 , and m are found by fitting measurements of LS and QR to Equation (6.2). The values of g 1 and g 2 determine the end members of LS at extreme values of QR (Figure 6.4, Generic Estuary). Examples of saltwater intrusion into estuaries are illustrated in Figure 6.4. In the St Johns River estuary, a variably stratified estuary on Florida’s northeastern coast, the value of m is around 0.35 and g 1 is between 1×107 and 1×108, which results from a relatively small range of river discharge. In this estuary, most of the freshwater discharge is related to groundwater discharge and, therefore, the narrow range in QR . The behavior of Equation (6.2) for the St Johns River estuary would predict large values of LS with decreased values of QR , below the typically expected values, that is, when QR becomes 8. Similarly, the eutrophic Neuse River Estuary in North Carolina consistently experiences responses

Biogeochemical Changes in Estuaries

wherein the initial load of organic matter and sediments suppresses primary production. After flow rates subside and light limitation is reduced, the nutrients delivered to the estuary along with OM further fuel plankton blooms (Paerl et al. 2018; 2019a; 2019b; 2020). Likewise, the flush of nutrients in lagoonal Mission Aransas Estuary after spring storms accelerated primary production which resulted in discrete changes to POM stable isotope and elemental values, as well as enabling cyanobacterial blooms (Reyna et al. 2017). Thus, estuaries tend to experience a duality of responses to high discharge events related to increased nutrient and OM fluxes into them, potentially offsetting biological responses. During drought, low river discharge restricts freshwater, nutrient, and OM inputs to estuaries. Generally, patterns of nutrient and OM concentrations in estuaries are lower, and primary production similarly responds with lower values, although such patterns are not always found. For example, in a 14-year time series for the river-dominated Apalachicola Bay, no change in chlorophyll abundance was detected between drought and non-drought years; in this case, enhanced light availability during droughts linked to lower turbidity and/or suspended sediment levels may have offset the effects of reduced nutrient supply on primary productivity (Geyer et al. 2018). Moreover, lower nitrate and phosphate inputs in drought years from reduced river flow were postulated to be offset by sediment resuspension (especially near oyster reefs) where elevated ammonium concentrations could satisfy phytoplankton N demand. Drought may also modestly improve oxygen dynamics in estuaries, thereby offsetting the development of hypoxia (Wetz et al. 2011). However, drought conditions via nutrient dynamics and salinity also can strongly influence phytoplankton community composition. In semi-arid Baffin Bay, Texas, drought conditions favored the development of nuisance brown tides dominated plankton biovolume. Interestingly, these trophodynamics were linked to the phase of ENSO with drought conditions prevalent in the non-El Nino phase (Cira et al. 2021). Wetter periods associated with the El Niño phase led to lowered salinities, increased nutrient inputs, and led to higher plankton community diversity. Release of DOM and nutrients was observed during low flow conditions in the Neuse River Estuary where wind-driven mixing was important (Dixon et al. 2014). In Suisun Bay, California, OM switching under a low flow of the Sacramento River was suggested to maintain the supply of DOC, perhaps released by shoals/tidal flats, or from POC (Murrell and Hollibaugh 2000). Thus, nutrient and OM retention and recycling within estuaries may result from low flow conditions, with attendant effects on estuarine trophodynamics. Drought coupled with SLR can also facilitate saltwater intrusion inland (Figure 7.1), impacting biogeochemical processes in surface and subterranean estuaries. Landward movement of seawater can serve as an effective physical barrier to material exports from coastal freshwater wetlands to adjacent estuaries (Ardón et al. 2016) and coastal

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aquifers in which the mixing of freshwater and saltwater defines the subterranean estuary (Moore and Joye 2021). However, the effect of SLR to move saltwater inland also causes biogeochemical changes in both surface and subterranean systems. Increasing salt content reduces DOC export as indicated by drying and salinization experiments which together reduced DOC concentrations in soil solutions by 50% and showed the synergistic effects of drought and SLR (Ardón et al. 2016). Other effects included enhanced POM formation from flocculation, which occurs at salinities 100 m) coastal seas and fjords, strongly stratified water columns result in virtually permanent hypoxia/anoxia, changing in size and position with decadal-scale variations in circulation (Zillen et al. 2008; Pakhomova et al. 2014). Importantly, shallow estuaries typically also have more dynamic hydrography than coastal seas and fjords, implying that a relatively small volume of O2-deficient water can affect larger areas of the estuary when winds generate advective transports in the bottom layer. While many of the published projections of climate-induced O2 variations have been done in seasonal or persistent hypoxic zones (Table 8.1), fewer studies have examined the effect of climate on shallower low-O2 environments. In the text that follows, we address each of the potential variables vulnerable to climate change and how they may affect hypoxia, followed by a discussion of their particular relevance in shallow nearshore systems.

8.2.1 Temperature Perhaps the most widely anticipated feature of climate change is warming, which has already been documented in several estuaries with water temperature increases of at least 1°C over the past 30 years (Nixon et al. 2004; Carstensen et al. 2014; Hinson et al. 2021; Whitney and Vlahos 2021). Warming has the potential to increase the intensity, duration, and spatial extent of hypoxia by increasing stratification, decreasing the solubility of O2 in seawater, and increasing the rate of organic matter remineralization and thus O2 consumption. At a given atmospheric pressure, O2 solubility is primarily determined by salinity and temperature, where solubility is reduced with increases in both factors (Garcia and Gordon 1992), thus causing high temporal variability in solubility. There are numerous examples of declining O2 concentrations in coastal waters that appear to have resulted from temperature increases. These include the St. Lawrence estuary (Gilbert et al. 2005), the northern

Coastal shelf

Partially mixed, coastal plain Partially mixed, coastal plain

Gulf of Mexico

Mid-Atlantic, USA North-Atlantic, USA

Fjord, Partially mixed

Large coastal sea

Temperature, SLR*, Flow

Seasonal

Seasonal

Seasonal

Episodic

Seasonal

Seasonal

Temperature, Elevated Oceanic and Anthropogenic Nutrient Loads, Boundary Deoxygenation&

Temperature

Temperature

Temperature, Flow

Temperature, Eutrophication Temperature, Population Growth, SLR*, Flow Temperature

Temperature

Seasonal

Seasonal/ Permanent Episodic

Temperature, SLR*, Flow

Temperature, Boundary Deoxygenation&

Major Driver(s)

Seasonal

Seasonal

Hypoxia Type/scale

+16%

–35%

+5–+30 days

+8–10%

+20%

+91 days

+14%

5–45%

+9%

+10–30%

+20–30%

Projected Change

Change from 2000 to 2095

2099

5° warming

2100

2100

2100

2100

Mid-21st century Change from 1995 to 2005 Mid-21st century

0.75 and 1.25°C warming

Projection Condition

Hypoxia Area/Duration

Range of Increase in Days with Daily Mean Oxygen < 2 mg/L Reduction in ManagementInduced Oxygen Concentration Increase

Hypoxia Severity (>60 days)

Hypoxic Area/Duration

Annual Days of Hypoxia

Lost Reduction in ManagementInduced Anoxic (< 0.2 mg/L) and Moderately Hypoxic (< 5 mg/L) Volume Hypoxic Area

Summer Hypoxic Volume

Hypoxic, Anoxic Volume

Hypoxic Volume

Hypoxia Metric

Khangaonkar et al. 2019

Whitney and Vlahos 2021

Lake and Brush 2015

Lehrter et al. (2017)

Lajaunie-Salla et al. 2018 Laurent et al. (2018)

Meier et al. 2011

Irby et al. 2018

Tian et al. 2021

Ni et al. 2019

Testa et al. 2021

Reference

Note: The reported changes represent responses to major drivers and the reported values are estimated ranges (not precise numbers) derived from figures and tables included in the cited publications. (* = Sea-level rise; & = boundary deoxygenation indicates import of lower oxygen water due to climate effects on adjacent waters). Note: Lake and Brush (2015) also reported simulations of +1, +2, and +3°C.

Salish Sea

Long Island Sound

Northern Gulf of Mexico Northern Gulf of Mexico York River

U.S. Pacific Northwest

Turbid, river estuary Coastal shelf

Gironde

Northern Europe Northern Europe Gulf of Mexico

Baltic Sea

Partially mixed, coastal plain Partially mixed, coastal plain Partially mixed, coastal plain

Mid-Atlantic, USA Mid-Atlantic, USA Mid-Atlantic, USA

Mainstem Chesapeake Bay Mainstem Chesapeake Bay Mainstem Chesapeake Bay

Partially mixed, coastal plain

Estuary Type

Mid-Atlantic, USA

Region

Chester River estuary

Estuary/System

TABLE 8.1 A selection of projected changes in metrics of dissolved oxygen in response to climate change reported in the literature

Hypoxia and Climate Change in Estuaries 147

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Gulf of Mexico (Laurent et al. 2018), the Chesapeake Bay (Ni et al. 2020), the Danish Straits (Conley et al. 2007) and many other systems. While there is ample evidence that rising temperatures will speed up the respiration of freshly deposited organic matter, ultimately respiration is determined by the amount of organic matter available. We will revisit these ideas in our case study section below.

8.2.2 Riverine Inputs There are multiple potential impacts of changes in freshwater input on O2 depletion. Larger discharges would increase stratification and nutrient loading, thus amplifying hypoxia and promoting eutrophication, but if larger discharges dilute nutrients (assuming constant load), hypoxia could potentially be limited. In many coastal regions experiencing hypoxia (e.g., the northern Gulf of Mexico, the East China Sea, and the Chesapeake Bay), salinity has a much stronger effect on stratification than temperature does, as suggested by the model simulations of Laurent et al. (2018) for the northern Gulf of Mexico and Irby et al. (2018) for the Chesapeake Bay. Precipitation amount, frequency, and intensity in the watershed are also major controls on riverine nitrogen (N) load (Lee et al. 2016). By combining an empirical model of N loading with the Coupled Model Intercomparison Project Phase 5 (CMIP5) climate projections, Sinha et al. (2017) predicted that projected increases in total and extreme precipitation will elevate total N loading in the continental USA by approximately 20% by 2100 in the business-as-usual scenario (RCP8.5; Riahi et al. 2011). Projected increases are especially pronounced for the northeastern USA and the upper Mississippi/Atchafalaya River basin, regions with historically high N fluxes. Sinha et al. (2017) also suggested that large portions of east, south, and southeast Asia, including India and eastern China, have similar risk factors and are therefore likely to experience large precipitation-driven increases in N load. Increasing precipitation will also enhance the phosphorus (P) load (Ockenden et al. 2017), which stimulates primary production in many estuaries, particularly in spring. For example, Jeppesen et al. (2009) estimated a P load increase between 3.3 and 16.5% by 2100 for Danish streams. Soil erosion is considered a major pathway for P in many catchments, and the expected higher frequency of flash floods may cause river bank erosion and, consequently, larger P loads to estuaries and coastal waters. However, due to the complexity of these processes, there are few assessments of climate change effects on the P cycles (Alewell et al. 2020). On the other hand, lower future discharges could reduce flushing rates in estuaries that may promote deoxygenation. Residence time is an important control on eutrophicationdriven hypoxia (Fennel and Testa 2019), and reduced river discharge has been related to increasing hypoxic conditions in small Mediterranean climate–type estuaries, which is expected to continue in the future (Cottingham et al. 2018; Warwick et al. 2018). In many coastal systems, vertical gradients in salinity and temperature cause density stratification, which is

Climate Change and Estuaries

often sufficient to isolate the bottom from surface waters (Pohlmann 1996). An important consequence of this vertical separation of water masses is the restriction of downward mixing of O2 from surface waters, thereby reducing physical replenishment and allowing depletion of bottom water O2 (Turner et al. 1987; Kemp et al. 1992). The buoyancy of the upper layer is increased and stratification is strengthened by seasonal inputs of freshwater and to some extent the warming of surface waters (Welsh and Eller 1991). The location of a permanent halocline at deeper depths in systems such as the Black Sea and Baltic Sea is maintained by subpycnal saltwater inflows and or limited external water exchange, and thus wind-induced vertical mixing has a minor effect on stratification. Weaker stratification can result from low freshwater inputs, shallow depths, or strong tides and can be disrupted by summer wind events in systems such as (in the USA) the Neuse River estuary (NC), Long Island Sound (NY/CT), and Mobile Bay (AL) (Turner et al. 1987; Stanley and Nixon 1992; O'Donnell et al. 2008). In years of relatively high freshwater input and/or surface water temperatures, enhanced stratification, which is more resistant to disruption by wind events, will lead to increased bottom water O2 depletion (Lin et al. 2008). Weakening of stratification and consequential ventilation of bottom waters may involve various complex mechanisms, where wind stress induces straining of density fields (Scully et al. 2005), lateral tilting of the pycnocline (Malone et al. 1986), alteration of far-field coastal circulation (Wiseman et al. 1997), or interaction with spring-neap tidal cycles (Sharples et al. 1994). In any case, stratification controls the downward mixing of O2, which is the primary mode of O2 replenishment for many coastal bottom waters.

8.2.3 Eutrophication as a Mediator Hypoxia is driven by the imbalance between O2 supply and consumption, and O2 consumption is strongly coupled to the respiration of organic matter that can be supplied from land (allochthonous) or produced within the estuary (autochthonous). Autochthonous organic matter production is enhanced by increased watershed nutrient loads. The first historical records of coastal hypoxia and anoxia were observed in urban estuaries associated with large inputs of O2-consuming organic and inorganic pollutants from sewage treatment plants (Andrews and Rickard 1980; Sharp 2010). As wastewater treatment plants have been upgraded in the vicinity of many estuaries worldwide, diffuse sources of nutrients from agriculture have become the dominant drivers of eutrophication (and associated hypoxia) in most estuaries. Although nutrient inputs have been reduced for a small but growing number of coastal systems worldwide (Riemann et al. 2015; Zhang et al. 2015; Taylor et al. 2020), nutrient-induced eutrophication has persisted in a large number of estuaries, sustaining extensive areas of hypoxia (Bricker et al. 2007; Díaz and Rosenberg 2008). Eutrophication is primarily limited by nitrogen (N) for the majority of marine systems, whereas both N and P

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are important for brackish systems such as the Baltic Sea (Vahtera et al. 2007). Transient P limitation also occurs in estuaries and river plumes, and therefore dual nutrient regulation is needed (Paerl et al. 2016b). The major factor determining N losses from many watersheds is N input through farming and other land use practices. Most developed countries have implemented measures to reduce excess N losses from agriculture, and reductions in diffuse N loads have been observed in the Baltic Sea (HELCOM 2018), the Chesapeake Bay (Zhang et al. 2015), to name a few. Seitzinger et al. (2010) showed that the future trajectory of N loading is highly dependent on socioeconomic drivers, where the Global Orchestration scenario (which includes intensive agriculture and rapid increases in fertilizer use in developing countries) leads to predictions of increasing N loads, while the Adapting Mosaic scenario (which assumes modest increases in global fertilizer use and moderate improvements in its efficiency) suggests N loads would decrease for many countries. Socioeconomic pathways to mitigate climate change (e.g., relying on biofuel) may also have a substantial effect on N loads (Sinha et al. 2019). Regardless of the scenario, large changes in nutrient loads to estuaries are expected in south, east, and southeast regions of Asia (Seitzinger et al. 2010; Sinha et al. 2019). Eutrophication and coastal hypoxia, which have been considered an environmental problem mainly in developed countries, will become more prominent across the entire world in the future unless technology and regulation are implemented to strongly reduce nutrient losses from agriculture.

8.2.4 Changes in Wind Patterns Projections of changes in future wind forcing are highly uncertain but can affect coastal hypoxia in significant ways. Alterations to wind forcing occur on a variety of timescales, including multidecadal oscillations, such as the NAO or PDO (Scully 2010a), seasonal shifts in pressure fields and wind directions (Connolly et al. 2010), event-driven mixing events (Li et al. 2007), and diel-scale sea breezes (Duvall et al. 2022). In eastern boundary current systems, wind forcing is not only the driver of nutrient supply to the continental shelf but also responsible for the upwelling of low-O2 source waters, which preconditions the Namibian and Oregon shelves to low-O2 conditions. A long-term shift toward more upwelling-favorable winds has been implicated in the recent emergence of hypoxic and anoxic conditions on the Oregon shelf (Grantham et al. 2004; García-Reyes et al. 2015). In the Chesapeake Bay and Long Island Sound, observational and modeling studies have shown the importance of both wind speed and direction in influencing the extent of hypoxia over multi-year timescales (Wilson et al. 2008; Scully 2010b; Li et al. 2015). Increases in wind strength could decrease hypoxia by eroding vertical stratification and aerating bottom waters, especially in shelf regions without permanent stratification, like the northern Gulf of Mexico (Yu et al. 2015). The direction of

the wind influences the size and location of hypoxic conditions in river plumes (Feng et al. 2012; Zhang et al. 2020), and therefore future changes may exacerbate or mitigate the effects of climate change on hypoxia (see case studies below).

8.2.5 Sea-Level Rise Sea-level rise can affect hypoxia in multiple ways. Sea-level rise increases the mean depth in an estuary and may result in stronger estuarine circulation (see Chapter 4), leading to stronger inflows of oxygenated shelf water into the estuary (Wang et al. 2017; Irby et al. 2018). On the other hand, sealevel rise may increase vertical stratification due to stronger salt intrusion and increases in bottom salinity, suppressing the downward diffusion of dissolved O2 into the bottom hypoxic water (Ni et al. 2019). Some modeling studies of the Chesapeake Bay showed that sea-level rise decreases dissolved O2 concentrations in the upper part of the water column and increases bottom dissolved O2 at least in parts of the deep channel in May–July (Wang et al. 2017; Irby et al. 2018). O2 increases could result from increases in the water depth that decrease water temperature if the air–sea heat flux stays constant, resulting in weaker respiration in bottom waters (St-Laurent et al. 2019). Other studies that considered warming showed that sea-level rise contributes to larger hypoxic volumes in bottom waters via elevated stratification (Ni et al. 2019; Ni et al. 2020) or a shallowing of the pycnocline (Cai et al. 2021). The net effect of sea-level rise on estuarine hypoxia ultimately depends on the competition between stronger vertical stratification, stronger inflow, and alterations of tidal mixing. Modeling investigations of the Baltic Sea, for example, showed a small effect of sea-level rise on hypoxia between 1850–2008, but a future sea-level rise over 1 m may lead to reinforced saltwater inflows into the deep water that increases stratification, expand anoxia and hypoxia, and thus elevate internal phosphorus loads from the sediment (Meier et al. 2017). Depending on the coastline management strategies in response to sea-level rise (e.g., construction of sea walls), the tidal range may increase or decrease with sea-level rise (Lee et al. 2017), with corresponding changes in tidal-induced turbulent mixing which will affect the vertical supply of dissolved O2 to the bottom water.

8.3 CASE STUDIES OF HYPOXIA IN RESPONSE TO CLIMATE CHANGE Here we provide a limited number of case studies that synthesize the expected response of hypoxia to climate change. We highlight these well-studied systems because sufficient model projections have been performed to provide a summary of expected changes. We provide a summary of projected changes in metrics of hypoxia from the literature in Table 8.1, which serves to illustrate the ranges of change projected for different scenarios of climate changes. These example systems include larger estuaries and continental

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shelves and represent deeper estuarine environments. Thus, they do not represent the shallow, nearshore estuaries where projections have rarely been made and where we aim to focus our discussion in the next section (see Section 8.4).

8.3.1 Chesapeake Bay Hypoxia in the Chesapeake Bay experienced a rapid expansion between the 1950s and 1980s, associated with large increases in nitrogen loading (Officer et al. 1984; Hagy et al. 2004; Kemp et al. 2009). Due to nutrient reduction efforts associated with the enactment of environmental protection laws, riverine nutrient concentrations have leveled off over the past three decades. Summer hypoxia, however, persisted and exhibited large interannual variations, with more extensive hypoxia in wet years than in dry years (Hagy et al. 2004; Bever et al. 2013). Several studies attempted to discern the physical and/or biogeochemical mechanisms driving this interannual variability in hypoxia. Scully (2016) conducted 30-year (1984–2013) hindcast simulations by coupling a 3D hydrodynamic model forced with observed variations in physical drivers and a constant biological consumption rate, and found that June–August wind speed and January–June Susquehanna River discharge are the two variables that have the highest correlation with the observed variability in the hypoxic volume time-series, consistent with statistical modeling analyses (Murphy et al. 2011; Lee et al. 2013). However, statistical correlations do not necessarily correspond to causal mechanisms. To discern the physical and biogeochemical drivers of the interannual hypoxia variability, Li et al. (2016) used a coupled hydrodynamic–biogeochemical model (Testa et al. 2014) to analyze the O2 budget of the bottom water over a 10-year period where both physical forcing and O2 consumption varied from year-to-year. Li et al. (2016) concluded that variations in water column respiration associated with changes in nutrient loading were the main mechanism driving interannual hypoxia variability, consistent with a eutrophication effect. Clearly, both physical and biogeochemical controls on hypoxia variability are relevant in the Chesapeake Bay, and due to the co-variance of many of these drivers (e.g., freshwater flow and nutrient input), discerning relative controls is difficult. Climate effects on hypoxia in the Chesapeake Bay may be realized through more than interannual variations and long-term trends. Retrospective analyses of water quality monitoring data since 1985 found no long-term trends in the magnitude of summer hypoxia but an apparent shift in the seasonal hypoxia cycle to an earlier peak. Murphy et al. (2011) and Testa et al. (2018) found significant increases in early summer hypoxia but a slight decrease in late summer hypoxia. Zhou et al. (2014) reached a similar conclusion, suggesting that the timing of the maximum hypoxic volume shifted from late to early July but no long-term trend was detected in the seasonal maximum of hypoxic volume. Both Zhou et al. (2014) and Testa et al. (2018) found that hypoxia terminated earlier in the fall but did not detect

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significant shifts in the timing of hypoxia onset. Testa et al. (2018) hypothesized that the earlier shift of the summer hypoxia cycle was linked to the altered external forcing such as nutrient loading and water temperature. To ascertain what drove the long-term changes in O2 in the Chesapeake Bay over the past 30 years, Ni et al. (2020) used a coupled hydrodynamic–biogeochemical model to conduct hindcast simulations between 1985 and 2016. After the removal of seasonal and interannual variations, model-predicted dissolved O2 concentration in all regions of the estuary showed a statistically significant declining trend: ~0.1 mg L –1 per decade (Figure 8.2a–e). Most of this decline occurred during winter and spring while May–August hypoxic volumes showed no changes and September hypoxic volume showed a slight decrease (Figure 8.2f–i), in agreement with the retrospective data analysis. Ni et al. (2020) found no statistically significant trend in the initiation of hypoxia in spring because the delayed onset resulting from nutrient reduction was offset by warming-induced O2 declines. However, both nutrient reduction and warming contributed to an earlier disintegration of hypoxia in the fall, as less organic matter accumulated to sustain hypoxia and warming increased early-season respiration rates. In another scenario modeling analysis for four select years (2016-2019), Frankel et al. (2022) found a larger impact of nutrient reduction on annual hypoxic volume (O2 90% of the total ecosystem carbon stock (TECS). In comparison with some marsh and mangrove biomass datasets, aboveground biomass averages less than 0.5 kg C m–2 in temperate tidal marshes (Byrd et al. 2018) and with means of 3–25 kg C m–2 across the world’s mangrove regions (Kauffman et al. 2020). Macroalgal and mudflat (tidal flat) environments lack surficial vegetation, but can still have organic stocks preserved in soils (Schile et al. 2019). Belowground pools dominate organic carbon stocks but are increasingly difficult to predict both in density and depth (Adame et al. 2017). Geomorphic conditions currently provide the best constraint for mangrove C stock distributions, leading to higher soil organic carbon (SOC) densities in carbonate settings and lower SOC densities in deltaic clastic soils (Rovai

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et al. 2018), but finer scale variability emerges as a function of both space and time (depth) in these dynamic environments (Breithaupt et al. 2020). Preservation of the soil C pool is primarily through hydrologically limited microbial activity, with a secondary role of stabilization through post-decay mineral association (e.g., Spivak et al. 2019). Vulnerabilities of legacy soil stocks to erosion and microbial mineralization are likely enhanced by climate and land use drivers. Models of soil carbon stock accumulation, calibrated by long-term data, illustrate the spatial and temporal variability of soil carbon densities, sensitivities of processes responsible for carbon stocks and fluxes, as well as their sensitivity to drivers acting at different time scales (e.g., Drexler et al. 2009). For example, depth age modeling in U.S. tidal wetlands shows us that carbon accretion accelerates with sea-level rise and after initial compaction, soil carbon densities remain stabilized at depth (Morris et al. 2016; Gonneea et al. 2019). Deep soil carbon in BCEs experiences limited variability in microbial activity, with very little evidence through models (Schoolmaster et al. 2022) or through isotopic examination of C emissions (Miller 2011) or hydrologic export (Gonneea et al. 2019) from subsurface soils, unless a disturbance has liberated Holocene soil pools (e.g., Sapkota and White 2021). Most biogeochemical and geomorphic models predict C impacts from climate disturbances through physical and microbiological impacts (Megonigal et al. 2019). Biologically, atmospheric warming leads to the stimulation of productivity (Kirwan et al. 2009) but also to the stimulation of microbial respiration (Carey et al. 2016), which typically overcompensates photosynthetic responses. Field experiments with temperature enhancement clearly stimulate evapotranspiration rates, thereby concentrating solutes in porewater and often cooling soils, in direct contrast with the warmer air (Carey et al. 2018). More clearly, whole soil warming experiments illustrate a strong temperature effect on productivity, especially on methane emissions, but no clear evidence of destabilizing millennial carbon (Smith et al 2022, Noyce et al. 2023). Still, deeper-rooted invasive species can expand the active layer and thus provide stimulation of microbial activity at depth (Muller et al. 2016). Geomorphically, climate impacts are especially driven by surface water flows, not just surface water levels. Windor channel-induced erosive processes can overtake depositional processes leading to particulate carbon export instead of import (Ganju et al. 2015). When channelization occurs, warmer waters can infiltrate deeper into marsh interiors (Enright et al. 2013) initiating a one-way erosive process. Storm activity, through increased discharge, wind waves, storm surges, etc., enhances surface water swells and as such the physical power to liberate surficial and channel deposits (Emmanuel 2005). Accelerated sea-level rise also introduces a more rapid “drowning” or lowering of the marsh surface in the tidal frame, leading to ponding and loss of shear strength, which can enhance sediment mobilization or not (Schoolmaster et al. 2022). Human-induced channels (mosquito ditches) or enhanced transgression

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upstream of the tidal network also enhance channel erosion and leads to interior marsh destabilization and carbon losses (Powell et al. 2020). Still, sea-level rise itself, is not necessarily a destabilizing influence on coastal soils, as it is the primary mechanism by which BCEs have accumulated such large carbon stocks over millennia. Validated geomorphic models, such as the Marsh Equilibrium Model (Morris et al. 2002, 2016) point to carbon stabilization and soil growth responses (Figures 11.1a,b). Paleorecords (e.g., Rogers et al. 2019; Gonneea et al. 2019) document that the rates of carbon accumulation are directly tied to the growth rate of soil, which is ultimately based on changes in accommodation space. Accommodation space at the land–ocean interface is primarily controlled by relative sea-level rise, both the rise of ocean elevation as well as the rise or fall of land elevation, through tectonics, or compaction. Because of the abundance of organic carbon, and biogeomorphic feedback of coastal vegetation, wetlands in equilibrium with the ocean accumulate enough elevation in their soils to fill the accommodation space and thus keep pace with sea-level rise. Soil cores around the world illustrate a wide range of growth rates and carbon densities, but also some consistencies and tendencies as well (Sanderman et al. 2018). It is the acceleration of sea-level rise (and concomitant land surface subsidence) that threatens a tipping point by which the accommodation space is too large to be overcome with equilibrium rates of soil building. Building purely organic soils (peats) is highly dependent on biomass production, and rates are likely to be maximized at 4–5 mm yr–1 of relative sea-level rise (RSLR, Morris et al. 2016). With mineral additions, the rates of accumulation can be enhanced, but only to an unconstrained tipping point. Much

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of the calculations for tipping points are based on relative elevation, and specifically whether a marsh is above or below vegetative optimum (Morris et al. 2002, Figure 11.1a). Vegetative optimums are related to both belowground biomass production as well as sediment trapping in the marsh surface, and thus these site-specific elevation: biomass relationships are the best predictors of marsh sustainability in the face of accelerated sea-level rise. These relationships may be dynamic, however, as nutrient enrichment and other stressors alter biomass distribution and their responses to RSLR (Deegan et al. 2012). Mangrove assessments have also pointed to a likely maximum of soil building, at 6–7 mm yr–1 of RSLR (Saintalan et al. 2020), but note that tidal wetland accretion rates can be as much as 2–4 times higher even with a lack of external sediment supply (e.g., Eagle et al. 2022). While not definitive, these metadata assessments, based on physical principles, are the most compelling quantification of BCE vulnerability under predictions of rapidly rising seas (Sweet et al. 2022), as loss is difficult to detect and often offset by gains (Murray et al. 2022). Potential responses to maintaining elevation are typically based on enhancing mineral supplies, such as through restoring river connectivity (Wasson et al. 2019) or thin-layer sediment addition (Sloane et al. 2021), to bolster soil building processes for sustained elevations and functions.

11.3 POTENTIAL SHIFTS IN BLUE CARBON ACCOUNTING: NET RADIATIVE BALANCE OF GHG EMISSIONS Part of preserving deep soil carbon is preserving the processes responsible for carbon sequestration, and the net negative emissions of this carbon storage function. The rates of

FIGURE 11.1  a,b. Lessons in tidal wetland organic soil accretion and C stabilization from Marsh Equilibrium Model sensitivity testing across a wide range of salinities and climates. Figure (a) produced by Monica Moritsch (USGS) from Morris et al. (2002) for plant and elevation constraints on future tipping points. Figure (b) soil carbon densities downcore and across the conterminous USA (Morris et al. 2016).

Blue Carbon in a Changing Climate and a Changing Context

net negative emissions are calculated as a Net Ecosystem Carbon Balance (NECB), and require boundaries in time and space. Czapla et al. (2020) illustrate the multiple fluxes necessary to calculate a budget for carbon exchanges between neighboring ecosystems in tidal settings. NECB = −NEE + FCO + FCH4 + FVOC + FDIC +FDOC +FPC(11.1) where NEE = net ecosystem exchange of C–CO2; FCO = net carbon monoxide (C–CO) absorption; FCH4 = net C–CH4 consumption; FVOC = net volatile organic C (C–VOC) absorption; FDIC = net dissolved inorganic C (C–DIC) to the ecosystem; FDOC = net dissolved organic C (C–DOC) input; and FPC = net lateral transfer of particulate (non-dissolved, nongaseous) C into the ecosystem. Carbon fixation fluxes are monitored by models of photosynthesis and/or gross primary productivity (GPP), and because they rely on chlorophyll and photosynthetically active radiation, GPP fluxes are among the most amenable to remotely sensed models (e.g., Feagin et al. 2020). In contrast, plant respiration, ecosystem respiration (Reco) and lateral exchange of organic and inorganic carbon are fairly insensitive to remotely sensed inputs (but see Joshi et al. 2017; Sanwlani et al. 2022) and require extensive monitoring to support budget calculations. Lateral fluxes are especially difficult in tidal systems, as the net exchange over a given time period (e.g., 1 year) involves a very small signal-to-noise of many flux measurements then summed to represent net exchanges in a bi-directional and subsurface flow environment (Bogard et al. 2020). Where monitored,

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tidal wetlands (marshes and mangroves) have been shown to export a much larger fraction of their net ecosystem productivity (>40%) to adjacent open waters than inland wetlands (e.g., Figure 11.2a,b). Alkalinity and carbonatesystem metrics, where measured, illustrate that the majority of exported DIC from anaerobic tidal soils is in the form of bicarbonate, and resistant to dissolution and emission, with some fraction entering the long-term deep ocean carbon sink (Santos et al. 2021). Alkalinity export can thus serve as both a sink for carbon and mitigation of ocean acidification in localized settings (Chu et al. 2018). Another anaerobic microbial process in tidal soils with climate impacts is methanogenesis. Methane (CH4) emissions – the fraction of methane produced that is emitted to the atmosphere prior to oxidation – is especially difficult to quantify, requiring long-term high-frequency measurements for accurate accounting. In a climate mitigation context, CH4 emissions are the least confident component of wetland accounting (Holmquist et al. 2018a; Saunois et al. 2020), owing to their outsized influence at short time scales (e.g., 45-fold more impactful than CO2 emissions over a 100-year period). In the absence of sufficient CH4 flux measurements, statistical models are used to estimate relative methane emissions on an annual basis. The most powerful predictive relationship available is the inhibition of methane emissions by porewater sulfate availability, as indexed by salinity (Poffenbarger et al. 2011). Sulfate-reducing bacteria outcompete methanogens for key substrates in most comparative studies. Synthetic studies show that negligible values (close to zero) of CH4 emissions are typically observed in saline ecosystems, yet there is a wide range of responses in more oligohaline environments (0–5 ppt salinity), often

FIGURE 11.2  a,b. Lateral flux exports of Net Ecosystem Production from tidal wetlands compared with other inland ecosystems. Figure reprinted from Bogard et al. 2020.

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exceeding inland freshwater environments (WindhamMyers et al. 2018). The highly varied responses among coastal settings and the mix of biological and physical controls on atmospheric flux rates limit our ability to extrapolate spatially and temporally (Rosentreter et al. 2021). However, experiments with elevated temperatures suggest strong feedback of temperatures supporting methanogenic productivity (Noyce et al. 2023), likely through increased supplies of organic substrate, and thus less competition for food. Leveraging atmospheric and hydrologic flux measurements is not only useful for closing ecosystem carbon budgets but also for assessing radiative balance through a “normalized” metric of carbon dioxide equivalents (CO2eq). Radiative balance (W m–2) can be calculated simplistically, at steady state assumptions, to illustrate the timepoint at which an ecosystem shifts from positive to negative emissions for climate mitigation. Radiative balance also can be used to compare systems, and management implications, whereby a more negative value (even if it is not net negative) is associated with a reduced climate impact (e.g., Neubauer 2021). While simple calculations of CO2 equivalents are useful to evaluate the net climate impacts of an action, a radiative balance approach is more relevant to decision-making. Two important considerations of manipulations on radiative balance are: (1) the relative effect of a management action; and (2) the time frame of the impact of a management action. First, manipulation does not have to achieve net negative C fluxes in order to produce a cooling effect; any change that lowers the net radiative balance is climate mitigation. Second, all manipulations should be considered across a time frame, and thus a cumulative metric of radiation balance; for example, a highly CH4 emitting freshwater wetland when considered over a millennial horizon is net neutral or potentially net negative considering the legacy CO2 uptake being preserved. A key consideration for evaluating radiative balance is “crossover time.” Figure 11.3a, which depicts modeled data from a well-studied freshwater tidal marsh (Virginia, USA,

Climate Change and Estuaries

Neubauer 2021), illustrates how cumulative radiative forcing can be linearly positive with time for CH4 and exponentially negative with time for CO2, resulting in a “crossover time” of, in this case, 850 years, at which point the system is in “radiative balance” before becoming net negative. The purpose of calculating crossover times is to illustrate the importance of timescales in cumulative radiative forcing, as emissions can be positive when a wetland is recently established or disturbed (e.g., Cameron et al. 2021) and net CH4 emissions are high, but may eventually become negative as the net carbon stored overwhelms CH4 emissions. A good range of examples for modeling radiative balance occurs in the San Francisco Bay-Delta, where wetland restoration through impoundment on subsided soils leads to extremely high methane emissions (Hemes et al. 2018) and may not recover the net cooling effect of the historical tidal wetlands, at least over short timescales. Figure 11.3b illustrates the radiative balance for three restoration sites analyzed by Arias-Ortiz et al. (2021). The sites vary in their crossover times, with the tidal restoration site (Eden Landing, south San Francisco Estuary) continuously negative due to tidal flushing and sedimentation, and the two non-tidal wetland restorations (Twitchell Island and Sherman Island) needing either 200 or 400 years (respectively) to reach a net negative emission for climate mitigation. Even so, compared to baseline conditions dominated by oxidation of organic soils, highly managed impoundments in the Delta for carbon mitigation exhibit a substantial improvement in the radiative balance (Hemes et al. 2019; Deverel et al. 2020).

11.4 POTENTIAL SHIFTS IN BLUE CARBON ACCOUNTING: MAP EXTENT AND CHARACTERISTICS Beyond the temporal dynamics of gas and stock flux accounting are the changing maps and wetland characteristics over time. Poised at the land–ocean interface, BCEs represent highly dynamic landforms that manifest processes allowing organic matter accumulation through productivity,

FIGURE 11.3  a,b. Cumulative radiative balance and crossover times from net warming to net cooling impacts. (a) Radiative balance assessed by summed cumulative radiative forcing (CH4 and CO2) for freshwater tidal marsh in Virginia, reprinted from Neubauer 2021. The blue circle represents the crossover point for a freshwater marsh in Virginia, where emission balances reach net zero after 850 years. (b) Radiative balance plotted as crossover curves by Arias-Ortiz et al (2021) for tidal (blue) and nontidal (red and black) restoration sites in California’s Sacramento–San Joaquin Delta. The tidal wetland radiative balance remains net negative.

Blue Carbon in a Changing Climate and a Changing Context

organic carbon trapping, and inhibition of microbial activity (Hopkinson et al. 2012). Their locations globally are physically determined, and yet also culturally constrained as more than 40% have been lost to human development (Barbier et al. 2011). Similarly, many acres of current tidal wetlands are located where human development enhanced sediment supplies to the coast, leading to the progression of marsh facies into openwater environments (Weston 2014). High-energy coastal environments are not conducive to organic matter accumulation and yet may produce algal carbon that is sequestered in sediments far from their site of production (Krause-Jensen and Duarte 2016). Further, roughly one-third of U.S. tidal wetlands are estimated to be impounded (Kroeger et al. 2017) or drained (Crooks et al. 2018), driving enhanced anthropogenic methane and CO2 emissions, and providing key opportunities for mitigation through the restoration of tidal flows. As conditions change along coasts – including losses of barrier islands (Ganju 2019), loss of water clarity in seagrass beds (Sherwood et al. 2019), and intrusion of saltwater into freshwater-dominated habitats (Charles et al. 2019), BCEs are threatened by conditions that do not allow continued C accumulation and preservation, such as coastal squeeze (e.g., Weis et al. 2021). On the other hand, as sea-levels rise, upland transgressions may continue as they have for millennia (Kirwan et al. 2016), generating expanded tidal wetland acreage and accretion of soil carbon in step with accommodation space. Deep learning approaches to mapping tidal wetlands have increased their representation, and better assessed regionally their rates of gains and losses. By including fresh and brackish tidal marshes in previous assessments of salt marshes, Murray et al. (2022) provide the most confident estimates of tidal wetlands area, with mangrove cover holding at roughly 137K km2, but tidal wetlands increasing from 54K to 90K km2 globally. Seagrass estimates still dominate BCE cover, with median estimates of 350K km2 (UNEP-WCMC, Short 2017), but with large uncertainty in the density and resilience of their carbon pools (e.g., Williamson and Gattuso 2022). Spatially explicit maps that constrain distributions and categories (e.g., tidal vs. impounded, saline vs fresh) are the most useful for quantifying past, current, and future climate mitigation benefits. Vulnerability to loss is especially useful, such as by considering the elevation capital of BCEs and any surrounding infrastructure that alters their equilibrium conditions (e.g. impoundments; Yang et al. 2022). Barriers to migration, whether human or natural topography, also constrain future distributions and thus potential continued climate mitigation benefits (Osland et al. 2022). Estimated map distributions for BCEs show global distributions of seagrass at all latitudes, with marshes concentrated in temperate latitudes, and mangroves in tropical latitudes (Howard et al. 2017). These different shorelines are all associated with different rates of CO2 sequestration, and historical carbon stocks, and yet some basic principles of soil accumulation and stabilization occur in all settings (e.g., Morris et al. 2016). Rates of methane emissions, nitrous oxide emissions, and the fate of carbon lost through lateral fluxes are the most difficult

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rates to apply in a spatially explicit manner (Rosentreter et al. 2021). Vulnerability assessments point to sea-level rise as a threat multiplier in sites with low potential for migration and relatively little elevation capital (Holmquist et al. 2021). In addition to better characterization of wetland status (salinity, hydrology, geomorphic setting), projections can be improved by assessing maps across management scenarios in four dimensions, by adding a timescale to resilience for climate mitigation and adaptation benefits. Temporally, we can consider their resilience both as recovery from “pulse” disturbances, or reorganization in response to “press” disturbances. Because climate change and changing contexts set the intensity and duration of stressful conditions across our varied and impacted coastlines, predictions are complicated and cross-disciplinary models will be needed to quantify the magnitude of impacts, whether biological, chemical, or physical.

11.5 ADVANCES IN PREDICTING CLIMATE EFFECTS ON BLUE CARBON STOCKS AND FLUXES It is widely accepted in coastal science literature that sea changes are coming to coastal ecosystems. In my current library of 2,844 coastal-affiliated publications, I have found no published analyses predicting zero effects on coastal settings from climate changes in the next century. Rather, physical and social scientists see active changes in populations, activities, and structure on global coastlines, even in far northern unpopulated regions (Williams et al. 2022). Many advances are occurring through a growing community of practice on Blue Carbon science, across scientific and social disciplines and management domains. Linking societal drivers with natural forces is increasingly complicated, but compounded impacts are reasonably projected with improved models, targeted measurements, and adequately classified maps. Management can be improved by greater attention to potential future scenarios, as tested through model forecasts. Managers also play a key role in calibrating expectations with hindcasts and sensitivity testing (e.g., Sherwood et al. 2019), and thus they provide necessary feedback to improved and actionable models, measurements, and maps.

11.5.1 Toward Better Maps Multiple global efforts have begun sharing information and approaches that optimize current knowledge on BCE classification needs. For example, the Coastal Carbon Research Coordination Network (CCRCN) has served as a receiving and sharing library of raw datasets on stock and flux assessments (Coastal Carbon Atlas), bringing coastal wetlands into the age of “big data” (e.g., Todd-Brown et al. 2022). Gridded data have been created globally using point measurements, machine learning, and remote sensing techniques that are both optical and physical (e.g., Sanderman et al. 2018).

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Whereas coastal land classifications have improved (Sayre et al. 2021; Murray et al. 2022), validated measurements remain underrepresented for many of the disaggregated classes now mapped. Further, some important and potentially “gamechanging” metrics, such as alkalinity export (Maher et al. 2018) or nitrous oxide emissions (Rosentreter et al. 2021), which are both strongly affected by microbial respiratory pathways, are currently unmappable.

11.5.2 Toward Better Models In all cases, understanding the modelable processes behind carbon storage, GHG emissions, and extent can help make stronger predictions and prioritizations as to what the impacts are of human and climatic disturbances on blue carbon metrics. Conceptual models (e.g., Kirwan and Megonigal 2013) have framed the needs and ranges of information required for decision-making. Today, physically modeled BCE elevation and migration sustainability thresholds, and their aggregated emission factors, are our current closest approximation to assessing the future of BCE carbon benefits. In the near future, however, public access to processing codes and tidal elevation maps (Holmquist and Windham-Myers 2022) will enable process-based integrated models for biogeochemical and geomorphic opportunities to test actions for responding to climate threats given situational contexts (WindhamMyers et al. 2019b). In closing, what we know about “Blue Carbon” we know through models. Carbon fluxes in BCEs are difficult to monitor due to their narrow and tenuous distribution between land and sea. Further, they are difficult to monitor due to the range (e.g. tidal to millennial) of timesteps necessary to reconcile land and sea drivers. Measurements and monitoring datasets are key to confidence in accounting. Hindcasts are validated through soil profiles and forecasts are based on process-based models of fluxes and their sensitivities. These advances provide the ability to assess the sensitivity of coastal carbon fluxes to climate change. The primary drivers of concern – accelerated sea-level rise, coastal storm intensity, and rising temperatures – will have mixed impacts across the wide range of BCE sites globally. Some ecosystems will see enhanced CO2 sequestration; some will see reduced CH4 emissions. With human actions, directly and indirectly, altering land and sea physical processes, BCE dynamics are inherently subject to management. Resist–Adapt–Direct (RADical) approaches (Lynch et al. 2021) to land management are one way to incorporate science into structural decision-making. While communities of practice will continue to improve confidence in BCE climate benefits, no-risk actions today include the preservation of remaining coastal carbon stocks. As climate mitigation tools, these soils represent an inheritance of climate cooling that needs recognition and protection as a means to reduce future CO2 loading of the atmosphere (Temmink et al. 2022). Initiating protection of these BCE services today, inherited from “B.C.E.,” may help mitigate climatic changes and ensuing consequences.

Climate Change and Estuaries

11.6 CONCLUSIONS 1. Blue Carbon Ecosystem soil stocks represent an inheritance of cumulative climate-mitigation impacts, but physical disturbances are capable of destabilizing soil stocks and reversing radiative balances toward positive emissions. 2. Warming climates have varied and counteractive effects on net carbon balances. For example, the stimulation of microbial methanogenesis by higher temperatures is also balanced by the inhibition of methanogenesis due to marine-based sulfate intrusion. 3. Today’s maps and models of Blue Carbon Ecosystems indicate their unappreciated extent and range of influence on global C fluxes, with varied conditions that likely mediate net greenhouse gas (GHG) fluxes and ecosystem resilience.

ACKNOWLEDGMENTS I acknowledge the support of the U.S. Geological Survey Water Resources Mission Area, and colleagues within the Coastal Carbon Research Coordination Network (Coastal Carbon Research Coordination Network | Smithsonian Environmental Research Center (si​.e​du)). The manuscript was improved by suggestions made in reviews by Kevin Kroeger (U.S. Geological Survey), Hans Paerl, Joey Crosswell, and Michael Kennish.

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214 Sloane, E. B., K. M. Thorne, C. R. Whitcraft, and V. Touchstone. 2021. Enhancing marsh elevation using sediment augmentation: A case study from southern California, USA. Shore Beach 89(4). https://doi​.org​/10​.34237​/1008943. Small, C. and R. J. Nicholls. 2003. A global analysis of human settlement in coastal zones. J. Coast. Res. 19(3): 584–599. Smith, A. J., G. L. Noyce, J. P. Megonigal, G. R. Guntenspergen, and M. L. Kirwan. 2022. Temperature optimum for marsh resilience and carbon accumulation revealed in a whole‐ ecosystem warming experiment. Glob. Change Biol. 28(10): 3236–3245. Spivak, A. C., J. Sanderman, J. L. Bowen, E. A. Canuel, and C. S. Hopkinson. 2019. Global-change controls on soil-carbon accumulation and loss in coastal vegetated ecosystems. Nat. Geosci. 12(9): 685–692. Sutton-Grier, A. E. and A. Moore. 2016. Leveraging carbon services of coastal ecosystems for habitat protection and restoration. Coast. Manag. 44(3): 259–277. Sweet, W. V., B. D. Hamlington, R. E. Kopp, C. P. Weaver, P. L. Barnard, D. Bekaert, et al. 2022. Global and regional sea level rise scenarios for the United States: Up-dated mean projections and extreme water level probabilities along U.S. coastlines. NOAA technical report NOS 01. Silver Spring, MD: National Oceanic and Atmospheric Administration, National OceanService. https://oceanservice​.noaa​.gov​/ hazards​/sealevelrise​/noaa​-nos​-techrpt01​-global​-regional​-SLR​ -scenarios​-US​.pdf. Temmink, R. J., L. P. Lamers, C. Angelini, T. J. Bouma, C. Fritz, J. van de Koppel, et al. 2022. Recovering wetland biogeomorphic feedbacks to restore the world’s biotic carbon hotspots. Science 376(6593): eabn1479. Todd-Brown, K. E., R. Z. Abramoff, J. Beem-Miller, H. K. Blair, S. Earl, K. J. Frederick, et al. 2022. Reviews and syntheses: The promise of big diverse soil data, moving current practices towards future potential. Biogeosciences 19(14): 3505–3522. Toscano, M. A., J. L. Gonzalez, and K. R. Whelan. 2018. Calibrated density profiles of Caribbean mangrove peat sequences from computed tomography for assessment of peat preservation, compaction, and impacts on sea-level reconstructions. Quat. Res. 89(1): 201–222. Troxler, T., J. Tang, and C. Hopkinson. 2018. Coastal blue carbon. Pp. 33–72. In: National academies of sciences, engineering and medicine, developing a research agenda for carbon dioxide removal and reliable sequestration. Washington, DC: The National Academies Press. https://doi​.org​/10​ .17226​/25259. UNEP-WCMC, Short, FT. 2017. Global distribution of seagrasses (version 6.0). Sixth update to the data layer used in Green and Short (2003). Cambridge: UN Environment World Conservation Monitoring Centre. http://data​.unepwcmc​.org​ /datasets​/7. Wang, F., X. Lu, C. J. Sanders, and J. Tang. 2019. Tidal wetland resilience to sea level rise increases their carbon sequestration capacity in United States. Nat. Commun. 10(1): 1–11.

Climate Change and Estuaries Wang, Z. A., K. D. Kroeger, N. K. Ganju, M. E. Gonneea, and S. N. Chu. 2016. Intertidal salt marshes as an important source of inorganic carbon to the coastal ocean. Limnol. Oceanogr. 61(5): 1916–1931. Ward, N. D., J. P. Megonigal, B. Bond-Lamberty, V. L. Bailey, D. Butman, E. A. Canuel, et al. 2020. Representing the function and sensitivity of coastal interfaces in Earth system models. Nat. Commun. 11(1): 1–14. Wasson, K., N. K. Ganju, Z. Defne, C. Endris, T. Elsey-Quirk, K. M. Thorne, et al. 2019. Understanding tidal marsh trajectories: Evaluation of multiple indicators of marsh persistence. Environ. Res. Lett. 14(12): 124073. Weis, J. S., E. B. Watson, B. Ravit, C. Harman, and M. Yepsen. 2021. The status and future of tidal marshes in New Jersey faced with sea level rise. Anthropocene Coasts 4(1): 168–192. Weston, N. B. 2014. Declining sediments and rising seas: An unfortunate convergence for tidal wetlands. Estuar. Coasts 37(1): 1–23. Williams, B. A., J. E. Watson, H. L. Beyer, C. J. Klein, J. Montgomery, R. K. Runting, et al. 2022. Global rarity of intact coastal regions. Conserv. Biol. 36: e13874. https://doi​ .org​/10​.1111​/cobi​.13874. Williamson, P. and J.-P. Gattuso. 2022. Carbon removal using coastal blue carbon ecosystems is uncertain and unreliable, with questionable climatic cost-effectiveness. Front. Clim. 4. https://doi​.org​/10​.3389​/fclim​.2022​.853666. Windham-Myers, L., W.-J. Cai, S. R. Alin, A. Andersson, J. Crosswell, K. H. Dunton, et al. 2018. Tidal wetlands and estuaries. Pp. 507–567. In: N. Cavallaro, G. Shrestha, R. Birdsey, A. M. Mayes, R. G. Najjar, S. C. Reed, et al. (eds.), Second state of the carbon cycle report (SOCCR2): A sustained assessment report. Washington, DC: US Global Change Research Program. https://doi​.org​/10​.7930​ /SOCCR2​.2018​.Ch15. Windham-Myers, L., S. Crooks, and T. G. Troxler (eds.). 2019a. A blue carbon primer: The state of coastal wetland carbon science, practice, and policy. Boca Raton, FL: CRC Press, Taylor and Francis. Windham-Myers, L., S. Crooks, and T. Troxler 2019b. Blue carbon futures: Moving forward on terra firma. Pp. 391–401. In: L. Windham-Myers, S. Crooks, and T. G. Troxler (eds.), A blue carbon primer: The state of coastal wetland carbon science, practice, and policy. Boca Raton, FL: CRC Press, Taylor and Francis. Yang, X., Z. Zhu, S. Qiu, K. D. Kroeger, Z. Zhu, and S. Covington. 2022. Detection and characterization of coastal tidal wetland change in the northeastern US using Landsat time series. Rem. Sens. Environ. 276. https://doi​.org​/10​.1016​/j​ .rse​.2022​.113047. Zabarte-Maeztu, I., F. E. Matheson, M. Manley-Harris, R. J. Davies-Colley, and I. Hawes. 2021. Fine sediment effects on seagrasses: A global review, quantitative synthesis and multi-stressor model. Mar. Environ. Res. 171: 105480.

12

Effects of a Changing Climate on the Physics of Estuaries L. Fernando Pareja-Roman and Robert J. Chant

CONTENTS Abstract............................................................................................................................................................................... 215 12.1 Introduction............................................................................................................................................................... 215 12.2 Estuarine Tides and Climate Change........................................................................................................................ 216 12.2.1 Tides in Idealized Estuaries without Friction................................................................................................ 216 12.2.2 Tides in Idealized Estuaries with Friction..................................................................................................... 217 12.2.3 Tides in Realistic Estuaries........................................................................................................................... 219 12.2.4 Coastal Lagoons and Sea-Level Rise............................................................................................................ 220 12.3 Estuarine Circulation, Dispersion, and Climate Change........................................................................................... 221 12.3.1 Tidal Dispersion (Kt )..................................................................................................................................... 222 12.3.2 Steady Shear Dispersion (Kex)........................................................................................................................ 223 12.3.3 Fjord Circulation............................................................................................................................................ 225 12.4 Synopsis: Climate Change Effects and Ecological Implications.............................................................................. 225 12.5 Conclusions................................................................................................................................................................ 226 References........................................................................................................................................................................... 227

ABSTRACT Estuaries are coastal bodies of water known for their biological productivity and the numerous ecosystem services they provide to humans. Over the last century, many estuaries worldwide have seen unprecedented stress and changes in their physical and ecological dynamics, especially those adjacent to urbanized centers and ports. The combination of climate change (mainly shifts in watershed dynamics, river discharge, tides, sea-level rise, and increasing temperatures) and anthropogenic interventions (e.g., urbanization and infrastructure development) will have a significant impact on the coupled physical and ecological dynamics of these coastal systems. In this chapter, we focus on the potential impact of climate change on the physics (i.e., hydrodynamics) of estuaries. We use simple, intuitive mathematical models to explore the ways sea-level rise and changes in river discharge may alter tides and density-driven circulation. First, we consider simple cases such as prismatic estuaries and increase the level of complexity by discussing coastal lagoons and fjords. For tides, the impact of sea-level rise and changes in river discharge depends largely on the type of estuary (short or long, shallow or deep, inlet-choked or unimpeded), as well as the nature of the shorelines. For salt intrusion and density-driven dynamics, we explore the drivers of turbulent dispersion in estuaries and some of the aspects that control residence time and flushing. We then include a discussion on potential ecological implications. One of the main takeaways is that rather than a universal estuarine response to climate change, the expected impacts DOI: 10.1201/9781003126096-13

tend to vary significantly based on the type of estuary and the projected anthropogenic interventions. Key Words:  Estuarine tides, exchange flows, tidal and shear dispersion, sea-level rise, climate change

12.1 INTRODUCTION Estuarine physics refers to the study of tides and densitydriven flows in estuaries from a hydrodynamic point of view. Tides are oscillatory fluctuations in sea-level and velocity caused by planetary-scale gravitational forces acting on the ocean. Estuarine (e.g., density-driven) circulation refers to the time-averaged exchange between fresh and saltwater, with river discharge as the main source of buoyancy. This circulation is often conceptualized as a two-layer flow whereby a seaward layer of freshwater lies on top of a landward layer of denser, saltier water. These mean flows are often spatially variable and can be sensitive to changes in tides, buoyancy forcing, and winds. In this chapter, we discuss the potential role of a changing climate on tides and estuarine circulation. For both topics, we revisit seminal theories known for their conceptual clarity; for example, the work on tides and hydraulic engineering by Ippen and Harleman (1966) and Dronkers (1964), and the long-standing papers on density-driven flows by Pritchard (1956) and Hansen and Rattray (1966). While we omit mathematical derivations, we present simple algebraic terms that show the dependency of tidal attenuation, propagation, and mixing on key variables such as depth, velocity, ocean salinity, 215

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estuary length, and width. An advantage of this approach is that it facilitates an intuitive understanding of how estuarine hydrodynamics may adjust to a changing climate. Undoubtedly, climate change involves multiple feedbacks among the earth’s physical components, and its effects are geographically variable. In this review, we omit geographical and temporal complexities and only consider the potential effect of sea-level rise (SLR) and shifts in river discharge from a highly idealized perspective. In addition to discussing climate effects, we bring the reader’s attention to concurrent issues that affect many estuaries worldwide today. For example, anthropogenic activities related to land use, water diversion, shoreline modification, and dredging often have a larger impact on estuarine tides and circulation than climate change. We also briefly discuss the importance of the morphological adjustment of estuaries to climate change and how algebraic predictions overlook this critical factor. More detailed predictions on how estuaries will evolve physically over the next century require numerical models that account for morphological evolution from hourly to decadal time scales. This chapter is organized into several sections. Section 12.2 focuses on estuarine tides and their response to increasing depth under hypothetical SLR scenarios. We consider prototypical channel-like estuaries with and without friction, as well as an idealized lagoon that is connected to the ocean by a single inlet. Section 12.3 is about density-driven dynamics: salt fluxes, dispersion, exchange flows, and salt intrusion. Section 12.4 presents a synopsis and some potential ecological implications of climate change in the physics of estuaries. We then present a summary that collects the key points of each part.

12.2 ESTUARINE TIDES AND CLIMATE CHANGE Gravitational forces within the sun–earth–moon system produce sea-level fluctuations in the ocean known as tides. These oscillations typically occur at a given location in the ocean once or twice a day, which is why we refer to them as semidiurnal or diurnal tides. In some regions, such as the USA east coast, semidiurnal oscillations dominate the temporal sea-level variability. In others, we may observe predominately diurnal or a combination of diurnal and semidiurnal tides in gauge records. Geographical differences in tidal patterns are due to the relative location of the continents and the distribution of rotation centers known as amphidromic points. Tides play a vital role not only in the gradation of habitat in coastal regions but also in the provision of ecosystem services for humans, especially in estuaries where the anthropogenic imprint has become conspicuous over the last decades. From a dynamical standpoint, we can describe tides as long, surface gravity waves that feature crests (high water) and troughs (low water), and whose propagation pattern is affected by the earth’s rotation. The Coriolis force traps the tide against the coastline such that the propagation occurs with the land on the righthand side in the northern hemisphere and on the left-hand

Climate Change and Estuaries

side below the equator – a characteristic mechanism of Kelvin waves (Pugh 1987). Along continental margins, part of the tidal energy is diverted into estuaries, bays, and lagoons where the difference between high and low water, or the tidal range, often differs from offshore values. Tidal transformations within estuaries can be attributed to a suite of hydrodynamic processes, but mainly due to the way in which the incoming tide interacts with the local topography (de Miranda et al. 2017). Here, “topography” encompasses a range of basin features such as bathymetry (e.g., whether the estuary has a flat bottom or a channel-shoal configuration), channel width, seabed roughness, and the characteristics of the basin’s boundaries (e.g., is the estuary bounded by wetlands or hardened shorelines?). In the Anthropocene, an era in which the changing climate is marked by urban development, agricultural growth, rising sea-levels, and changes in topography, tides have changed in many estuaries of the world (Khojasteh et al. 2021; Talke and Jay 2020). Examples include the Hudson (Ralston et al. 2019), Ems (de Jonge et al. 2014b), and Cape Fear River estuaries (Familkhalili and Talke 2016), the Delaware Bay and San Francisco Bay (Holleman and Stacey 2014; Lee et al. 2017), Boston Harbor (Talke et al. 2018), and the Changjiang Estuary in China (Li et al. 2020a). Understanding the role of anthropogenic modifications, climate change, and SLR on tidal shifts is critical for water resource management, resiliency building, and climate change adaptation (Talke and Jay 2017). Next, we review the fundamentals of tidal dynamics to examine how estuaries may respond to a changing climate, especially rising sea-levels. For simplicity, we will ignore non-tidal flows to focus on tidal variability. We will consider idealized channel-type estuaries and then briefly discuss the potential effect of SLR in coastal lagoons.

12.2.1 Tides in Idealized Estuaries without Friction To understand why and how topography can alter the tidal range in estuaries, it is useful to revisit conservation laws from physics and use some of the resulting algebraic expressions as guides. Conservation of mass implies that the timeaveraged amount of water that flows through any cross section must always remain the same. Thus, if the cross-sectional area of an estuary decreases in the landward direction, mass conservation requires that the velocity or water elevation change accordingly so that the amount of flowing water per unit of time does not change. The narrowing of the area, which some authors call “topographic funneling” (Jay 1991), explains why the tidal range increases in the landward limit of some convergent estuaries, such as the Delaware, Thames, and Tamar estuaries (Friedrichs and Aubrey 1994). At the same time, friction with the seabed and shorelines counteracts the funneling effect and tends to reduce the tidal range. Intuitively, whether the tidal range increases or not along a segment of an estuary depends on the competition between friction and convergence. A useful classification arises from these basic tidal responses: if convergence outcompetes

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friction, the estuary is “hypersynchronous,” and it is “hyposynchronous” if the opposite is true (de Miranda et al. 2017). In the relatively rare case where convergence and friction have equal magnitude, the estuary is in “near equilibrium” (Friedrichs 2010) or “synchronous.” At this point, we can anticipate that the main role of SLR will be to change the extent of tidal “synchronicity” in estuaries. In a discussion of estuarine geometry and tides, we should mention that funneling is not necessary for the tidal range to increase along the channel. In relatively rectangular or weakly convergent segments, amplification can also occur if the incoming tide is sufficiently reflected at the end of the channel, which we call the “head.” This is typical in estuaries where the head is defined by a hydraulic structure such as a weir or a sharp geological feature – see, for example, a discussion on tidal reflection at a weir in the Guadalquivir River estuary (Díez‐Minguito et al. 2012) and at the head of the Hudson River estuary (Georgas 2012). In summary, convergence and reflection favor the amplification of tidal range, while friction tends to reduce it. The main role of SLR, as we will see in more detail shortly, is to alter these competing factors by increasing depth. As mentioned earlier, the tide is essentially a long gravity wave. As such, it has a period (T ), phase speed ( gh ), and wavelength l = T gh , where g is the gravitational acceleration and h is depth. The tidal frequency is w = 2p / T . These definitions arise from the derivation of the wave equation (Ippen and Harleman 1966). At any given time, the tide is the combination of semidiurnal (M2, S2, N2), diurnal (O1, K1), and other minor constituents with different amplitudes and time lags. For simplicity, the discussion hereafter will focus on a single astronomical constituent. At first glance, one of the effects of SLR would be an increase in phase speed gh , and subsequently in l . As the phase speed increases, we can also expect that high water will occur sooner along the estuary. We may observe that the wavelength grows as well, but the implications of increased wavelength on tidal range depend on the ratio between l and the length of the estuary, L, which we assume is constant. To explore the dynamical implications of increasing l , we need to apply Newton’s second law, which states that the sum of forces acting on a parcel of fluid is proportional to its acceleration. In this case, the acceleration is caused by the pressure gradient force induced by changes in tidal elevation. If we schematize an estuary as a rectangular channel with no river discharge at its head and no friction, we can find an expression for the tidal amplitude (h ), or half the difference between high and low water:

words, the amplification is maximized as cos kL tends to zero or when l tends to 4L. This phenomenon is known as quarter-wave resonance and has been observed in the Gulf of Thailand, Long Island Sound in the USA, and the Bay of Fundy in Canada, among others. Following our earlier assumption that SLR changes l and not L, we may also see that the role of increasing depth on resonance depends on the starting value of L / l . If L / l is currently greater than one-quarter in an example estuary, SLR will increase this ratio, and the system will move closer to resonance and the tidal range will increase. If the initial L / l ratio is smaller than one-quarter, SLR will make the ratio even smaller. The tidal response then will move away from resonance, and the tidal range will decrease. The effect of increasing l on amplification is described in detail by Talke and Jay (2020) who show some estuaries and gulfs of the world in terms of their estimated L / l ratios and their tidal amplification relative to the ocean. The authors illustrate the role of dredging, which is analogous to SLR in that it involves increased channel depth, on the resonant response of the Ems estuary (the Netherlands). Given the importance of L / l on estuarine tidal dynamics, an estuary is often considered “short” or “long” not in terms of its actual length L but on the L / l ratio. Thus, we can say that a short estuary is one where the length is short relative to the tidal wavelength (Li and O'Donnell 2005). The main dynamic role of SLR is then to “shorten” the estuary.



h = h0

cos kx (12.1) cos kL

where x > 0 is the along-estuary direction, h0 is the amplitude at the mouth, and k = 2p / l is the wavenumber. This expression for h contains the interesting result that the amplification relative to the mouth (h / h0 ) increases as the cos kL term in the denominator decreases. In other

12.2.2 Tides in Idealized Estuaries with Friction Previously, we considered an idealized case in which the tide propagates into a channel with no friction. Next, we consider a more realistic scenario in which we account for friction, as is typical in the relatively shallow coastal plain or alluvial estuaries. Specifically, we now assume that friction attenuates the tidal amplitude at an exponential decay rate m , which we call the attenuation coefficient. With this new assumption, the momentum and continuity equations can be combined to yield



æ ç h = h0 ç ç è

1

1 ö2 (cos 2kx + cosh 2 m x ) ÷ 2 ÷ (12.2) 1 (cos 2kL + cosh 2 m L ) ÷ 2 ø

through some algebra, we can show that the attenuation coefficient is inversely proportional to the phase speed; m = r / 2 gh , where r = ( 8 / 3p ) C DU / h is a friction coefficient that involves the tidal current amplitude U ; and C D , a coefficient that is proportional to the seabed roughness. By replacing r in m , we find that the attenuation coefficient is proportional to h -3 / 2

æ 4C U ö - 3 m = ç D ÷ h 2 (12.3) ç 3p g ÷ è ø

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The expression above indicates that increasing depth decreases the damping effect. A more accurate derivation in which we allow the tidal current amplitude to change with depth, U = (h / h ) gh , shows that m is proportional to h -2 . Figure 12.1 illustrates how m decreases with h according to Equation (12.3), and Figure 12.2 presents examples for the along-channel tidal response to different SLR scenarios based on Equation (12.2). In a hypothetical SLR case where the initial depth h0 increases by Δh, the proportional change in the attenuation coefficient is Δμ µ (1 / h 0) Δh / h 0. This suggests that the damping response is more significant in relatively shallow estuaries where Δh / h 0 is large, or where Δh due to SLR is an important fraction of the initial depth. Talke et al. (2021) show that large values of Δh / h due to dredging resulted in the doubling of the tidal range in the Cape Fear River estuary and also highlight that the tidal and flooding response to Δh in weakly convergent estuaries can vary spatially. FIGURE 12.1  Attenuation coefficient ( m ) as a function of channel depth, Equation (12.3) for different current amplitude scenarios (U = 0.5 m/s; 1.0 m/s) and C D = 2 × 10 -3 .

FIGURE 12.2  Tidal amplitude in idealized estuaries with friction under SLR scenarios from 0 m to 1 m in 0.2 m increments. The estuary lengths are (a) 50, (b) 80, (c) 100, and (d) and 120 km. Results are based on U = 0.5 m/s and C D = 2 × 10 -3 . The tidal amplitude at the mouth is set to 0.6 m for all cases (dashed line).

Effects of a Changing Climate on the Physics of Estuaries

12.2.3 Tides in Realistic Estuaries Simple mathematical frameworks of prismatic estuaries show that SLR would tend to reduce the damping effect and increase tidal range, especially in shallow systems. In more realistic estuaries, the topography is difficult to schematize mathematically as it may include shallow shoals, shipping channels, intertidal flats, and vegetated substrates such as kelp forests and seagrass meadows. The presence of spatially variable topographic features complicates the analysis of the attenuation coefficient, and more elaborate models are often needed. One of the aspects we neglected in the discussion of friction is the shape of the shorelines, which may be critical for the overall tidal response to SLR. In this context, it is customary to classify shorelines as “hard” and “soft.” Concrete seawalls and bulkheads are examples of “hard shorelines,” and they tend to streamline tidal flows, reduce energy dissipation, and increase tidal reflection. The term “soft shoreline” refers to vegetated riparian sills, marshes, or living shorelines such as constructed oyster reefs. In the algebraic expressions we have used so far, an underlying assumption is that the estuarine shorelines are hard, which means that frictional dissipation is only due to bed roughness. In a more realistic scenario, SLR may increase the mean sea-level to a point where land inundation can occur during a portion of the tidal cycle. New intertidal areas can increase tidal attenuation over shallow, often vegetated flats (Nepf 1999, 2012), which would also alter the duration of ebb and flood. In hardened estuaries, tidal propagation is confined to the channel where currents are stronger during the flood phase (Friedrichs 2010). This is because the depth at high water is greater than at low water, in which case the estuary can be referred to as “flood dominant.” In the case where the tide inundates low-lying areas, the propagation of high water is delayed by the flood expansion over the flats while the tidal prism is quickly evacuated in the deeper channels during ebb. This mechanism produces an asymmetry in the duration of ebb and flood (Speer and Aubrey 1985). Thus, another potential consequence of SLR is a shift in the flood-ebb duration asymmetry. Changes in tidal asymmetry are important for long-term sediment dynamics (Allen et al. 1980; Cheng et al. 2010; Chernetsky et al. 2010; Scully and Friedrichs 2007), which in turn shapes estuarine morphology. Assessing the morphological evolution of estuaries as sea-level rises poses a significant challenge in numerical modeling given the relatively long time scales involved and the suite of hydrodynamic feedbacks that affect sediment erosion and deposition (Burchard et al. 2018). For example, both tidally driven (barotropic) and density-driven (baroclinic) flows can change under SLR, and they both contribute to the morphodynamic evolution of estuarine basins (Olabarrieta et al. 2018). A more accurate representation of morphological responses under SLR would require three-dimensional models that allow bathymetry to adjust to changing sealevels and account for sediment infilling.

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Despite inherent challenges in numerical modeling, there are several examples in the literature that discuss the role of shoreline characteristics on inundation induced by SLR. For example, Lee et al. (2017) conducted numerical experiments to evaluate the response of the Chesapeake and Delaware Bays to SLR under shoreline hardening and softening. In the hardening case, SLR increased the overall depth keeping the initial shoreline configuration, while the softening scenario allowed the inundation of low-lying areas. With hardened shorelines, the authors found that the tidal range increased, especially in the landward, narrow reaches of each estuary. This is consistent with our previous discussion of the reduced attenuation coefficient under increasing depth. In contrast, when low-lying land inundation was allowed, the tidal range decreased because the reduced dissipation in the channel was compensated by the increased friction over shallow water. These results indicate that shoreline softening can help attenuate the tidal range as the sea-level rises and provide a basis to guide climate adaptation efforts. In San Francisco Bay, another study considered several modeled scenarios, among them a “leveed” hardened shoreline case and another that allowed inundation of low-lying flats (Holleman and Stacey 2014). Similar to Lee et al. (2017), these authors reported tidal amplification in the leveed case and attenuation when inundation was allowed. In addition to SLR, tides can also produce a residual (i.e., time-averaged) sea-level tilt in estuaries. In some cases, the timing of the maximum tidal current speed at a certain location nearly coincides with that of high water, in which case we say that the velocity and elevation are in phase and that the tide is progressive. This generally occurs in long estuaries where the tidal phase (i.e., time of high water) advances landward. When this happens, there is a net forward movement of water referred to as the “Stokes drift,” and a sea-level tilt may develop in the landward direction. The residual elevation in this case is then the mean sea-level (including any SLR effects) plus the Stokes-induced tilt. At the head of a long estuary, the sea-level tilt due to the Stokes drift can be on the order of 10 cm. If the maximum current speed and elevation are fully out of phase (i.e., the tide is a standing wave), the Stokes drift becomes negligible. Common to the studies by Lee et al. (2017) and Holleman and Stacey (2014) is a discussion of how SLR alters the phase lag between elevation and velocity. For the Delaware and Chesapeake Bays, Lee et al. (2017) found that shoreline softening reduced the phase lag while hardening increased it. In the San Francisco Bay, changes in phase lag under SLR differed between subareas since the estuary features a complex, convergent topography ringed by numerous tidal sloughs and ponds. While we have examined different estuarine responses to SLR, we have not yet considered the role of river discharge. The river discharge can be an important driver of the tidal and residual dynamics of tidal rivers. At the head of the estuary, the seaward river velocity enhances the ebb and competes against the flood. As a result, there is an

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asymmetry in the river-tide velocities that is often referred to as “river–tide” interaction (Godin 1985; LeBlond 1978). At the head of the estuary, and from a tidally averaged perspective, it is reasonable to assume a balance of forces between the depth-averaged friction and the pressure gradient produced by the river water surface slope (Buschman et al. 2009). In this balance, increasing depth is reflected in a relaxation of the river slope for the same river discharge, similar to the response to dredging (Ralston et al. 2019). The total mean sea-level budget at a given point then consists of contributions from SLR, any Stokes-driven residual tilt, and the river water slope. In certain cases, land subsidence also needs to be factored into the residual elevation budget. With SLR, we can then intuitively expect increased conveyance and reduced flood risk. However, the river discharge can increase in the future due to climate-driven shifts in watershed water cycles. The multicomponent nature of river-tide interactions and its connection with watershed hydrology is a subject of ongoing research. One of the potential consequences of changing estuarine tides and rising sea-levels is the increase in the frequency of high tide floods. These are non-catastrophic, relatively minor floods that disrupt daily activities and can strain urban infrastructure systems in the long term. While these events can occur due to astronomical tides alone in areas such as Boston Harbor (Ray and Foster 2016), studies have shown that SLR and changing tides can increase their frequency. Li et al. (2021) found that SLR is the main driver of increased nuisance flood frequency in the USA, and that long-term changes in tides can also contribute.

Climate Change and Estuaries

FIGURE 12.3  Schematic of a coastal lagoon in the Northern Hemisphere. The main geometric features are the inlet length (Li ), depth (h), surface area ( Ae ), and Mean Sea-Level (MSL) datum. The tidal elevation offshore is h0 . Modified from Spaulding (1994).

effect is also called “impedance” (Mann and Mehta 1993) due to its analogy to impedance in an electrical circuit. The strength of the choking effect can be measured with the “choking number” (P ): 1

12.2.4 Coastal Lagoons and Sea-Level Rise In previous examples, estuaries were simplified as channels where the tide is not locally restricted or attenuated at the mouth. We now briefly discuss the response of coastal lagoons, which are semi-enclosed, coastal bodies of water in which the exchange with the ocean takes place through one or several narrow inlets (Kjerfve and Magill 1989). Lagoonal systems comprise ~10% of the world’s coastlines and nearly 25% of the global estuary area (Dürr et al. 2011). They are common along USA east coast, Gulf Coast, and Pacific northwest coasts (Mulhern et al. 2017). The orientation of coastal lagoons is often parallel to the coast so that they are separated from the ocean by a thin strip of land such as a barrier island or peninsula. Two key features of these lagoons are their restricted exchange with the ocean and their relatively low freshwater input. A key assumption often made to conceptualize coastal lagoon hydrodynamics is that the tide does not propagate as it does in a channel but that the sea-level in the lagoon responds uniformly to ocean forcing. Relevant geometric features of coastal lagoons are shown in the following schematic (Figure 12.3) based on Spaulding (1994). The inlet, which is often shallower than the inner continental shelf, can produce a “choking” effect that locally attenuates the incoming tide (Stigebrandt, 1980). This



æ gB2h3T 2 ö 2 P=ç (12.4) 2 ÷ è C D Lih0 Ae ø

which depends on the inlet width (B), depth (h), and length (Li ); tidal period (T ), drag coefficient (C D ), offshore tidal elevation (h0 ), and lagoon surface area (Ae ). Hill (1994) indicates that the system is choked when P 0.3, which is indicative of a shallow, well-mixed estuary. The role of intertidal flats and marshes is important as well; increasing intertidal area tends to increase ebb-dominance. Therefore, changes in water depth and intertidal storage will have an influence on net sediment transport direction: deepening will favor ebb-dominance while decreasing intertidal area will favor flood-dominance. The reader is directed to Zhou et al. (2018) for a thorough discussion of these simplified tidal asymmetry metrics, and guidance for proper application to realistic estuarine geometries. The analytical solutions described above function as general concepts for relating total sediment load with tidal propagation and estuarine morphology. They provide important guidance for evaluating future changes in sediment transport under climate change, however, other detailed aspects of tidal transport are equally relevant, especially in deeper, stratified estuaries. Vertical and lateral density gradients interact with the flow to induce flood–ebb asymmetries that influence mixing and near-bed stress on the sediment bed, and ultimately control net sediment transport direction (Geyer and MacCready 2014). Tidal pumping (Geyer et al. 2001) is defined as the net sediment transport induced by tidal-timescale correlations between velocity and suspended-sediment concentration (SSC). For example, in a system with no tidal velocity asymmetry (i.e., balanced flood–ebb velocities), increased SSC on flood tide will lead to net landward transport. The underlying cause of the SSC increase will depend on the system. In a spatial context, wind-wave resuspension (see Section 13.3.2.1) can increase SSC in one region of the estuary (Ganju and Schoellhamer 2006), providing a mobilization mechanism that partners with advection from that region to yield a net transport towards elsewhere in the system. In a temporal context, differences in the vertical velocity profile between flood and ebb can yield differences in bed shear stress that result in locally enhanced resuspension on one phase of the tide (Scully and Friedrichs 2007). Lastly, net sediment transport can also arise from durational asymmetries in water level despite a constant settling velocity (settling lag; Chernetsky et al. 2010). Coarser sediment, specifically sands, have settling velocities exceeding 1 cm/s and tend to move along the bed with minimal presence in the upper water column. The net transport direction of this bedload will generally match the direction of tidal dominance as described above, because of the non-linear effect of velocity on sediment transport. In addition to saltation (rolling) of grains along the bed, the larger-scale movement of bedforms can have an important influence on sediment supply and morphology. Ripples

Climate Change and Estuaries

and sand waves form and migrate in response to combined waves and currents, and can play an important role in sediment supply to estuaries (Fitzgerald et al. 1983). Estuarine-suspended sediment dynamics are often best characterized by the estuarine turbidity maximum (ETM). This common feature represents the convergence of nearbed flows and suspended sediment (Burchard et al. 2018), and results in a zone of sediment trapping, enhanced suspended-sediment concentrations in the water column, and the formation of a mobile fine sediment pool. Conceptual geologic models of estuarine evolution (discussed later) cite the ETM as an important aspect of morphological development, chiefly as a depocenter and mobile sediment supply for other parts of the system. Furthermore, the ETM also serves as a source and sink for sediment-bound contaminants, nutrients, and organic detritus. The response of estuarine biogeochemistry and food webs often relates to the spatial and temporal nature of the ETM. Numerous studies have investigated the ETM’s response to tidal asymmetry, river inflow, and sediment parameters (Jay et al. 2007; Talke et al. 2009), and in a general sense, Burchard et al. (2018) differentiate between four general ETM types. First, those created at the salinity intrusion limit form through estuarine circulation driving a near-bed landward sediment transport and termination of the salinity gradient. In this case, the strength of the estuarine circulation is a function of river flow, depth, and tidal mixing. A second type of ETM is formed in the freshwater zone simply due to tidal asymmetry leading to landward transport, that is countered by river discharge, again leading to convergence of currents and suspended-sediment. Topographic transitions are responsible for a third type of ETM, where stable bathymetric features induce local intensification of baroclinic circulation and stratification (due to the rapid change in depth), leading to near-bed convergence of flow and sediment. Lastly, lateral trapping of sediment due to lateral asymmetries in depth and stratification can generate near-bed convergence that is heterogeneous in the cross-estuary direction. Each of these ETMs will be influenced differently by changes in sea-level, river flow, sediment supply, and tidal propagation. The effect of salinity on flocculation, both within and outside of the ETM, is outside of the scope of this chapter, though we can assume this effect will act in concert with the movement of ETMs associated with salinity transitions.

13.2.2 Non-tidal Processes The aforementioned discussion of the ETM involves estuarine circulation (i.e., gravitational circulation). The estuarine circulation is not a tidal process, though there are tidal processes that have an impact on the strength of the circulation and its spatiotemporal evolution (Geyer and MacCready 2014). Chapter 12 covered estuarine circulation in-depth, so here we only point out that the relative strength of the estuarine circulation and its convergence will have an impact on near-bed sediment convergence as well as the export of near-surface sediment to the ocean. The response

Climatic Drivers of Estuarine Sediment Dynamics

of sediment dynamics induced by the estuarine circulation will depend on river flow magnitude, timing, and watershed sediment load, which are also non-tidal processes. The relative “length” of the estuary, inversely related to tidal amplitude and river flow (Lerczak et al. 2009), also plays a role in sediment trapping of river-borne sediment. Rivers are responsible for transporting watershedderived sediment to the estuary and ultimately the ocean. The annual load to an estuary is typically described as a power law (Müller and Förstner 1968), where representative sediment concentration C is predicted as aQb, where Q is the volumetric flow rate, indicating that sediment load (Q × C) is aQb+1. Factors a and b are related to the watershed characteristics (supply), and mean river condition (transport potential), but may also be temporally dynamic within an individual river system (Warrick 2015). Later we will discuss how these parameters may change in a system in response to external forcing associated with climate change. Here it is important to note the non-linear nature of the sediment load in response to river flow, and how that non-linearity influences total load with varying hydrograph shapes. For example, in a weakly non-linear system with b = 0.1 (at the extreme low end of typical values), compressing an entire year’s flow into one day will nearly double the total sediment load (in comparison to spreading the flow evenly throughout the year). Most rivers, even those which are heavily regulated, have some characteristic hydrograph shape. This leads to higher sediment loads during the “wet” season, and lower loads in the “dry” season. These transitions affect the seasonal net sediment transport magnitudes and directions within the estuary. Watersheds and climatic zones with a pronounced seasonal hydrograph, for example, those feeding the Hudson, Sacramento–San Joaquin, Mekong, and Amazon River estuaries, will deliver most of their annual sediment load to the estuary in a few events (Geyer et al. 2001; Walling 2008; Schoellhamer 2011). The timing of these events relative to other seasonal forcings is important in dictating the ultimate fate of the delivered sediment. The size of the event also dictates whether the sediment is exported directly to the ocean/nearshore environment, or trapped in seaward reaches of the estuary, as does the estuarine length. Longer estuaries may trap all sediments during the largest events, while shorter estuaries allow for most sediment to bypass the estuary and reach the seaward boundary. Trapping of river-borne sediment within the tidal reaches can confound the interpretation of sediment loads and annual sediment budgets, as there is not necessarily complete export during a given wet season (Ralston and Geyer 2017). Interannual variability in river flow and sediment load can also, in some cases, cause net landward transport over the year due to estuarine circulation and spatial gradients in resuspension and supply (Ganju and Schoellhamer 2006). Winds are important sediment transport drivers through two primary mechanisms: wind-waves and wind-driven residual velocities. Wind-waves have two major impacts on estuarine sediment dynamics: resuspension of bed

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sediment (mobilization), and erosion of shoreline sediment (supply). Resuspension occurs through shear stress imposed by the near-bed wave orbital velocity. Even in shallow estuaries with limited fetch, small (~ 0.2 m), short-period waves can induce sufficient shear stress to mobilize fine sediment (Ganju et al. 2020). Once mobilized, sediment is available for advection via currents to elsewhere in the estuary. The seasonal patterns of winds can lead to seasonal changes in geomorphology, in a manner similar to seasonal ocean-facing beach profiles. Intertidal flats may have seasonal profiles that oscillate from convex to concave depending on the wind-wave climate and sediment characteristics as well as biotic factors (Le Hir et al. 2000). Waves are also capable of increasing sediment supply to the estuary through erosion of the estuarine coast (Hopkinson et al. 2018). Marsh edges and bluffs can be eroded through the wave thrust (force per unit length) and its direct destabilization of the soil-root matrix. The magnitude of the wave thrust is a function of the wave energy (Leonardi et al. 2016), but the exact magnitude of sediment liberation is a complex function of soil type, vegetation presence/type, and likely many other unknown factors (Bloemendaal et al. 2021). Later in this chapter, the importance of future wave-induced edge erosion, and its role as a new sediment source to the estuary will be discussed in the context of geomorphic evolution in response to sea-level rise and wave climate. Wind-driven flow can have a significant impact on net sediment transport, especially during events with strong along-estuary winds. Estuaries with even minimal lateral gradients in bathymetry will experience a net downwind flow over shallower areas, and a compensatory return flow in the upwind direction over deeper channels (Csanady 1973). While these flows are relatively weak (~0.1 m/s) in the context of shear stress and sediment transport, they can be responsible for a significant advective flux of sediment already resuspended by wind-waves during the same wind events (Nowacki and Ganju 2018). Coastal storms impact estuaries through high water levels (via low atmospheric pressure, i.e., “reverse barometer effect”), winds, and intense precipitation and runoff. Each of these has a separate influence on sediment dynamics, and their combined effect depends on the strength of each force. For example, storms leading to intense runoff may cause massive sediment export through the estuary and to the open ocean (Wright and Nittrouer 1995; Ruhl et al. 2001), while in other systems sediment is trapped within the estuary even during the most extreme events (Ralston et al. 2013; Du et al. 2019). In some cases, the mobile pool of sediment may remain outside the estuary post-storm, and advect back into the estuary later (Downing-Kunz et al. 2021). Systems with limited freshwater input, such as lagoons and back-barrier estuaries, experience very different impacts from storms. Storm surge and waves can overtop barrier islands, leading to overwash events and sediment deposition into the estuary, as well as the creation of new inlets that provide additional conduits for water and sediment exchange between the estuary and ocean (Miselis et al. 2016).

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13.2.3 Geologic-Timescale Processes Classical conceptual models of estuarine sediment transport and geomorphology are often based on facies-models and geologic interpretation of depositional records that are derived from sediment cores or direct observations (Fletcher et al. 1990; Dalrymple et al. 1992). Models such as those presented by Dalrymple et al. (1992) place estuaries within broad geomorphic categories (e.g., deltaic vs. tide-dominated), and identify the relative strength of rivers, waves, and tidal forces. These models describe the evolution of estuaries in response to forcing and sediment supply, how the evolution is manifested by depositional facies of sediment from different sources, and the spatial distribution of sediment re-worked or transported by external forces. The interpretation is based on the influence of sediment delivery from the watershed, reworking of that sediment within the estuary, and export or import from the oceanic boundary. These models necessarily bypass the details of tidal-to-annual timescale sediment dynamics, and instead attempt to capture the overall estuary response to gradual sea-level rise and infilling. There is also an underlying implication that systems simply translate landward if sediment supply is less than the rate of sea-level rise and other external forcings are unchanged. For the purposes of this chapter, we are concerned with the variability of external forcings induced by climate change, so the shorter timescale processes become important for understanding future evolution. For a deeper understanding of these models and useful conceptual diagrams, the reader is directed to Dalrymple et al. (1992) for a general model, and Fletcher et al. (1990) for a specific model applied to a coastal plain, funnel-shaped estuary (Delaware Bay, DE). The example of Fletcher et al. (1990) is remarkable for its inclusion of the estuary turbidity maximum in both the conceptual model and the interpretation of depositional records. They describe the landward migration of the ETM depositional region as a function of sea-level rise over the last 10,000 years, which is a useful merging of a tidal-timescale process with the geologic record, and pertinent to the overall focus of this chapter. Furthermore, the case of the Delaware Bay estuary highlights the importance of source sediment. Fletcher et al. (1990) detail the importance of storms for depositing estuarine sediment on landward wetlands, while supply to the estuary (and ETM) is dependent not only on watershed supply, but edge erosion as well. In this case the shoreline retreats along existing, pre-transgression topography.

13.3 CLIMATIC DRIVERS AND RESPONSES 13.3.1 Drivers of Watershed Sediment Delivery 13.3.1.1 Sediment Yield Despite the appearance of a simple relationship between sediment load and river discharge presented above, the power-law formulation is an approximation that can vary

Climate Change and Estuaries

substantially within a single system. Warrick (2015) detailed the non-stationary aspect of sediment rating curves, indicating that the relationship between sediment discharge and flow, for a given system, may not be constant in time or even through a single event. Climate change specifically will induce changes in the rating curve by modifying the environmental conditions that control sediment yield. Warming of the atmosphere increases the carrying capacity for water vapor, and therefore the intensity of rainfall can increase in wetter regions (Trenberth 2011). Direct scaling between temperature and precipitation/flood intensity decreases from humid to arid climates, but the overall intensity is nonetheless expected to increase globally (Tabari 2020). Annual precipitation totals, and ensuing runoff to higher-order (i.e., smaller) streams, increases sediment loading to rivers that ultimately empty into estuaries (Langbein and Schumm 1958). The intensity and duration of rainfall, regardless of the mean precipitation, is ostensibly linked to sediment mobilization, but recent studies in various environments have found that these relationships may be site-specific (Wu et al. 2018; Almeida et al. 2021). Within a given system, it can be expected that increasing mean precipitation will increase sediment yield, but there can be a variable response to precipitation intensity even within a single event (Zabaleta et al. 2007). Most likely the larger influence of climate change on sediment yield will lie with the local sediment matrix conditions. Climatic elements, especially air temperature and drought frequency/intensity, have initiated land change responses including conversion of ecosystems (e.g., poleward expansion of mangroves, changes in forest types) that have implications for sediment yield and ensuing supply to estuaries. The soil matrix within a watershed naturally responds to non-anthropogenic changes in vegetation, agricultural practices, urbanization, and wildfire. Vegetation distribution over landscapes plays a strong role in sediment yield; however, landscapes with similar cover can have different yields depending on where vegetation is present, and direct relationships can be elusive (Braud et al. 2001; Rey 2003). Assuming some natural shifts in vegetation type with temperature and precipitation changes, the direct influence on sediment yield may be of secondary importance to anthropogenic changes. Direct conversion of land from natural vegetative cover to agriculture, in response to climate-induced demand for crops, will likely increase sediment yield (Shao et al. 2013), though modern soil management practices may modulate the impact. Similarly, urbanization (or de-urbanization) in response to population shifts induced by climate change alters loading to streams and rivers. Urbanization can increase sediment loads through increased direct runoff (Myronidis and Ioannou 2019) or cause large fluctuations in load due to land disturbance and post-disturbance storage in the system (Wolman and Schick 1967). The largest direct effect of climate on the sediment matrix may be through wildfires. The duration of fire seasons and fire-prone areas has increased globally in recent

Climatic Drivers of Estuarine Sediment Dynamics

decades (Jolly et al. 2015) in response to precipitation and temperature extremes. Wildfires effectively eliminate the soil-stabilizing and flow-routing effects of plants, and ultimately provide a potential increase in sediment yield to downstream estuaries (Shakesby and Doerr 2006). Areas with an increase in fire season length may therefore have increased sediment yield to the estuary, though the precise timing of precipitation events post-fire will influence the yield per unit discharge (Warrick et al. 2012). Therefore, the interaction between wildfire and precipitation in wet/ dry climates (i.e., the Mediterranean) may see the greatest changes in sediment yield in response to climate change. The creation of dams in watersheds is usually to secure water resources and/or produce hydroelectric energy. Climate change, through changing precipitation patterns and the potential for prolonged droughts, may lead to an increase in water storage needs worldwide (Ehsani et al. 2017). Simultaneously, many restoration projects target dam removal as a key aspect of restoring connectivity between ecosystems (Bednarek 2001). Dams inherently trap sediment, reducing reservoir capacity over time and regulating sediment yield to downstream estuaries (Wright and Schoellhamer 2004; Schleiss et al. 2016), though the magnitude of trapped sediment may be minimal compared to the mass transported downstream (Ralston et al. 2021). Conversely, dam removal releases reservoir-trapped sediment and the return of unregulated flow may increase downstream sediment transport as well (Randle et al. 2015). 13.3.1.2 Streamflow Timing and Magnitude Direct precipitation on the terrestrial land surface drives sediment input to higher-order streams, but once those streams merge into lower-order rivers, sediment is carried downstream suspended in the water column (suspended load) or along the bed (bedload). The total load ostensibly follows the aforementioned non-linear (i.e., power law) relationship with river flow, though the relationship can vary with time. As mentioned before, climate change (through temperature increases) directly increases water vapor capacity in the atmosphere, thereby increasing precipitation. In the tropics, precipitation extremes associated with cyclones have increased (Maxwell et al. 2021); in temperate regions, air temperature and precipitation increases may shift the spatiotemporal distribution of rain and snow, or create more “rain-on-snow” events, which increase peak runoff (Surfleet and Tullos 2013). Observations over the past decades show evidence that these changes can lead to larger, and earlier peak flows during the annual runoff season (Knowles et al. 2006). The increase in peak flow, even with unchanged annual total flow, may lead to increased sediment loading to estuaries earlier in the annual cycle through the non-linear relationship between flux and river flow. However, in heavily regulated systems such as the Columbia River estuary, reduction in peak spring flows has decreased sediment loads, and anthropogenic influence (through flow regulation) outweighs climatic signals (Naik and Jay 2011).

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A secondary effect of timing shifts is the phasing of sediment arrival with other environmental forcings, such as wind-waves and salinity distribution. For example, in the San Francisco Bay estuary, spring river flows and sediment loads are followed by a reliable wind-wave season responsible for widespread resuspension; the longitudinal salinity gradient drives landward sediment flux via estuarine circulation during the dry season (Ganju and Schoellhamer 2006). Earlier deposition of watershed sediment and an altered salinity gradient may allow for a longer season of wind-wave resuspension and ensuing landward transport. Earlier spring freshwater and sediment pulses, coupled with the timing of other resuspension mechanisms, may change the net export of sediment from the estuary (e.g. earlier delivery will result in a longer exposure of new bed sediment to wind-wave resuspension and possibly trapping efficiency). Anthropogenic influences on sediment dynamics within the estuary are discussed at length later in the chapter, but here we also note the impact of human activities on watershed sediment load. Syvitski et al. (2005) quantified the global terrestrial sediment loads to the ocean, under both pre-Anthropocene and present-day conditions. Humans are responsible for significant mobilization and trapping of sediment, through land-use change and reservoir construction, such that the overall effect is to increase sediment mobilization in the watershed, but also trap sediment within reservoirs before reaching the ocean (though there are spatial variations in rates globally). Furthermore, Syvitski and Kettner (2011) point out that the human influences on these processes likely outweigh projected climatic effects. Nonetheless, one can consider the climatic effects as a further perturbation on top of human activities and account for them in future projections of watershed sediment loads.

13.3.2 Drivers of Redistribution and Resuspension 13.3.2.1 Wind-Waves Climate-induced changes in wave intensity may arise from both air temperature changes modulating the sea breeze (caused by differential warming of land and water; Miller et al. 2003), and storm frequency and intensity changes (Knutson et al. 2010). As noted earlier, waves will have two primary impacts on sediment transport: resuspension of bed sediment and erosion of shorelines. While there may be secondary impacts including changes in wave-current interaction, the influence of waves on bedforms, and others, we will consider those outside the scope of this chapter. Bed resuspension is a function of wave-induced shear stress, caused by near-bed orbital velocity (Grant and Madsen 1979). The change in wave height and period will have a direct influence on this near-bed stress, as will changes in water level due to SLR. Increasing wave-induced shear stress will cause more mobilization (if sediment supply is ample), thereby increasing redistribution within the estuary and possibly increasing concentration gradients between the estuary and landward/seaward end members. With

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increasing wave energy, intertidal flat shapes will tend towards concavity which will encourage landward penetration of waves, though sediment source and tidal hydrodynamics will also play a role (Le Hir et al. 2000). As noted later in this section, ambient sediment concentrations in the water column can mitigate such evolution, provided that deposition can compensate for increased sea-level rise. Nonetheless, in a sediment-poor, SLR-impacted system, greater wave energy will impact the shoreline. Therefore, in both dynamic and geologic conceptual models, the more relevant long-term impact of waves may be on edge erosion (i.e., shoreline change). In conceptual models of estuarine evolution, the landward movement of the coast in response to sea-level changes is not strictly due to changes in water level, but also the erosive impact at the land-water interface (and overwash in the case of back-barrier systems). Recent work has shown that the erosion rate within a system is a linear function of wave power (the product of wave energy and group velocity; Leonardi et al. 2016). Wave power may likely increase in most systems as fetch distances and depth increase with SLR, and larger waves will penetrate into the estuary and towards the shoreline. For muddy and sandy coasts, larger waves will mobilize more sediment in the nearshore “swash” zone, and sediment-poor systems will experience net erosion as sediment is advected to lower energy components of the system (deeper channels and basins). For vegetated shorelines with more vertical edges (e.g. bluffs or marsh scarps), the wave thrust is maximized when the water level is level with the edge. Given that marsh plants colonize above mean sealevel (Redfield 1972), sea-level rise will tend to move the average water level farther up the face, thereby increasing wave thrust and edge erosion. This increase in erosion will then liberate more sediment from the shoreline, resulting in deposition either in the estuary, along wetland plains, or further offshore (Hopkinson et al. 2018). The landward movement of the shoreline and increase in estuary area will then also increase the tidal prism, leading to an adjustment of inlet and channel cross-sections to compensate (Fitzgerald et al. 2018). Additional feedbacks include less sediment trapping as vegetated fringes of the estuary are lost (Donatelli et al. 2020), though landward migration of marshes may partially compensate for this loss (Schieder et al. 2018). The interplay between sea-level rise and wave-induced erosion is an important yet often elusive concept in estuarine geomorphic evolution. In an estuarine-marsh system with no sea-level rise, waves will nonetheless attack the shoreline and mobilize sediment either through edge erosion or intertidal flat resuspension. In most cases that sediment will advect to a lower energy component of the system: either the marsh plain (during high tide inundation) or deeper portions of the estuary with lower shear stress (channels or basins). The edge will retreat, and the flat will erode, unless there is an external source of sediment to resupply both components, or a mechanism that returns those mobilized sediments to the original areas (e.g., a large storm event).

Climate Change and Estuaries

In a sediment-poor system, this process alone will cause geomorphic evolution; in a sediment-rich system there may be a quasi-equilibrium that is reached over some timescale (Fagherazzi et al. 2013). Now consider the same system, with sea-level rise imposed on top of these processes. Greater water depth allows for more wave penetration to the flat and marsh edge and accelerates the geomorphic evolution relative to the non-sea-level rise case with the same given external sediment supply. Models of such coupled processes, covered in detail for estuarine salt marshes by Fagherazzi et al. (2012), are useful for understanding the parameter space under which estuaries and marshes evolve in tandem in response to sea-level rise and sediment supply. Note that both of the aforementioned processes can be modulated by biotic processes, including bed biostabilization by benthic algae (Wood and Widdows 2003) and seagrasses (Donatelli et al. 2018a), and shoreline stabilization by salt marsh and mangrove species (Leonardi et al. 2018). Shifts in coverage of these ecosystem features, as well as efforts to re-establish their coverage through direct intervention (e.g., “living shorelines,” Bilkovic et al. 2016) may have an equivalent impact on sediment transport as will changes in physical forcing. Orbital velocities and bed shear stresses can be reduced by over 60% in the presence of seagrass, leading to a similar reduction in suspended-sediment concentration (Hansen and Reidenbach 2013), while the critical shear stress threshold can increase by an order of magnitude in the presence of benthic algae and their associated extracellular polymeric substances (de Brouwer et al. 2005). The influence of vegetation and the associated root-sediment matrix on edge erosion and shoreline change is not yet well-understood (Brooks et al. 2021), but intuitively plants and their root structures tend to encourage stability of the shoreline in comparison to unvegetated sandy or muddy coasts. 13.3.2.2 Hydrodynamics and the Estuarine Turbidity Maximum As noted, estuarine turbidity maxima (ETMs) are ubiquitous features, characterized by locally increased SSC. Geologic time-scale models gloss over tidal-to-annual timescale effects and simply assume a net landward movement of the ETM with sea-level rise and transgression. While that may indeed be the case over centennial scales, a finer understanding of ETM movement on shorter timescales is important in the context of climate change, ecosystem response, and contaminant fate. We can evaluate the response of ETMs to climate change by addressing the general types separately. For all cases, the presence of sediment either stored within the estuarine seafloor or delivered from watershed/marine end-members is a precondition for ETM formation. In systems where mobile bed sediment is depleted and/or sediment delivery is reduced, ETMs will not exist. The reader is directed to Burchard et al. (2018) for a comprehensive schematic of ETM behavior. ETMs created at the salinity intrusion limit are controlled by the estuarine circulation, i.e., the two-layer circulation that leads to near-surface seaward flow, and near-bed

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landward flow. The salinity intrusion limit will therefore be controlled by the strength of the near-bed flow and the distance to which it can penetrate. The classical formulation of Hansen and Rattray (1966), detailed in Chapter 12, shows that the gravitational circulation increases with the longitudinal salinity gradient and depth, so sea-level rise alone (assuming sediment supply does not keep up with sea-level rise) will increase the salinity intrusion limit and move the location of the ETM landward. However, increased river flow may also decrease sediment trapping within the ETM by reducing the estuarine length (Jay et al. 2007). Under a condition of vanishing or reduced freshwater flow, this ETM would likely cease to exist. ETMs created within the tidal convergence zone occur in freshwater, where tidal currents vanish relative to the compensatory river outflow. The location and strength of these ETMs will depend on the precise tidal hydrodynamics of the system, which will be affected by the morphology and response of the tidal constituents to friction within the estuary, as well as the strength of river flow (Burchard et al. 2018). Transitions in bathymetry create ETMs through local maxima in near-bed flow and sediment convergence of flow and sediment. These ETMs will likely change as sea-level rise modifies bathymetric gradients, and new ETMs may be created where new bathymetric gradients form in response to sea-level and geomorphic evolution. Analytical models, such as that presented by Talke et al. (2009), provide a useful context for first-order estimation of ETM strength and position. For example, increasing depth (through sea-level rise) and sediment settling velocity move the ETM upstream, while increasing freshwater discharge and vertical mixing move the ETM downstream. Within a given system, this type of framework can be useful for understanding the future trajectory of the ETM. Once again however, note that anthropogenic effects can trump physical forces: Winterwerp and Wang (2013) also showed with an analytical model that the dredging and deepening of tidal channels can amplify tides and increase landward transport of suspended sediment. The deposition of that sediment smooths the bed, reduces frictional effects, and amplifies the tides further, causing a runaway condition characterized by “hyperturbidity.” Indeed this condition has been observed in several systems globally (Burchard et al. 2018). It is possible that sea-level rise and associated tidal amplification in some systems may shift a system to hyperturbidity.

control the distortion of the tide via friction, and depending on the precise geometry and present dominance of the estuary, distortion may be reinforced or weakened. Ultimately each estuary will respond differently to sea-level rise in the context of tidal dominance, and simplified models can estimate the potential change in dominance. Concurrently, expansion of the estuarine area in response to shoreline change will increase tidal prism and tidal velocities through inlets, as well as adjustment of inlet morphology and flood/ebb tidal deltas if present (Fitzgerald et al. 2018). Khojasteh et al. (2021), following van Rijn (2010), presents an illustrated diagram (see Figure 13.4) of tidal response that describes how propagation is manifested in different systems; sea-level rise can shift systems from one state to another (e.g., amplification to resonance) if the geometry is sufficiently modified. Ultimately, fully resolved two- or three-dimensional numerical models can properly estimate future tidal responses to sea-level rise under given geometry and anthropogenic influence (Holleman and Stacey 2014); however, modeling the sediment dynamic response is still an uncertain task (see Research Gaps).

13.3.2.3 Tidal Currents and Dominance As water level changes due to sea-level rise, systems may switch from flood to ebb dominant (or vice-versa) based on their underlying morphology and basin geometry. This will then lead to a change in net sediment transport direction, in the direction of tidal dominance. Friedrichs et al. (1990) identified the geometric parameters controlling tidal response to sea-level rise: the ratio of tidal amplitude to mean channel depth (a/h) and the ratio of intertidal to subtidal estuarine volume (Vs/Vc). These two parameters

Though the inner shelf and nearshore zone are outside the scope of this chapter, climate change will induce large geomorphic changes at the seaward boundary that should be accounted for. Marine sediment supply can be the dominant source to some estuaries (Dronkers 1984; Thomas et al. 2002), and coupled with tidal asymmetry and/or flooddominance, net import of sediment may offset sea-level rise for a period of time. This, of course, depends on the available sediment supply at the seaward boundary, which is a complex function of alongshore and cross-shore transport,

13.3.2.4 Wind-Driven Currents In lagoonal and funnel-shaped estuaries, winds drive a downwind residual current on shoals, and a compensatory upwind flow in channels (Csanady 1973). In weakly tidal systems with minimal river input, such as back-barrier lagoons, this signal can drive significant sediment flux. Increased wind velocity will tend to then intensify this aspect of transport and development, thereby accelerating estuarine evolution over longer timescales. For example, in a relatively unimpacted system such as the Chincoteague Bay, VA, the sediment transport in the system is largely driven by episodic wind events that move sandy, subaqueous overwash features in the downwind direction (Ganju et al. 2017), and finer channel sediment upwind (Nowacki and Ganju 2018). If storminess, specifically peak wind speed and duration, increases with climate change this would accelerate the morphological development of these systems, which are present on much of the world’s coastline. Simultaneously, larger storms would increase barrier-island breaching and overwash, thereby accelerating the transgression process in concert, assuming sufficient sediment supply in the nearshore.

13.3.3 Drivers of Marine Sediment Delivery

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anthropogenic actions in the nearshore (stabilization, beach nourishment, sand mining), and the evolution of flood and ebb tidal deltas. The conceptual model of Fitzgerald et al. (2018), as applied to barrier island systems, is a thorough synthesis of the feedback between sea-level and sediment transport over flood and ebb tidal deltas, and is recommended as a resource for connecting tidal-to-annual timescale transport with geologic-timescale evolution of the outer boundary of back-barrier estuaries. Briefly, their conceptual model suggests ongoing deterioration of barrier islands in response to sea-level rise and coastal storms, and a transfer of this sediment pool to both ebb deltas and the estuaries. Fringing wetlands within the estuary will contract in response to sea-level rise, leading to a larger tidal prism and increased flood-dominance in some cases. Ultimately, conceptual and quantitative models of the component-specific sediment budgets will be required to holistically evaluate the coupled response of the nearshore and estuarine geomorphic response. Sea-level rise and storms directly impact sediment supply and geomorphology on the seaward boundary of backbarrier estuaries, in a more obvious fashion than coastal plain or river estuaries. More frequent water level extremes with sea-level rise in addition to waves will lead to more overtopping of barrier islands and increased sediment flux from the estuarine seaward boundary. Over geologic timescales, transgression will erode estuarine sediment and deliver some portion landward and the remainder to the continental shelf as the coupled barrier-estuary system transgresses over older environments (e.g., Swift 1968). Idealized barrier-island models, such as that developed by Nienhuis and Lorenzo-Trueba (2019a), are useful for estimating sediment fluxes over long timescales across coupled systems, in response to sea-level rise. Though this model is not explicitly representing the estuarine system, it provides insights into the fate of the seaward boundary, in particular, the role of tidal inlets as opposed to overwash. Using this model, Nienhuis and Lorenzo-Trueba (2019b) found that ephemeral tidal inlets are important for driving landward sediment flux into the back-barrier system, allowing for migration of the entire system in response to sealevel rise (Fitzgerald 1988). This, of course, is dependent on a lack of human intervention; inlet stabilization as well as hardening of the coast “stovepipe” these ocean-estuary sediment pathways and may decrease overall resilience. In fact, Miselis et al. (2016) showed that developed barrier island shorelines directed storm-induced overwash sediment into deeper estuarine basins, while undeveloped back-barrier areas with marsh vegetation may have trapped sediment, thereby fortifying the barrier island.

13.4 COMPLEX FEEDBACKS 13.4.1 Anthropogenic Processes Absent from most geologic-timescale conceptual models, as well as process-based models, is consideration of the

Climate Change and Estuaries

explicit intervention by humans. Over sufficiently long timescales, it is unlikely that humans will have the final say on estuarine geomorphology across all systems. Certain systems, due to their socioeconomic role, may be controlled in the short term and therefore it is important to consider those explicit actions and feedback. 13.4.1.1 Storm Surge Barriers Storm surge barriers, intended to reduce the impact of extreme water levels within the estuary, have been constructed or considered across many inlets worldwide (Bijker 2002; Chen et al. 2020; Kirshen et al. 2020). The impact on sediment transport depends strongly on the geometry of the modified inlet, but also on the operational details of the barrier, neither of which have been studied in great detail. We can speculate on the influence of barriers under non-storm, open conditions where the inlet cross-sectional area has been reduced. A reduction in cross-sectional area, given an unchanged tidal prism landward of the inlet, will lead to higher velocities and potentially enhanced sediment transport in the inlet itself; however, the attenuation of the tide throughout the estuary can also reduce the tidal prism, reduce withinestuary velocities, and reduce morphological change (Eelkema et al. 2013). Whether an open barrier leads to increased flood or ebb tidal shoals will depend on the local source of sediment, whether it is the nearshore zone or the scoured inlet. The tidal dynamics near the inlet, specifically the distortion of the oceanic tide via friction, will control tidal dominance and net sediment transport direction. The stabilization of the inlet itself is also an important factor. Preventing inlet movement and evolution has the overall effect of blocking natural sediment transport pathways in both the alongshore and cross-shore direction, so there may be indirect effects including sediment starvation of down-drift areas both in the nearshore and in the estuary. During barrier closure, tidal and storm hydrodynamics are effectively halted. Assuming a storm with no associated increase in watershed-derived sediment via river input, the closure will result in reduced sediment input to the estuary from the marine end-member. Conversely, if the estuary has significant freshwater and sediment input from the watershed, the closure will result in enhanced sediment trapping that would not typically occur. The reintroduction of tidal hydrodynamics after re-opening will provide new energy that will likely mobilize material that may not have been retained under normal conditions. This could result in enhanced SSC post-storm. Storm surge barriers are only closed during forecasted high water events, during storms or spring tides. With sea-level rise, however, the operational threshold for closing a barrier will ostensibly be crossed more frequently in the future. Therefore, the sediment transport impacts due to closure will become more dominant, and the relative influence on marine and watershed sediment input will be system and operationally specific.

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13.4.1.2 Dredging and Sediment Disposal Dredging and “beneficial re-use” of estuarine and marine sediment for restoration is an accelerating area of practice (Ulibarri et al. 2020). While these practices are spatially limited and will not be applied comprehensively across all estuaries, the actions themselves are specific responses to external forces including climate change and must be considered. Dredging of existing navigation channels is often a response to either infilling caused by over-deepening channels beyond their natural depth (to facilitate larger vessels), sediment deposition in an area of convergence (e.g., ETM), or removal of sediment for coastal nourishment projects. Agencies that conduct dredging are often limited in their ability to dispose of such sediment; therefore, there has been recent interest in using this sediment to increase wetland elevation in areas threatened by sea-level rise (Thorne et al. 2019), place the sediment within tidal channels for redistribution elsewhere (Baptist et al. 2019), or create entirely new intertidal systems by reversing reclamation (Brunetta et al. 2019). Sea-level rise is the overarching context that governs these practices. Degradation of coastal wetlands and infilling of over-deepened estuaries is exacerbated by sea-level rise, so ostensibly both of these responses will be intensified. The economic value of navigable channels and ports clearly indicates ongoing dredging. Rates of wetland loss are increasing; therefore, the practice of beneficial reuse will also increase assuming no change in regulatory forces. The addition of sediment to degraded wetland areas for increasing elevation and/or aiding vegetative colonization is intended to prolong the lifespan of the wetland and its associated ecosystem services. In fact, the act of moving sediment from one component (over-deepened channels) to another (intertidal areas) represents a proxy for marine transgression in response to sea-level rise. This presents an opportunity to mimic natural processes (transgression) by strategically placing sediment in areas that are vulnerable to open-water conversion along the cross-shore geomorphic continuum. This would imply that placing sediment in the most seaward areas only is likely unsustainable, and optimal management would aim to restore coherently along the cross-shore gradient. 13.4.1.3 Shoreline Protection Increased construction of estuarine shoreline protection can disrupt natural sediment dynamics, including the aforementioned transgression process. Open-coast shoreline protection has long been known as an obstacle to natural sediment transport pathways (Komar 2018); however, similar techniques are being applied in estuaries with substantially less attention devoted to the side effects (Ganju 2019). The type of shoreline protection varies, from levees, solid concrete bulkheads/seawalls, boulders and rip-rap, oyster castles and reefs, and/or restored marsh and seagrass fringes (Bilkovic et al. 2016). The latter are typically denoted as “living shorelines,” with the implication that they function as nature-based features. However, any attempt to stabilize a shoreline effectively reduces sediment transport and

eliminates a potential source of sediment to other components of the system. In the case of an estuary, an intentionally protected shoreline will reduce wave energy and sediment mobilization as compared to the unprotected case. That sediment may have moved seaward in the cross-shore direction to supply the intertidal flat or channel, or in the alongshore direction to a neighboring habitat, or landward to supply a wetland fringe. The reduction of transport to the other areas may benefit the locally protected shoreline, but decreases the resilience of other components by blocking a natural sediment transport pathway. Essentially, the intentional protection of a shoreline is somewhat counter to the dredging/beneficial re-use case discussed above. While the latter appears to accelerate the effect of climate change by moving sediment from the subtidal estuarine areas to intertidal areas, the former aims to lock sediment in place and maintain shoreline position. This may lead to less resilience overall if sediment input is insufficient and vertical accretion cannot keep up with the rate of rise. The importance of a three-dimensional context for sediment transport is critical for evaluating such actions in response to climate change: sediment moves in both alongshore and cross-shore directions, and stabilizing a sediment pool in one location may prevent alongshore or cross-shore transport to another location.

13.4.2 Intertidal and Subtidal Vegetation From an ecosystem services perspective, perhaps the greatest value of estuarine systems lies in the vegetated portions. Seagrass meadows, mangroves, and salt marshes all provide habitat, carbon storage, and wave attenuation in varying amounts. Furthermore, all are essential geomorphic features that interact with physical forces and sediment transport to shape estuarine geomorphic evolution. The impact of climate change on these habitats is covered in other chapters of this book; in this chapter, we focus on how potential changes in these habitats will affect sediment dynamics and feedback between sediment supply and habitat evolution. Seagrasses generally colonize subtidal areas where water column and sediment substrate conditions are suitable for growth. Both the water column and substrate conditions are directly related to sediment transport: suspended sediment in the water column influences light availability and bed resuspension directly modulates sediment composition within seagrass meadows. Climate change, through temperature and sea-level changes, will influence the future coverage of seagrass meadows (Scalpone et al. 2020). In the case of seagrass areal loss, we can expect less wave attenuation, higher bed shear stresses, and more sediment resuspension over newly bare areas (Hansen and Reidenbach 2013). This increases light attenuation and overall will continue in a negative feedback loop that prevents the reestablishment of seagrass colonies. However, van Katwijk et al. (2010) demonstrated that seagrasses can subtly alter the sediment-light environment by changes in stem density

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that feedback with turbulence in resuspension (“muddification” and “sandification”), suggesting that these complex feedbacks need to be considered in the context of climate change as well. Nonetheless, the potential reduction of seagrass coverage due to “coastal squeeze” (see below) will reduce wave attenuation, increase wave energy reaching the shoreline, accelerate edge erosion, and increase sediment supply to the estuary (Donatelli et al. 2018a). Salt marshes colonize the intertidal areas of the estuary, with different species adjusting to different elevation and salinity regimes. Their potential evolution in response to climate change is covered in detail elsewhere in this volume, but the direct interaction with sediment is discussed briefly here. Salt marshes require some input of external, inorganic sediment to both accrete vertically on the marsh plain and to allow for expansion into unvegetated intertidal areas. The expansion of marshes represents a sediment sink, because vegetated plains trap sediment, and the plant roots stabilize the substrate. Conversely, marsh loss represents a sediment source to the estuary (and potentially other vegetated areas). Marsh loss is often regarded as a vertical process (i.e., submergence), but it is more appropriately both a vertical and lateral process (Ganju et al. 2020). There are instances of direct vertical-only loss, such as episodic overwash completely smothering and permanently burying a salt marsh, but the chronic loss is three-dimensional. Edge erosion, pond expansion, and permanent submergence are relatively slower processes that liberate sediment through plant dieback and subsequent erosion of a weakened sediment substrate. An additional feedback of marsh loss is the alteration of tidal hydrodynamics. Marsh loss changes the estuarine tidal prism by increasing subtidal surface area, which can alter M2–M4 patterns and tidal dominance, as well as modulate tidal-prism/inlet-area evolution (Donatelli et al. 2018b; Fitzgerald et al. 2018). Mangroves similarly colonize intertidal estuarine and deltaic flats, but can also establish on sediment-poor, carbonate platforms. Chapter 20 of this volume connects many aspects of climate change with sediment supply to mangroves; Woodroffe et al. (2016) also provide a thorough review of mangrove interactions with sediment supply and sea-level rise and discuss a continuum of sediment supply conditions under which mangroves persist. In sediment-rich deltaic regions, mangroves colonize intertidal flats and follow the topographic changes induced by sediment dynamics. Conversely, in sediment-poor, carbonate regions mangroves have historically maintained elevation through peat accumulation. With climatic change, changes in sediment supply to sediment-rich regions will influence potential habitat distributions, while sea-level rise will challenge the autochthonous accumulation at sediment-poor sites. Mangroves that exist in sediment-dependent intertidal areas will influence sediment dynamics by stabilizing the substrate: shifts in mangrove distribution in response to temperature and/or salinity may therefore provide a sediment-stabilizing effect in newly colonized areas, and a destabilizing effect in areas of mangrove loss.

Climate Change and Estuaries

13.4.3 Coastal Squeeze When we consider the evolution of vegetated habitats in concert with anthropogenic responses to climate change, there is an eventual impasse between the forces of nature and the human desire to maintain the landscape in a fixed position. Seagrasses, as described above, will have a tendency to contract from deeper, light-limited areas, and expand into shallower areas with more light availability. In the context of sea-level rise and insufficient sediment supply, this will ostensibly lead to the migration of seagrass colonies landward to the shallower reaches of the estuary, and the loss of seagrass at the deeper seaward ends of the estuary. In a natural, unmodified system, this would mirror the classical conceptual model of transgression in response to sea-level rise. However, in a modern urbanized estuary, the hardened shoreline presents a barrier to landward migration of seagrasses, leading to “coastal squeeze” of both vegetated subtidal habitats and unvegetated intertidal flats (Pontee 2013). Similarly, on the vegetated intertidal plain (e.g., salt marsh and mangroves), a natural shoreline will move landward under the combined action of edge erosion, intertidal flat loss, and sea-level rise. Sea-level rise and storms will then modulate the inundation frequency, salinity exposure, and habitat suitability of the upland, leading to marsh migration upslope (Kirwan et al. 2016). At this end of the system, urban and agricultural landscapes will prevent the colonization of marsh plants (though salinization of groundwater may render agriculture moot). Salt marsh habitats will then be “squeezed” between the estuarine shoreline and the hardened upland. There is evidence that marsh loss at the estuary edge can be offset by gains made through upland conversion (Schieder et al. 2018). However, a comprehensive assessment of salt marshes globally will require more detailed observational data and modeling over decadal timescales (Molino et al. 2021).

13.5 RESEARCH GAPS Throughout this volume, there are numerous examples of how climate change has already modulated estuarine characteristics and functions. Some ecological communities, for example, respond rapidly to subtle temperature shifts, and decades of on-the-ground field studies have led to observable trends and direct responses to climate. Unfortunately, sediment transport is an inherently chaotic process that varies over all spatiotemporal scales: from flocculation at the tidal timescale, to episodic evolution during storms, to decadalscale anthropogenic forces, to gradual infilling in response to sea-level rise. Documenting these responses has been hampered by observational challenges and model limitations.

13.5.1 Observational Methods Estuarine sediment dynamics begin on the turbulent timescale, and cascade onwards to the tidal timescale, decadal timescale, and beyond. Therefore, an understanding of the

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shortest timescale is conceptually necessary to address decadal-scale evolution in response to broad climate shifts. Our ability to measure turbulent timescale sediment dynamics at a point has evolved in recent decades, and we now have a better understanding of both water column (flocculation) and seabed (resuspension) dynamics in response to external forcing. Of course, it is impossible to capture these processes in a spatially complete context over an entire estuary. Therefore, these measurements are most useful for informing numerical models that can represent larger spatial scales. Perhaps the most useful observational data for sediment dynamics, in the context of climate change, is bathymetric change. Variations in bathymetry, if measured “perfectly” over a given timescale, can reveal net transport directions and mechanisms, sediment budgets, and the geomorphic trajectory. The issue, in this case, is the signal-to-noise ratio: tidal timescale bathymetric changes are on the order of millimeters and undetectable by any method over the estuary spatial scale. Even bathymetric changes from the largest events (e.g., Hurricane Sandy in the northwest Atlantic) are obfuscated by instrument error and environmental conditions (Ganju et al. 2017). Currently, bathymetric change observations have only proven robust over decadal timescales (Jaffe et al. 2007). Future efforts using more robust vessel-based sensor technologies may reduce error, while satellite-based methods may provide greater temporal resolution. Regardless of improvements in sensor technology, numerical models (properly informed by observations) are ultimately necessary to evaluate combined responses to climate change.

13.5.2 Modeling Methods Much of our current understanding of climate change impacts is derived from numerical models. They provide a controlled environment to explore how physical processes interact and how natural systems respond to external forcings. The uncertainty in models, specifically those concerned with hydrodynamics and geomorphology, is covered in detail by Roache (1997), and Haff (1996). Additionally, Fringer et al. (2019) cover the state of estuarine modeling in detail. Their review covers all aspects of estuarine modeling, and research gaps in processes such as turbulence and waves will cascade to sediment transport as well. In terms of sediment transport alone, they identify several major gaps in our current understanding, which we briefly re-iterate here. First, the inherent complexity in representing a heterogenous sediment distribution both in the bed and water column is difficult, and idealizations for the true grain-size distribution and resulting settling velocity are always necessary. The precise behavior of the sediment bed in response to shear stress is idealized with formulations that are relatively primitive, and cannot account for complex interactions (mixed non-cohesive/cohesive; bioturbation, biotic stabilization) without ad-hoc empirical formulations. Lastly, our ability to model geomorphic change in response to external forcing is confounded by the preceding issues, which compound with time, and in

some cases, computational expense (though improvements in high-performance computing technology may alleviate this). In addition, of course, the paucity of observational data to calibrate models of these processes, as mentioned above, still limits the validity of the models.

13.6 CONCLUSIONS Sediment transport is a notoriously complex field of research, given the need to translate the movement of individual particles over multiple timescales into net fluxes, and ultimately geomorphic change. Nonetheless, the simpler concepts involving sediment sources, mobilization and advection mechanisms, sediment budgets, and their responses to external forcing are valuable for assessing future trajectories under projected climate change. Every estuary will respond differently to changes in precipitation, sediment yield, winds, sea-level rise, and storm intensity/duration. Predicting these changes requires a baseline understanding of present-day sources and transport mechanisms, and a conceptual understanding of how they might change in that specific system with climate change. That conceptual understanding should lead to the development of appropriate analytical and numerical models that can evaluate multiple future scenarios of change, for both a furthering of basic research and application to critical resource and management questions. As humans increasingly intervene in the coastal zone to protect communities and ecosystems from storms, sea-level rise, and coastal change, it will also be imperative to build socioeconomic drivers and associated actions into our mechanistic models.

ACKNOWLEDGMENTS I thank the editors for inviting me to contribute to this book, the U.S. Geological Survey Coastal and Marine Hazards/Resources Program for supporting my participation, and Ben Gutierrez, Peter Swarzenski, David Ralston, and David Schoellhamer for input on drafts of this chapter. Any use of trade, firm, or product names is for descriptive purposes only and does not imply endorsement by the U.S. Government. Symbols used in the schematic were obtained from the Integration and Application Network (ian​.umces​ .edu​/media​-li​brary) and vecteezy​.co​m.

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14

Climate Change Effects on Intertidal and Subtidal Environments Impacts, Projections, and Management Kerrylee Rogers, Janine Adams, Nicole Cormier, Jeffrey Kelleway, and Neil Saintilan

CONTENTS Abstract............................................................................................................................................................................... 249 14.1 Introduction............................................................................................................................................................... 249 14.2 Estuaries and the Distribution of Intertidal and Subtidal Ecosystems...................................................................... 250 14.3 Conceptualizing the Impact of Sea-Level Rise on Intertidal and Subtidal Environments....................................... 250 14.4 Influence of Other Human-Induced Climate Change Drivers................................................................................... 255 14.5 Modulating Effect of Humans................................................................................................................................... 258 14.6 Evidence of Human-Induced Climate Change Impacts............................................................................................ 259 14.7 Projected Impacts of Sea-Level Rise (Models from Paleo and Contemporary Records).........................................260 14.8 Management Actions and Policy Decisions to Improve Adaption............................................................................ 262 14.9 Conclusions................................................................................................................................................................266 Acknowledgments...............................................................................................................................................................266 References...........................................................................................................................................................................266

ABSTRACT Intertidal and subtidal environments occupy the land and sea interface and are highly susceptible to the influence of human-induced climate change and other drivers of change on both marine and terrestrial processes. As sea-level has an overwhelming influence on estuary evolution and geomorphology, sea-level rise is likely to have a profound impact on the structure and function of intertidal environments and, together with other climate change drivers, will also influence temperature and water quality. Changes in temperature and water quality will be greater in subtidal environments, although they are also modified by sea-level rise. The influence of sea-level rise on intertidal and subtidal environments can be conceptualized by considering the negative feedback between sedimentation, organic matter addition, autocompaction, subsidence, sea-level rise, and accommodation space. Other climate change drivers, such as increasing air and sea surface temperatures, extreme events, and elevated atmospheric carbon dioxide concentrations, will interact with and amplify the effects of sea-level rise. Similarly, direct human impacts on estuaries, including engineering projects and nutrient enrichment, can modulate the effects of climate change. Reducing the impacts of humans and restoring intertidal and subtidal structure and function will provide opportunities to enhance coastal adaptation, ecosystem resilience, and ecosystem services. DOI: 10.1201/9781003126096-15

Key Words:  climate change, sea-level rise, accommodation space, sedimentation, organic matter, mangrove, salt marsh, seagrass, macroalgae

14.1 INTRODUCTION Intertidal and subtidal environments in estuaries support halophytic vegetation, including mangroves, salt marshes (or tidal marshes), and seagrasses; macroalgae may also grow where conditions are favorable. Intertidal mudflats are devoid of vascular plants, but they can support benthic microalgae. Together, these ecosystems provide a range of ecosystem services that benefit people and deliver intrinsic values, including coastal protection, fisheries sustainability, carbon sequestration, nutrient cycling, raw materials, tourism, recreation, cultural heritage, education, and research services (Barbier et al. 2011). These benefits and values are increasingly recognized and underpin coastal blue economies (Spalding 2016; Mulazzani and Malorgio 2017). The risk that human-induced climate change places on intertidal and subtidal environments is concerning, and it warrants ongoing research to ensure that planning and management improve their resilience. The objective of this chapter is to consider the likely effects of climate change on intertidal and subtidal estuarine environments and to discuss the challenges of planning for and managing the uncertain future of these environments. 249

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Being situated at the land and sea interface, estuaries and their intertidal and subtidal environments are particularly sensitive to processes that influence both marine and terrestrial processes and conditions, including changes in precipitation, storm frequency and intensity, and elevated air temperatures and atmospheric carbon dioxide concentrations. However, as the formation and morphology of estuaries are associated with changes to sea-level (Roy 1984; Boyd et al. 1992; Dalrymple et al. 1992), sea-level rise is likely to be amongst the most profound climate change drivers impacting estuarine intertidal and subtidal environments, particularly when coupled with extreme weather events (storms, flood, droughts) that compound impacts. This chapter initially focuses on the effects of sea-level rise on estuarine intertidal and subtidal environments and the primary ecosystems that occupy these environments. Consideration is given to the geomorphological, hydrological, and ecological processes likely to change with sea-level rise, the influence of other climate drivers, including increases in atmospheric and dissolved CO2 concentrations, altered storm frequency and intensity, and changes in precipitation and temperature, and the compounding effect of multiple climate drivers. Evidence of the impacts of human-induced climate change on intertidal and subtidal environments is provided. Finally, management actions and policy decisions that have been implemented to ameliorate the impacts of climate change are noted.

14.2 ESTUARIES AND THE DISTRIBUTION OF INTERTIDAL AND SUBTIDAL ECOSYSTEMS In this chapter, we take a geomorphological and sedimentological perspective when examining intertidal and subtidal environments in estuaries. As such, an estuary is considered to extend from the landward limit of tidal facies to the seaward limit of coastal facies (Dalrymple et al. 1992). Based on sedimentological characteristics, the history of sediment dispersal is preserved within facies. The extent of an estuary can also be defined on the basis of salinity (Pritchard 1967), which has a fundamental control on the distribution of ecosystems that occupy intertidal and subtidal environments, as do tides, which influence inundation depths, duration, and frequency. Within an estuary, riverine processes converge with coastal and marine processes arising from waves and tides, and these aspects have been conceptualized at global (Boyd et al. 1992; Dalrymple et al. 1992) and regional scales (Roy 1984; Harris et al. 2002). The interaction of these processes occurs along a continuum. At the extremes of this continuum are: 1) river systems where river flow exceeds tidal and wave processes; 2) estuaries arising along wave-dominated coastlines where the delivery of coastal and marine sediments to estuary entrances promotes coastal barrier development; and 3) estuaries along coastlines where tidal energy is high and the estuary entrance form is more funnel-like. The relative magnitude of river, wave, and tidal energy is controlled by the interaction between tidal forcing

Climate Change and Estuaries

and climatic factors that influence rainfall and runoff, wind climates, and wave regimes. The coastal geological structure has fundamental control over the shape of bedrock valleys and influences the egress of marine processes into estuaries (Roy et al. 1980; Roy 1984). Moreover, bedrock valleys with deep or unconstricted entrances facilitate less impeded propagation of wave and tidal energy through estuary entrances. Geological characteristics also influence the size and shape of drainage basins, and modulate freshwater discharge to an estuary and coastal waters. As climate change alters river, wave, and tidal processes, it is crucial to understand the influence of these processes on intertidal and subtidal environments before the modulating effect of human-induced climate change can be considered. River, wave, and tidal processes interact across three primary zones: (1) the outer marine-dominated zone; (2) the central zone where energy is dissipated; and (3) the inner river-dominated zone. Intertidal environments occur across each of these zones, bounded by the lowest astronomical tide (LAT) at the lower limit. At the upper limit of normal tidal inundation, the highest astronomical tide (HAT), the intertidal zone transitions to the supratidal zone. Variation arises in the ecological character of these zones as individual species exhibit varying tolerance to their physicochemical properties (Figure 14.1). The combination of astronomical tides with other hydrological drivers (e.g., rainfall, runoff, groundwater, storm surge, sea-level rise) makes these boundaries diffuse; and ecotones may develop as a result.

14.3 CONCEPTUALIZING THE IMPACT OF SEA-LEVEL RISE ON INTERTIDAL AND SUBTIDAL ENVIRONMENTS As global mean sea-level (GMSL) is projected by the Intergovernmental Panel on Climate Change (IPCC) to rise between a median of 0.38 m (0.28–0.55 m likely range) under a lower sea-level rise scenario (SSP1-1.9) and 0.77 m (0.63–1.02 m likely range) under a high sea-level rise scenario (SSP5-8.5) by 2100 (Arias et al. 2021), it is anticipated that sea-level rise will significantly affect sediment supply, sedimentation, inundation and salinity regimes, and the geomorphology of intertidal and subtidal environments and their associated ecosystems and ecology. Critically, the input of organic material, largely from in situ vegetation, but also transported from catchments and adjacent marine environments, also contributes to substrate volumes (Allen 2000; Woodroffe et al. 2016). In coastal wetlands, peat formation has been found to occur synchronously with sea-level rise (Redfield 1972; McKee et al. 2007). The addition of both mineral and organic material to substrates within upper intertidal environments occupied by mangroves and salt marshes is connected by a negative feedback with inundation depth. This negative feedback serves to stabilize the geomorphological evolution of intertidal environments

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FIGURE 14.1  Energy distribution and operation of marine and riverine processes along an estuary, and the influence this has on subtidal, intertidal, and supratidal habitat distribution.

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(Pethick 1981; Fagherazzi et al. 2012). Moreover, an increase in inundation frequency and depth, often termed hydroperiod (French 2019; Kumbier et al. 2021), will incrementally follow an increase in sea-level. This increases the time available for mineral and organic sediment addition from tides, and in situ vegetation may respond by increasing productivity and adding organic material to substrates, with the outcome being an increase in the substrate volume (Morris et al. 2002; Kirwan and Murray 2007). The aforementioned negative feedback between the addition of mineral and organic material to substrates and tidal inundation leads to the self-organization of substrate elevations. The result of this self-organization is the development of estuarine floodplains that asymptote towards a depth below maximum tide level where surface elevation gain matches relative sea-level rise (Fagherazzi et al. 2012). Alternatively, these feedback responses may be disrupted by increasing hydrodynamic energy at shorelines, leading to shoreline erosion. Depth-constrained species of seagrass and macroalgae can increase substrate elevations as sea-levels rise. Seagrasses that form meadows within the intertidal zone (especially Zostera spp., Halophila spp., Phyllospadix spp.) increase elevation through accretion of mineral and organic material (Koch 2001; Madsen et al. 2001). Aboveground biomass also baffles flow velocities, dampening sediment resuspension; in addition, belowground root and rhizome structures bind sediments. The capacity to do this is dependent on flow velocities that are sufficient for the settling of particles from the water column and the ongoing survival of seagrass meadows (De Boer 2007). As intertidal seagrass meadows increase elevation over time, they may become increasingly exposed during low tides (Madsen et al. 2001), and substrate conditions may transition towards elevations suitable for other macrophytes (e.g. mangroves and salt marshes). Hence, an increment of sea-level rise will reduce risks of desiccation associated with overexposure, providing sedimentation does not outpace sea-level rise. Seagrass, macroalgae, and bryozoan meadows occupying deeper subtidal waters also baffle flow velocities and trap sediments (De Boer 2007), although a decrease in water depth arising from vertical accretion of mineral and organic material increases sheer stress, creating a negative feedback between sedimentation and erosion that generates a self-regulating balance (Koch 1999). Accordingly, they are less likely to be affected by a vertical increase in sea-level, and the influence of climate change drivers on water temperature and water quality will be more pertinent. However, sea-level rise will influence tidal dynamics and salinity regimes, providing conditions favorable for the expansion of seagrass to the middle and upper reaches of estuaries. This expansion may be periodically set-back by high rainfall events that remove seagrass and submerged aquatic vegetation (Adams 2016; Adams et al. 2016). Accommodation space is a useful concept for considering the response of intertidal environments to sea-level rise as a range of processes influencing the relationship between substrate elevations and sea-level can be incorporated.

Climate Change and Estuaries

Additionally, the concept also applies to subtidal environments, as the seaward limit of accommodation space is delimited by sediment deposition. Accommodation space, initially described by Jervey (1988), was conceived as a geological concept that incorporated sedimentation, subsidence, and eustatic sea-level change. Refinement of the concept has followed in response to increasing knowledge of processes influencing substrate volumes, such as autocompaction (i.e., consolidation of mineral and organic sediments, decomposition of organic material), and the character and source of material filling accommodation space (i.e., sediment composition) (Muto and Steel 2000; Rogers 2021). Accommodation space is bounded by: (1) the landward limit of tidal inundation at contemporary timescales or sea-level over geological timescales; (2) the basement or bedrock geology; and (3) a seaward zone where hydrodynamic energy restricts the deposition of sediments, vegetation growth is restricted, and accommodation space is “unavailable” (Figure 14.2). Mineral and organic material that has accumulated above the basement occupy “realized accommodation space”, and “available accommodation space” demarcates where accumulation of material could occur. While changing sea-levels are of primary concern when considering the effect of climate change on intertidal environments, it is important to consider other processes influencing accommodation space (Rogers 2021). In particular, subsidence of the basement, autocompaction of sediments, and decomposition of organic material within realized accommodation space can increase available accommodation space, resulting in increased inundation depth and frequency. When this occurs, the effects of sealevel rise are amplified. Conversely, the addition of mineral and organic material via a negative feedback with inundation increases substrate volumes, offsets the effects of sealevel rise, and serves to stabilize substrates. Large changes in rainfall and run-off may amplify sediment supply or increase erosion of deposited sediments, further modifying the volume of realized accommodation space. Non-depositional shorelines, such as rocky intertidal zones, have limited capacity to adapt to sea-level rise and will be submerged. Depositional intertidal and subtidal shoreline adjustment is dependent upon the relationship between the following elements: (1) the rate of increase in accommodation space (∆A), resulting from sea-level rise, autocompaction and subsidence; and (2) the rate of vertical increase in substrate volume resulting from the addition of mineral and organic material (∆S), a concept known by sedimentary geologists as the A/S ratio (Schlager 1993; Muto and Steel 1997). The outcomes of sea-level rise for the intertidal zone are typically considered in the context of sediment supply due to the overwhelming influence of this process on the evolution of intertidal substrates (Allen 2000; Woodroffe et al. 2016); however, organic matter inputs significantly contribute to elevation adjustment where sediment supply is limited and in situ vegetation productivity is high (Redfield 1972; McKee et al. 2007).

Climate Change Effects on Intertidal and Subtidal Environments

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FIGURE 14.2  (a) Intertidal and subtidal accommodation space, delimited by highest astronomical tide, basement or bedrock geology, and hydrodynamic conditions favorable for mineral and organic matter accumulation; and (b) the influence of sea-level rise, other climate change drivers and human modifications on processes operating within the negative feedback loop that underpins the geomorphological stability of intertidal and subtidal environments.

At the seaward limit of the intertidal zone, it is anticipated that where sediment supply is low and the addition of organic material inputs are low due to reduced vegetation productivity, shoreline retreat will occur (A/S>1) (Figure 14.3d). However, when the rates of sediment supply and organic matter addition exceed the rate of sea-level rise, subsidence, and autocompaction (i.e. A/S95% of standardized temperature anomalies) per decade (Lima and Wethey 2012), and the frequency, severity, and impacts of marine heatwaves in coastal zones have increased by more than 50% during the past century (Oliver et al. 2018; Smith et al. 2021). As with many of the other stressors to be discussed, the large temporal variability of the estuarine environment makes heat stress difficult to predict, as broad latitudinal gradients do not always reflect differences in thermal conditions of specific habitats, which might be more dependent upon the timing of tidal cycles or cloudiness, for instance (Helmuth 2009; Helmuth et al. 2010). A study by Helmuth et al. (2006) details this phenomenon using biomimetic sensors to document a

mosaic of habitat-specific thermal conditions for the intertidal Mytilus californianus (coast mussel) along a 14° latitudinal stretch of the U.S. west coast, from southern California to northern Washington. Affixing thermistors inside empty M. californianus shells, these researchers quantified the body temperatures of mussels under both submerged and aerial exposures to define the range of thermal conditions experienced across different habitats. Interestingly, their findings indicate that maximum body temperatures did not follow a smooth latitudinal gradient, but rather showed a patchwork of hot and cold localities that depended on factors such as wave exposure and tidal cycling (Helmuth et al. 2006). Indeed, in 2002, the maximum mussel temperature recorded in Boiler Bay, Oregon – the second-most poleward site – was ~0.5°C higher than any temperature recorded in the most equatorward site, Coal Oil Point, California, despite a difference of >1200 km (10° latitude) between the two locations (Helmuth et al. 2006). Thermal “hotspots” such as Boiler Bay could be sites of local extinction and/or barriers to poleward expansion under future climate change (Helmuth et al. 2006), or alternatively, more frequent exposures to brief heat-stress

Estuarine Shellfish and Climate Change

could confer a greater level of thermotolerance than in other less variable locations, as has been shown for Nucella canaliculata (channeled dog whelk) populations across the same region (Kuo and Sandford 2009; Somero 2010). Of course, subtidal shellfish will be less prone to localized thermal extremes. Experiments simulating future warming by elevating temperatures beyond the average conditions observed in local habitats by offsets of up to 6°C have revealed a spectrum of direct impacts on individuals that can vary across taxa and life stage, altering basal metabolic functioning and impacting growth, reproduction, and survival (O’Conner et al. 2007; Talmage and Gobler 2011; Waller et al. 2017). Elevated temperatures can negatively impact the fertilization success of some broadcast spawners, such as the purple urchin Heliocidaris erythrogramma at a 6°C offset (Byrne et al. 2009), small giant clam Tridacna maxima at a 4°C offset (Armstrong et al. 2019), and Crassostreas gigas (Pacific oyster) at 30°C, an offset of 4–12°C above natural spawning temperatures (Parker et al. 2010). Other shellfish exhibit a wide range of thermotolerance among gametes and during fertilization as evidenced by the heat shock method to induce gamete release in an aquaculture setting (Helm et al. 2004). Moreover, natural gamete release is also often tightly controlled by temperature, potentially preventing gametes from exposure to highly elevated temperatures owing to shifts in the phenology of spawning. Larvae often demonstrate increased sensitivity to elevated temperatures relative to gametic and embryonic stages. Negative effects on larval growth, development, and/or survival have been detected for some species, including delays in metamorphosis, decreased lipid synthesis, decreased size, and decreased survival in the bay scallop Argopecten irradians and hard clam Mercenaria mercenaria after 20 days of 4°C-elevated temperatures (Talmage and Gobler 2011), and decreased survival in the American lobster Homarus americanus (Waller et al. 2017) and Chilean kelp crab Taliepus dentatus (Carreja et al. 2016) under 3–4°C increases in temperature, as well as decreased egg hatch success in northern shrimp Pandalus borealis under comparable temperature increases (Arnberg et al. 2013). Elevated temperatures also resulted in increased rates of biological processes such as feeding, oxygen consumption, and larval development rates of P. borealis (Arnberg et al. 2013) and H. americanus (Waller et al. 2017), in addition to the larval development rates of more than 50 other invertebrate species (O’Connor et al. 2007). Such increases in metabolic rate could result in decreased scope for growth for shellfish (Sokolova 2013). Reductions in developmental time also suggest that planktonic dispersal distances would also be reduced, and could result in less connected, more fragmented populations (Duarte 2007). Juveniles and adults of some shellfish are also impacted by warming temperatures. Some bivalve species exhibit reductions in shell growth and/or survival (Almada-Villela et al. 1982; Talmage and Gobler 2011; Speights et al. 2017; Stevens and Gobler 2018), for instance, and elevated temperatures can also alter shell morphology (Martinez et al.

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2018) and impair immune system function (Mackenzie et al. 2014; Nardi et al. 2018). The duration of exposure to elevated temperatures can also influence biological responses. Juvenile C. virginica exposed to temperature offsets of 6°C over 30 days showed no difference in shell growth (Stevens and Gobler 2018), for example, but offsets of 4°C significantly reduced shell growth over 45 days (Talmage and Gobler 2011), and linear reductions in growth over 1–3°C-offsets were detected over a 5-month period (Speights et al. 2017). Similarly, warming has also been shown to negatively impact C. virginica survival in some cases (5 months, Speights et al. 2017), and not others (45 days, Talmage and Gobler 2011), and to have timescaledependent negative effects on weights (Talmage and Gobler 2011; Speights et al. 2017) as well. Additionally, some species and populations may benefit from future warming. Rising temperatures result in increased juvenile Callinectes sapidus (blue crab) survival and growth (Brylawski and Miller 2006; Glandon and Miller 2017) which may benefit Chesapeake Bay populations by limiting winter mortality (Glandon et al. 2019), for example. In general, any individuals existing at maximal temperatures that are below their thermal optimum should benefit from warming temperatures, provided warming does not alter their habitat or interspecific interactions in a negative way. The effects of increasing temperatures on individuals can scale up to impact populations, resulting in distributional shifts and local extinctions. For example in Long Island Sound (LIS), the once thriving >USD30,000,000 yr–1 New York/Connecticut H. americanus fishery vanished over the first two decades of the 21st century after failing to recover from a heat stress-induced mass mortality event in 1999 (Pearce and Balcom 2005). Over that time, 99% of lobster fishermen working in LIS went out of business, and the combined New York and Connecticut lobster landings fell by 98% from a maximum in 1996 of >5.5M kg to 50% A. irradians population declines if host estuaries reach 800 µatm pCO2. For another scallop species, Placopecten magellanicus (sea scallop), one of the largest fisheries in the northeast United States, an integrated assessment model combining population, biogeochemical, and socioeconomic models, predicted that projected population decline under future acidification (and warming) will translate into a USD100M loss in annual revenue in the year 2050 (Cooley et al. 2015). These results indicate the potential for disastrous outcomes for certain bivalve shellfisheries under future acidification. Most experimental research approaches to study community-wide OA impacts use either mesocosms or partially vented field enclosures. Development of free ocean CO2 enrichment (FOCE) systems, based on the terrestrial free air CO2 enrichment (FACE) systems, have emerged as one promising experimental approach to measure communitywide metrics such as net calcification or competitive and/ or symbiotic interactions, in communities associated with shellfish (Gattuso et al. 2014; Wallace 2020). FOCE systems deliver CO2 gas to experimental benthic chambers that are open to the sediment and partially vented to allow flowing ambient water to be enriched in CO2 at a designed offset, usually meant to simulate future OA scenarios (Gattuso et al. 2014). Thus far, FOCE systems deployed in seagrass beds and other coastal environments have revealed no measurable impacts on seagrass growth or physiology (Cox et al. 2016), or associated epiphytic communities (Cox et al. 2017b), but have revealed decreases in juvenile growth and survival of some seagrass-associated bivalves such as A. irradians (Wallace 2020), while not impacting others such as Lyonsia sp (Cox et al. 2017a). Experiments comparing locations with natural variation in CO2 have also provided insight into the population- or ecosystem-level impacts of ocean acidification. Natural gradients in CO2, controlled by distance from volcanic CO2 vents off the Italian coast revealed decreases in calcareous algae, sea urchin density, and gastropod density/shell strength associated with increased CO2, for example (HallSpencer et al. 2008). Experiments utilizing variations in pCO2 across habitats have also intimated some adaptation potential for certain shellfish species, as M. edulis individuals from separate populations that varied in habitat CO2 displayed differential fitness under high CO2 conditions (Thomsen et al. 2017).

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Estuarine Shellfish and Climate Change

23.2.3 Dissolved Oxygen Warming sea surface temperatures have contributed to large-scale ocean deoxygenation by reducing oxygen solubility and altering ocean stratification (Altieri and Gedan 2015; Schmidtko et al. 2017; Breitburg et al. 2018). Since 1960, global ocean DO has decreased by 2% (Schmidtko et al. 2017), with coastal deoxygenation outpacing that of the open ocean (Gilbert et al. 2010), reflecting in part the combined influence of climate change processes such as warming and regional precipitation changes, both of which enhance water column stratification, thereby reducing the diffusion of oxygen to bottom waters (Breitburg et al. 2018). Indeed, over this time, reports of coastal hypoxia have increased by over an order of magnitude (Diaz and Rosenberg 2008; Diaz et al. 2011) spurring numerous mass mortality events worldwide (Swanson and Sindermann 1979; Boesch and Rabalais 1991; Escobar et al. 2013; Ram et al. 2014). Lethal and sublethal DO thresholds for fish and other taxa of 2 or 3 mg O2 L –1 commonly mark the upper bounds of what is termed “hypoxic” water, although some species experience harmful effects at levels above these thresholds (Vaquer-Sunyer and Duarte 2008). Hypoxic episodes can also last from hours to years, with most hypoxia in estuaries occurring seasonally in the warm summer months when the stratification of the water column is greatest (Diaz and Rosenberg 2008). A recent meta-analysis of over 700 experimental comparisons revealed that hypoxic events are the only global change stressor that consistently yielded negative effects on growth, development, metabolism, reproduction, survival, and abundance in both mollusks and crustaceans (Sampaio et al. 2021). For example, low oxygen conditions can weaken the immunity of C. virginica and leave them more vulnerable to parasites (Burnett 1997; Breitburg, et al. 2015), decrease growth (Baker and Mann,1992; Keppel et al. 2016) reduce feeding (Burnett 1997), slow metabolism (Burnett 1997), decrease settlement (Baker and Mann 1992), and result in mass mortality of oyster reefs (Lenihan and Peterson 1998). Hypoxia can also influence the behavior of shellfish, resulting in changes in food preference (Neff et al. 2020) and habitat preference (Pihl et al. 1991) in mobile species such as C. sapidus, or causing infaunal species like M. mercenaria to leave their burrows and/or extend their siphons well above the sediment surface, increasing the risk of predation (Diaz and Rosenberg 1995). In some adultstage bivalves such as A. irradians, hypoxic conditions negatively impact feeding rates, contributing to the prevention of metabolic maintenance (Tomasetti et al. 2023). Hypoxia and anoxia can also decimate entire shellfish populations. Mass mortalities of the commercial surf clam, Spisula solidissima have been reported in conjunction with a sustained period of low oxygen in the New York Bight (Boesch and Rabalais 1991), and similar events involving bivalve populations have occurred in the Black Sea (Steckbauer et al. 2011) and elsewhere (Rosenberg 1985). Recurrent or persistent hypoxia can lead to communities

dominated by opportunistic species (Diaz and Rosenberg 1995), “boom-and-bust” populations (Diaz and Rosenberg 2008), and the diversion of energy from higher trophic levels to the microbial loop (Baird et al. 2004). Hence, the loss of functional traits associated with some shellfish species such as suspension-feeding and/or bioturbation can perpetuate hypoxia through positive feedback loops. In some cases, the loss of benthic fauna can create a hysteresis-like recovery response, or alternative stable states (Diaz and Rosenberg 2008; Steckbauer et al. 2011), making it even more difficult to remediate. The restoration of native species in degraded coastal systems (i.e., constructed oyster reefs or clam sanctuaries) have had mixed success but have largely failed to recover fully (Lotze et al. 2006), and often the recovery of populations to prior abundances is impeded by reduced oxygen concentrations (Baird et al. 2004; Schulte et al. 2009). Despite the universally negative impacts of hypoxia on shellfish growth, condition, and survival, a greater tolerance of hypoxia among some bivalves relative to their predators in some instances may permit population growth under hypoxia if it offers a refuge from predation (Galligan et al. 2022).

23.2.4 Salinity Climate change–induced increases in total precipitation and the intensity of precipitation (Jackson et al. 2001; Scavia et al. 2002; Konapala et al. 2020) will likely alter the temporal and spatial patterns of salinity in estuaries and will interact with other climate change stressors, for example, increasing stratification and potentially exacerbating hypoxia. Within river-dominated estuaries, intense precipitation events have the potential to move freshwater into regions that had been typically brackish zones and were hospitable to many shellfish species (Gibson et al. 2002; Najjar et al. 2010; Levinton et al. 2011). As an example, C. virginica are sessile shellfish that thrive at low salinities in estuaries (5–10 psu), in part due to lower predator and disease prevalence in such regions, and can be vulnerable to sudden freshening of estuaries due to intense rainfall (Levinton et al. 2011). In a Texas estuary, flood events resulting in sharp salinity declines caused reductions in oyster abundance, larval settlement, and filtration rates (Pollack et al. 2011). The occurrence of severe flooding associated with successive tropical storms produced severely depressed salinities in the upper reaches of Delaware Bay that caused historically high mortalities for the oyster stock (Munroe et al. 2013). Levinton et al. (2011) examined eastern oysters in the Hudson River Estuary and reported trade-offs associated with lowered salinity, as it benefited oysters by reducing disease (Perkinsus marina) and predator prevalence, but also increased vulnerability to mass mortality events associated with sudden salinity declines from excessive rainfall. The group then modeled expected future rainfall and salinity patterns and found that regions that had historically been a low salinity refuge to disease and predators will have lower survival rates under climate change (Levinton et al. 2011).

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The abundance and distribution of many estuarine species is tightly linked to salinity limits (Gunter 1961), and the ecological importance of these species (as bioturbators, suspension-feeders, habitat formers, and keystone species) suggests that shifts in salinity associated with altered precipitation, stratification, and/or sea-level rise may have disproportionate effects on ecosystem functioning. A small increase in salinity of ~3 psu in Ringkøbing Fjord, Denmark, associated with a change in sluice management, for instance, increased recruitment and abundance of Mya arenia (soft-shell clams) by >100×, shifting the benthic community from primarily deposit feeders to suspensionfeeders (Petersen et al. 2008). The study by Petersen et al. (2008) further documented significantly reduced phytoplankton and zooplankton abundances and altered abundance and distributions of submerged aquatic vegetation impacting herbivorous birds in the years following.

23.3 MULTIPLE STRESSORS IN VARIABLE ENVIRONMENTS Importantly, the climate change syndromes of warming, acidification, and deoxygenation do not occur in isolation (Gruber 2011; Boyd and Hutchins 2012). Estuaries are additionally influenced by a myriad of other anthropogenic impacts, including agricultural and waste-management practices that increase nutrient inputs to coastal waters and impact ecosystem metabolism; coastal development and hardened shorelines that reduce coastal wetlands, mangroves, salt marshes, seagrasses, and other biologically important habitats; and community food web alterations or direct depletion brought about by overfishing. Regionalscale anthropogenic impacts can result in already degraded estuarine systems where resident organisms must cope with varying durations and intensities of distinct combinations of stressors (Tyler et al. 2009; Rabalais et al. 2010; Wallace et al. 2014; Rheuban et al. 2019). The co-occurrences of some combinations of coastal/ climate change stressors are well-documented (Boyd and Hutchins 2012; Boyd et al. 2018). For example, many systems with low DO levels also have high levels of CO2 and resultant low pH levels, with both processes driven by accelerated rates of respiration (Cai et al. 2011; Wallace et al. 2014). Moreover, habitats subjected to hypoxia also often face prolonged warming (Rabalais et al. 2002; Whitney and Vlahos 2021). Other combinations of climate co-stressors, such as the co-occurrence of harmful algal blooms (HABs) with other climate stressors, are emerging as new actualities under anthropogenic climate change, for which further research is greatly needed (Griffith and Gobler 2020). Concurrent stressors can have converging physiological targets, therefore experiments assessing combinations of stressors are required to accurately document the impacts of climate change on shellfish species (Boyd and Hutchins 2012; Boyd et al. 2018). One approach involves exposing organisms to stressors both individually and in combination, to enable experimenters to quantify the difference

Climate Change and Estuaries

of combined stressors from (1) the individual effects of each stressor, and (2) the hypothetical sum of the individual stressors (Gobler and Baumann 2016). Results can be categorized as additive where the effect of the combination of stressors is equal to the effect of each individual stressor, synergistic where the effect of the combination of stressors is more severe than the sum of each individual stressor, or antagonistic where the effect of the combination of stressors is less severe than the sum of each individual stressor. Of particular concern is the potential for additive and synergistic negative effects, as these effects suggest impacts that are more severe than suspected under only one stressor. Bioenergetic frameworks that integrate the effects of combined stressors on specific species and life stages can be useful for predicting future change (Pörtner 2010; Sokolova 2013). Under these frameworks, concurrent stressors can constrict the energy available for activities beyond basal metabolism (i.e., growth, immunity, reproduction). Hence, persistent or extreme stress under multiple stressors may demand more energy than is available for an organism and lead to a negative aerobic scope and, ultimately, mortality. Sites with thermal, oxygen, and/or carbonate conditions that result in consistent energy limitation and overall reductions in fitness, growth, and/or reproductive impairment, can cause distributional limits of a given species and can be used to predict ecosystem-level change for a given species under future conditions (Sokolova 2013). Below we document some of the established and newly emerging climate change stressor pairs, reporting the effects of both chronic and fluctuating exposures on coastal shellfish species.

23.3.1 Coastal Hypoxia and Coastal Acidification In estuaries, the shallow waters, proximity to terrestrial nutrient sources, and often complex topography grant watershed and biological processes a large influence over coastal DO conditions. In many of these systems intense biological respiration consumes O2, depleting the water of DO, and generating CO2, impacting the carbonate chemistry via increasing dissolved inorganic carbon (DIC) and ultimately lowering the pH (Burnett 1997; Baumann et al. 2015; Baumann and Smith 2017). Hence, when excess carbon is metabolized in situ it can stimulate coastal acidification (Howarth et al. 2011; Wallace et al. 2014, 2021), and hypoxic waters are indeed both reduced in DO and pH. Hypoxic waters high in CO2 exhibit reduced buffering capacities relative to oxic waters, so in hypoxic systems, each additional increase in DIC will result in a successively greater reduction in pH (Cai et al. 2011; Sunda and Cai 2012). Thus, organisms exposed to hypoxic stress must also cope with acidification stress that will continue to grow more severe under future OA. Unfortunately, decades of research on hypoxic stress sparged seawater with nitrogen gas to establish low DO conditions, an approach that not only displaced O2 but also

Estuarine Shellfish and Climate Change

expelled CO2 from experimental vessels and, therefore, disregarded any additional acidification stress coastal organisms may endure under hypoxia, or at worst, inflated pH values to unrealistic, basified conditions (Gobler et al. 2014; Tomasetti and Gobler 2020; Steckbauer et al. 2020). To address this shortcoming, more recent experimental work involving coastal hypoxia often bubbles CO2 gas with N2 gas to recreate the combined hypoxic and acidified conditions commonly exhibited in estuaries. In addition, coastal shellfish, given their (1) economic and ecological importance, (2) physiological requirements for CO32–, and (3) high likelihood of exposure to hypoxic/acidified habitats, are among the earliest and most commonly tested groups to be investigated to date (Gobler and Bauman 2016; Steckbauer et al. 2020; Figure 23.4). A meta-analysis of past and recent experimental investigations testing the biological responses to hypoxia and acidification individually and in combination revealed an overwhelming predominance of additive followed by synergistic negative effects in mollusks, echinoderms, and crustaceans, signifying that the outcomes were most severe when both stressors were experienced together (Steckbauer et al. 2020). Larval M. mercenaria and A. irradians under hypoxia and acidification demonstrated additive and synergistic negative responses that included lower growth rates, stunted development, and enhanced mortality (Gobler et al. 2014). Additive effects on growth and survival were also observed in larval C. virginica (Clark and Gobler 2016), and for Mytilus coriscus (thick shell mussel) the combination of hypoxia and acidification can increase vulnerability

FIGURE 23.4  General trends in experimental findings for shellfish under decreasing oxygen (DO) and pH, individually and in combination (Gobler and Baumann 2016; Steckbauer et al. 2020). The solid arrows indicate the directions of DO and pH change in estuarine ecosystems.

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to crab predation (Sui et al. 2015) and negatively affect the immune and general defense responses (Sui et al. 2016; 2017). Combined impacts on reproduction include significant reductions in the percentage of motile sperm and fertilization success of the purple sea urchin Paracentrotus lividus (Graham et al. 2015) but remain largely unstudied across other species and taxa. In crustaceans, mixed responses have been detected at various life stages. High CO2 concentrations alleviated fatigue associated with exercise in benthic-stage C. sapidus under hypoxia (Stover et al. 2013) by increasing oxygen binding affinity (Mangum and Burnett 1986; Lehtonen and Burnett 2016), an antagonistic effect that may help individuals escape hypoxic conditions (Lehtonen and Burnett 2016). Concurrent exposure to low pH and low DO can also have significant negative effects on blue crabs, for example impairing their defense system and immune response system (Holman et al. 2004; Tanner et al. 2006). For larval stage individuals, low DO and low pH significantly decreased survival in an additive manner (Tomasetti et al. 2018). The few studies that have assessed the biological responses to hypoxia and acidification of other crustaceans beyond C. sapidus have detected a varied mixture of antagonistic and additive metabolic responses (Steckbauer et al. 2015, 2020; Fontanini et al. 2018). Short, repeated declines to hypoxic and acidified conditions can occur in association with tidal/diel cycles, particularly in shallow environments where stratification is rare, and are generally expected to become more persistent or severe under continued climate change as warming temperatures favor a disproportionate increase in the rate of community respiration relative to production (Brown et al. 2004; Harris et al. 2006). Studies with larval and juvenile mollusks have revealed negative effects associated with fluctuating hypoxia and/or acidification, including decreased growth (Breitburg et al. 2015; Clark and Gobler 2016; Gobler et al. 2017), decreased survival (Clark and Gobler 2016; Gobler et al. 2017), and increased Perkinsus infection prevalence (Breitburg et al. 2015; Keppel et al. 2015), relative to static, control conditions. In one study with larval C. sapidus, significant mortality associated with fluctuating DO/pH would have been concealed otherwise under static scenarios with equivalent DO/pH daily averages (Tomasetti et al. 2021). Pooling data from nine experiments, exposures of 96 h or less to diel cycling hypoxic conditions routinely caused more severe larval mortality than static treatments with the same average DO values (Tomasetti et al. 2021; Figure 23.5). By quantifying the timescales and magnitudes of dual hypoxia- and acidification-impairment in their local estuaries, and recreating that variability under controlled laboratory conditions, this study highlights the need to incorporate fluctuating exposures to multiple stressors into experimental designs. These studies will be of increasing importance under future climate change, as episodic extremes will likely become more frequent, persistent, and/or severe (Waldbusser and Salisbury 2014).

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FIGURE 23.5  Survival of Callinectes sapidus zoea I larvae (relative to control conditions) under static vs fluctuating dissolved oxygen (DO) conditions. Each point represents one experimental low DO treatment from five experiments conducted by Tomasetti et al. 2021 and four experiments conducted by Tomasetti et al. 2018. The shade of red indicates the duration of hypoxia (~2.0 mg L –1) over a diel cycle within the fluctuating treatment. Reprinted from Tomasetti et al. 2021.

23.3.2 Warming and Hypoxia Of the numerous emerging threats to coastal estuaries, warming ocean temperatures and low or decreasing DO concentrations have been identified as some of the most pervasive and impactful (Diaz and Rosenberg 2008; Halpern et al. 2008; Diaz et al. 2011). Warming temperatures are likely to exacerbate hypoxia in many coastal regions by decreasing oxygen saturation and increasing stratification (Figure 23.6). In a long-term study in Danish coastal waters, a modest 3.2% decrease in DO saturation consistent with the increase in temperature over that time caused hypoxic thresholds to be exceeded in the system, contributing to the largest and most severe hypoxic event in the region despite nearly two decades of reductions in nutrient loading prior to the event (Conley et al. 2007). In a similar vein, reductions in nitrogen loads to the Long Island Sound – a result of two decades of systematic upgrades to sewage treatment plants – were unable to reverse DO declines during the first decade of improvements (Wilson et al. 2008). While sediments may act as a reservoir of nutrients causing a time delay in the change of DO levels, climatic processes, namely an increase in stratification, were responsible for the persistence of hypoxia in the estuary (Wilson et al. 2008). Interestingly, an improvement in the severity and extent of hypoxic impairment was recently detected in the Long Island Sound after additional analyses of the most recent decade of data, although future warming threatens to annul the recorded improvements (Whitney and Vlahos 2021). Regional increases in precipitation can also influence the volume of freshwater (and associated nutrients) entering coastal estuaries, with higher inflow resulting in increased stratification and primary production (Justić et al. 1996; Scavia et al. 2002). For instance, the “Mississippi River flood” of 1993 increased net primary production by

Climate Change and Estuaries

an order of magnitude and nearly doubled the areal coverage of the Gulf of Mexico dead zone when compared to the associated seven-year average (Justić et al. 1996). Similarly, increased extreme rainfall and flooding events were shown to increase organic matter loading in the Neuse River Estuary, accounting for up to 80% of annual loads in certain years (Paerl et al. 2020), potentially having massive impacts on DO concentrations in downstream habitats. Warming temperatures will also result in an increased likelihood of tilting systems toward heterotrophy because the increase in the rate of heterotrophic respiration would exceed the concurrent increase in the rate of primary production (Brown et al. 2004; Harris et al. 2006). Taken with simultaneous changes in DO saturation, these processes could cause hypoxic thresholds to be exceeded in declining systems, or result in an increased duration, frequency, or intensity of hypoxia in systems already exhibiting some level of hypoxic impairment. In addition to influencing the extent and severity of hypoxia, warming temperatures also affect the vulnerability of shellfish to low oxygen concentrations. A meta-analysis assessing hypoxic thresholds in marine benthic organisms, including benthic crustacean, mollusk, and echinoderm species, revealed that as temperature increased, time until death (under hypoxic conditions) decreased, and the DO concentrations at which mortality occurred increased (Vaquer-Sunyer and Duarte 2011). Concurrent warming and hypoxia may induce synergistic negative effects on aerobic scope, simultaneously increasing metabolic oxygen demands while decreasing the availability of oxygen to benthic species (Figure 23.6). Yet, there are limited studies that assess the effects of these two stressors concurrently (see

FIGURE 23.6  General trends in dissolved oxygen (DO) under rising average coastal temperatures with implications on estuarine shellfish. The solid arrows indicate the direction of DO change with increasing temperature in estuarine ecosystems.

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Estuarine Shellfish and Climate Change

Sampaio et al. 2021). Recently, analysis of decadal-scale warming, summer estuarine heatwaves, high-frequency oxygen records, and field/laboratory experimental results on adult A. irradians revealed complex linkages between concomitant thermal- and oxygen-extremes and bay scallop mass mortality events in the Peconic Estuary (New York, United States), the southernmost edge of commercially viable northern bay scallop populations (Tomasetti et al., 2023). Laboratory experiments on coastal invertebrates have additionally indicated interactive negative effects of warming and hypoxia on developmental success (Vasquez et al. 2015) and the antioxidant response system (Khan et al. 2021). Furthermore, a recent field study surveying the abundance of larval bivalves across various tropical lagoonal sites revealed that the combination of high temperature and low DO reduced larval abundance, indicating that future warming could strengthen these trends and result in a reduction in larval settlement/recruitment in areas susceptible to hypoxia (Weinstock and Collin 2021).

23.3.3 HABs, Phytoplankton, and Other Climate Stressors Harmful algal blooms (HABs) are the overgrowth of a specific phytoplankton species that results in harm to organisms and/or ecosystems, and HABs are well known for both making shellfish toxic and thus unharvestable/inedible/unsafe, as well as causing direct harm to shellfish, including mortality (Shumway 1990). There is a scientific consensus that the number of reported HABs and the economic and societal impacts of HABs have increased in recent decades (Anderson et al. 2012; Hallengeaff et al. 2021). The IPCC Special Report on the Ocean and Cryosphere in a Changing Climate (Bindoff et al. 2019) indicated that HABs have displayed a range expansion and increased frequency in coastal areas since the 1980s partly attributed to the effects of surface warming, heatwaves, deoxygenation, eutrophication, and pollution, and that HABs have negatively impacted economies, food security, tourism, and human health. Models have predicted increases in HABs due to climate change in the near-term (Glibert et al. 2014; Moore et al. 2015; BoivinRioux et al. 2021), although such an outcome will always be ecosystem and species-specific. Regardless, given the certainty of future climate change (IPCC 2021), the net effects of climate change and HABs on shellfish in the future will be more intense. Studies assessing the combined impacts of HABs and other climate change stressors on shellfish are rare (n~4; Griffith and Gobler 2020). Interestingly, these limited studies have reported strong species-specific interactions between HABs and climate change co-stressors, yielding outcomes for bivalves and other aquatic organisms that could not have been predicted based on investigations of these factors individually. For example, the survival of A. irradians larvae was synergistically suppressed by the coexposure to brown tide (Aureococcus anophagefferens) and

acidification (Talmage and Gobler 2012). Bivalve larvae whose parents underwent gametogenesis under acidification experienced a higher rate of mortality when exposed to the dinoflagellate Cochlodinium (Margalefidinium) polykrikoides compared to individuals reared from normal parents (Griffith and Gobler 2017). Sail mussels (freshwater; Hyropsis cumingii) exposed to hypoxia and the HAB toxin microcystin, which is made by freshwater cyanobacteria but is prevalent in estuaries (Preece et al. 2017), experienced a reduced scope, diminished immune response, and increased cellular damage compared to control mussels (Hu et al. 2016; Wu et al. 2017). Like many climate change co-stressor combinations, many more studies are needed to understand how climate change processes will interact with HABs to impact shellfish. High priorities will be identifying regions that will be newly or more prone to specific HABs in order to minimize effects on shellfish and shellfish products, as warming temperatures can make ecosystems more or newly vulnerable to HABs (Glibert et al. 2014; Moore et al. 2015; Gobler et al. 2017; Boivin-Rioux et al. 2021). Assessments of even the most common of co-occurring stressors (e.g. hypoxia and HABs on estuarine bivalves; Griffith and Gobler 2020) are much needed as no such studies have been performed to date. Beyond HABs, there are likely to be changes in the total amount and types of phytoplankton available as food for bivalves in estuaries. Warming surface waters can increase stratification, decrease the availability of nutrients to surface waters, and decrease phytoplankton biomass and mean cell size (Behrenfeld et al. 2006; Boyce et al. 2010). Climate change–induced increased precipitation may also enhance stratification and reduce phytoplankton biomass (Jackson et al. 2001; Scavia et al. 2002; Konapala et al. 2020). Conversely, more precipitation may accelerate nutrient loading and enhance phytoplankton productivity and biomass and alter community composition (Paerl et al. 2018; 2020). Hence, while it seems certain that climate change processes will alter phytoplankton populations in estuaries (see Chapter 17) and thus the food supply for bivalves, the precise interaction and thus outcome is likely to be ecosystem dependent.

23.3.4 Disease and Other Climate Stressors Chronically stressed animals are more vulnerable to disease, as immune function can weaken under energetic strain (Coates and Söderhäll 2021). As discussed previously, the collapse of the H. americanus fishery in Long Island Sound was the result of temperature stress and infection with the parasitic amoeba N. pemaquidensis (Pearce and Balcolm 2005). Warming temperatures can also favor the growth of common shellfish pathogens like Haplosporidium nelsoni (causing MSX in oysters) and P. marinus (causing Dermo in oysters) and lead to range expansions, like for the parasite P. marinus across the US east coast (Ford 1996; Ford and Chintala 2006). Conversely, some parasite infections can

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become less intense under warmer temperature (e.g., QPX in M. mercenaria; Dahl et al. 2011). Hypoxia and acidification have also been shown to influence infection susceptibility and prevalence. Across 14 sites in the Chesapeake Bay system, variation in P. marinus infection prevalence was explained by the interaction of salinity and the daily DO minimum, with sites containing lower daily DO minima exhibiting higher total infection prevalence (Breitburg et al. 2015). Infection intensity was best explained by the frequency of exposure to diel cycling hypoxia and acidification (Breitburg et al. 2015). Laboratory experiments with M. mercenaria revealed an increased susceptibility to infections by Vibrio (bacteria) species in larval and juvenile clams exposed to OA, and elevated pCO2 also favored the growth of Vibrio (Schwaner et al. 2020). Interestingly, in one case, parasite infections by any one of three species of parasitic trematodes actually improved the survival of Zeacumantus subcarinatus (“southern creeper” gastropod snail) relative to noninfected individuals under acidified conditions (MacLeod and Poulin 2016). This suggests that the responses of the host-parasite relationship under OA are complex, and at times unpredictable, warranting more research attention.

23.4 CONCLUSIONS The acceleration of climate change, pervasiveness of coastal degradation, and legacy of overharvesting collectively present a significant suite of challenges to coastal shellfish. Already faced with combinations of global change stressors over varied temporal and spatial scales, communities and researchers are increasingly working in collaboration to protect shellfish resources through data-driven restorative, conservation, and policy advancements. Aquaculture has emerged as one promising approach toward shellfish restoration and conservation, particularly in estuaries where natural populations are recruitment-limited (Froehlich et al. 2017; Ridlon et al. 2021). Aquaculture techniques to restore oyster populations of C. virginica (Brumbaugh et al. 2000) and Ostrea lurida (Ridlon et al. 2021) can also improve ecosystem health via increasing filtration capacity and improving water quality. Or alternatively, in some cases, dense concentrations of calcifying organisms can act as a net source of CO2 to the environment via reductions in TA during shell production (Morris and Humphreys 2019; Yang et al. 2021). Shellfish aquaculture, however, may increase resilience of individuals and populations to future change. A study on the Sydney rock oyster, S. glomerata, indicated that aquaculture helped to ameliorate the negative impacts of OA through selective breeding (Fitzer et al. 2019). Macroalgae aquaculture production has sharply increased in recent decades (Ferdouse et al. 2018) and has been increasingly integrated into multitrophic aquaculture systems involving the co-culture of shellfish (Fernández et al. 2019; Liu et al. 2021; Figure 23.7). Many species of cultivated macroalgae readily assimilate nutrients (Ahn et al.

Climate Change and Estuaries

FIGURE 23.7  Integrated kelp aquaculture on an oyster farm in Moriches Bay, New York. Pictured left to right: Stony Brook University research scientist, Michael Doall with the owner of Great Gun Shellfish, Paul McCormick. Photo credit: Stephen Tomasetti.

1998; Kim et al. 2014, 2015; Marinho et al. 2015) and CO2 (Fernández et al. 2019; Xiao et al. 2021), particularly to the benefit of bivalves (Wahl et al. 2018; Young and Gobler 2018; Young et al. 2022). A recent study has also shown the potential for commercial-scale production of some species of macroalgae to mitigate HABs via the release of allelopathic compounds (Sylvers and Gobler 2021), which are responsible for annual surplus shellfishery losses estimated in the millions (USD) globally (Hoagland and Scatasta 2006; Diaz et al. 2019). Recent international policy to mitigate climate change (Paris Accord) and to promote the conservation and sustainable use of marine/coastal resources (United Nations’ Sustainable Development Goal 14) reflects a growing emphasis on managing the global threats to coastal/estuarine natural resources (Duarte et al. 2020). In the last few years national, subnational, and tribal leaders have established advisory and action groups consisting of researchers, state officials, fishing/aquaculture industry representatives, and other community members to share information and coordinate a response to the issue of ocean/coastal acidification (see https://www​.oaalliance​.org/). Improved protection will also be facilitated via the revision of existing policies and regulatory standards. Policies that consider the effects of co-stressors rather than single stressors, for example, can improve the protection of shellfish species by accounting for additive and synergistic negative effects (Tomasetti and Gobler 2020). At the local scale, purposeful introduction of loose and crushed bivalve shells has been shown to effectively buffer sediments and significantly increase sediment porewater calcium carbonate saturation states to the direct benefit of infaunal shellfish such as M. mercenaria (Curtin et al. 2022). As climate change accelerates, autonomous observing platforms, high-frequency sensor arrays, and long-term continuous monitoring programs and datasets will be invaluable

Estuarine Shellfish and Climate Change

to quantify estuarine biogeochemical and ecological changes. There is still much to learn about the chronic effects of climate change syndromes on coastal shellfish, particularly in variable environments under multiple co-occurring stressors; most experiments to date have collectively served to establish a foundation, considering the effects of static, isolated climate stressors over short timescales. Investigations that consider multiple generations and utilize omic approaches will reveal more about the adaptive capacity of various species and populations under rapid coastal change. To that end, impacts across larger scales of biological organization, such as population- and ecosystem-level impacts are still largely unstudied and much needed to ultimately forecast future climate change impacts on estuarine shellfish.

ACKNOWLEDGMENTS We would like to thank the editors of this volume for their helpful feedback. This chapter was supported with support from New York Sea Grant, the Chicago Community Trust, and the Laurie Landeau Foundation.

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473 J. K. Petersen, and Ø. Strand (eds.), Goods and services of marine bivalves. New York: Springer International Publishing. Wilberg, M. J., M. E. Livings, J. S. Barkman, B. T. Morris, and J. M. Robinson. 2011. Overfishing, disease, habitat loss, and potential extirpation of oysters in upper Chesapeake Bay. Mar. Ecol. Prog. Ser. 436: 131–144. Wilson, M. A., R. Costanza, R. Boumans, and S. Liu. 2005. Integrated assessment and valuation of ecosystem goods and services provided by coastal systems. Pp. 1–24. In: M. A. Wilson (ed.), The intertidal ecosystem: The value of Ireland’s shores. Dublin, Ireland: Royal Irish Academy. Wilson, R. E., R. L. Swanson. 2005. A perspective on bottom water temperature anomalies in Long Island Sound during the 1999 lobster mortality event. J. Shellfish Res. 24(3): 825–830. Wilson, R. E., R. L. Swanson, and H. A. Crowley. 2008. Perspectives on long‐term variations in hypoxic conditions in western Long Island Sound. J. Geophys. Res. C: Oceans 113(C12). https://doi​.org​/10​.1029​/2007JC004693. Winder, M., J. Carstensen, A. W. E. Galloway, H. H. Jakobsen, and J. E. Cloern. 2017. The land-sea interface: A source of high-quality phytoplankton to support secondary production. Limnol. Oceanogr. 62(S1): S258–S271. Woodside, C. 2018. Long Island Sound to lobsters: Is this farewell? Wrack Lines 18: 14–17. Wu, F., H. Kong, Y. Shang, Z. Zhou, Y. Gul, Q. Liu, and M. Hu. 2017. Histopathological alterations in triangle sail mussel (Hyriopsis cumingii) exposed to toxic cyanobacteria (Microcystis aeruginosa) under hypoxia. Aquaculture 467: 182–189. Xiao, X., S. Agustí, Y. Yu, Y. Huang, W. Chen, J. Hu, C. Li, K. Li, F. Wei, Y. Lu, C. Xu, Z. Chen, S. Liu, J. Zeng, J. Wu, and C. M. Duarte. 2021. Seaweed farms provide refugia from ocean acidification. Sci. Total Environ. 776: 145192. Yang, B., X. Gao, J. Zhao, Y. Liu, L. Xie, X. Lv, and Q. Xing. 2021. Summer deoxygenation in a bay scallop (Argopecten irradians) farming area: The decisive role of water temperature, stratification and beyond. Mar. Pollut. Bull. 173(B): 113092. Young, C. S. and C. J. Gobler. 2018. The ability of macroalgae to mitigate the negative effects of ocean acidification on four species of North Atlantic bivalve. Biogeosciences 15(20): 6167–6183. Young, C. S., L. H. Sylvers, S. J. Tomasetti, A. Lundstrom, C. Schenone, M. H. Doall, and C. J. Gobler. 2022. Kelp (Saccharina latissima) mitigates coastal ocean acidification and increases the growth of North Atlantic bivalves in lab experiments and on an oyster farm. Front. Mar. Sci. 9. https://doi​.org​/10​.3389​/fmars​.2022​.881254. Zeebe, R. E. 2012. History of seawater carbonate chemistry, atmospheric CO2, and ocean acidification. Annu. Rev. Earth Planet. Sci. 40(1): 141–165. https://doi​.org​/10​.1146​/ annurev​-earth​- 042711​-105521.

24

Climate Change Effects on Fish Populations Alan K. Whitfield, Bronwyn M. Gillanders, and Kenneth W. Able

CONTENTS Abstract............................................................................................................................................................................... 475 24.1 Introduction............................................................................................................................................................... 475 24.2 Fish Guilds in Estuaries............................................................................................................................................. 477 24.3 Climate Change Drivers for Fishes in Estuaries....................................................................................................... 478 24.3.1 River Flow..................................................................................................................................................... 478 24.3.2 Salinity Regime.............................................................................................................................................480 24.3.3 Temperature Changes.................................................................................................................................... 481 24.3.4 Sea-Level Rise............................................................................................................................................... 482 24.3.5 Estuary Connectivity..................................................................................................................................... 483 24.3.6 Declining Dissolved Oxygen, Increasing Carbon Dioxide, and Lower pH Values.......................................484 24.3.7 Spreading Diseases and Parasites..................................................................................................................484 24.4 Global Examples of Changing Estuarine Fish Populations....................................................................................... 485 24.4.1 Temperate Northern Atlantic and Northern Pacific Estuaries...................................................................... 485 24.4.2 Tropical Atlantic and Indo-Pacific Estuaries................................................................................................. 488 24.4.3 Temperate Southern Atlantic and Southern Indo-Pacific Estuaries.............................................................. 489 24.4.4 Polar Estuaries............................................................................................................................................... 493 24.5 The Way Forward...................................................................................................................................................... 494 Acknowledgments............................................................................................................................................................... 496 References........................................................................................................................................................................... 496

ABSTRACT Estuaries are a key aquatic ecosystem most likely to be influenced by climate change. These systems will be direct recipients of major changes to their physico-chemical conditions by riverine, marine, terrestrial, and atmospheric drivers associated with climate change. In addition to these changes, fishes in estuaries are increasingly subjected to habitat degradation, water pollution, and fishery exploitation, all of which are creating stressful conditions for an ichthyofauna already under significant natural environmental pressures. The major climate change drivers are alterations in river flow, salinity, temperature, dissolved oxygen, sea-level, and connectivity between habitats and ecosystems. The drivers likely to have less influence on estuaryassociated fishes this century are increases in dissolved carbon dioxide, declines in pH, and the spread of tropical diseases and parasites. Alternatively, we simply have insufficient information to understand their influence. A review of climate change impacts on estuarine fish assemblages around the world indicates that these impacts are not distributed equally, i.e., they vary among estuaries and associated habitats, processes therein, as well as across biogeographic regions. Yet, one common aspect to all regions is the lack of baseline understanding of the biology of many DOI: 10.1201/9781003126096-26

estuarine fish species, including their environmental sensitivities and tolerances. This is especially the case in polar regions where even documentation of the estuary-associated fish assemblages is largely missing, but initial information suggests these are becoming dynamic and increasingly functional. Looking to the future, research should focus on filling the above gaps, as well as how other anthropogenic impacts in estuaries will interact and compound climaterelated changes. The cumulative effects of such impacts will likely increase the vulnerability of fish species, communities, and estuarine ecosystems to climate change, but this has not been quantified. Harnessing the power of modeling approaches to anticipate the impacts of global and regional climate change on estuarine environments at ecologically relevant scales will be vital to assisting estuarine fish assemblages to weather the storm of climate change that is already upon us. Key Words:  fishes, estuaries, global, climate, tropical, temperate, polar, change

24.1 INTRODUCTION Global aquatic scientists, through their societies, associations, and organizations, issued a formal online assessment 475

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during the latter half of 2020 in which they indicated that the scientific evidence for human-caused climate change is overwhelming, and is accelerating the degradation of aquatic ecosystems and the services that they provide (American Fisheries Society et al. 2020). Since approximately 40% of the world’s population lives within 100 km of the coast, and many of these people are clustered around estuaries (e.g., New York, London, Lagos, Sydney, and Hong Kong), these aquatic systems in particular are under major pressure. The impacts on fishes in estuaries, which include habitat degradation, pollution, and exploitation, are in addition to the already demanding natural physicochemical conditions that are part of the lives of all fishes living in estuaries (Fonseca et al. 2014). The question then arises – given the combined impacts of natural variability, climate change, and other global changes, how far can fish in estuaries be pushed before the “envelope” bursts? Ultimately sea-level rise will probably have the greatest long-term impact on estuarine fishes via the potential loss, reduction, or impairment of essential habitats. If the actual loss of fish habitats due to human developments in and around estuaries is added to the above potential loss, then a concerted management effort will be required to restore essential fish habitat in estuaries and prevent the “double whammy” of climate and other global changes (Cloern et al. 2016) from causing the loss or severe decimation of fishes in estuaries. The same principal applies to increasing water pollution and reduced river flow into estuaries due to human use – these additional global changes (Cabral et al. 2019; Gudmundsson et al. 2021) can be magnified in their negative effect on fishes in estuaries if climate change impacts are not drastically reduced or reversed. The extent to which human-induced climate change is responsible for the decline in species richness and the abundance of fishes within estuaries is difficult to determine, primarily because other anthropogenic pressures are

Climate Change and Estuaries

operating in parallel with climate change. This view is eloquently expressed by Ficke et al. (2007) for freshwater systems when they stated “Currently, the magnitude of global climate change is such that most of its effects on freshwater fisheries could be easily masked by or attributed to other anthropogenic influences, such as deforestation, overexploitation and land use change.” In this review, we will focus on the impact of climate change on estuary-associated fish assemblages but will also place these changes in the context of some of the broader global change pressures that are occurring in tandem with climate change. Global climate change is expected to lead to long-term, broad-scale changes in a suite of environmental drivers that pose emerging threats to estuarine biodiversity and ecosystem functioning, as well as geomorphologic and biophysical processes within these systems. An example of how coastal temperature changes due to climate change may cause a “squeeze” on marine fish species associated with warm-temperate estuaries along the South African coast is illustrated in Figure 24.1. The increased tropical Agulhas Current water temperatures along the east coast associated with global warming, and increased upwelling on the west coast associated with changing prevailing wind patterns, will combine to squeeze warm-temperate endemic fish species into a smaller coastal biogeographic area (Whitfield et al. 2016). This “biogeographic squeeze” is more likely to occur in temperate global estuaries with an east-west coastal orientation than those situated on coasts with a north-south orientation. Some climate-induced changes are likely to be swift whilst others will be protracted but, in both cases, sound knowledge of estuarine fish communities and ecosystem functioning will be pivotal to recognizing effects and predicting future impacts in a timely manner (Gillanders et al. 2021). Only then will it be possible to safeguard the myriad of ecological functions and services that estuaries provide

FIGURE 24.1  Diagrammatic representation of the major distributional changes that are occurring in the three biogeographic zones along the southern African coastline. The MODIS satellite sea surface temperature image is a 10 day average from February 2009 (after Whitfield et al. 2016).

Climate Change Effects on Fish Populations

(Boerema and Meire 2017), in particular, their nursery role (Vasconcelos et al. 2011; Whitfield 2020; 2021a; Baker and Sheaves 2021) which underpins the resilience and persistence of coastal fisheries for numerous commercial species that depend on estuaries to complete their life cycles (Able 2005). Only when we understand the interactions between climate change, global change, and fishes in estuaries will we be able to restore and manage these systems in a manner that is beneficial for the ecology and productivity of the ichthyofauna, and for the coastal economies that depend on viable and healthy fish populations (Allison et al. 2009).

24.2 FISH GUILDS IN ESTUARIES At global and regional scales, the composition of fish assemblages in estuaries is shaped by assembly processes, namely dispersal limitation, environmental filtering, and biotic interactions operating at global, regional, and local scales (Figure 24.2, Henriques et al. 2017a). Understanding these processes is central to evaluating and responding to climate change. Patterns of taxonomic composition emerge at the global scale, i.e., among marine biogeographical realms. This is because large marine regions share

FIGURE 24.2  Hierarchical order of ecological processes that operate at (A) global, (B) regional, and (C) local scales to determine the composition of estuarine fish assemblages (after Henriques et al. 2017b).

477

evolutionary histories and exposure to common drivers (i.e., mainly water temperature and historical isolation), yield high levels of endemism, and have internally coherent biotas at high taxonomic levels (Spalding et al. 2007). Climate change may lead to changes in the distributions and population densities of estuary-associated fishes, but to understand how and where, it is necessary to understand the drivers behind estuarine fish population patterns at global, regional, and local scales. The species composition of biological communities is influenced by a hierarchy of ecological mechanisms, including dispersal limitation and environmental filtering. Dispersal limitation is the main predictor of fish species composition in estuaries at global and regional scales (Henriques et al. 2017b). Dispersal limitation results primarily from biogeographic barriers that isolate evolutionary processes (e.g., speciation and extinction) by limiting dispersal and thus giving rise to spatial differences in species composition among communities (Barton et al. 2013). In addition to marine biogeography which drove evolutionary processes associated with historical isolation and dispersal, the main drivers of taxonomic composition and richness of fish assemblages in estuaries are primarily the global temperature gradient, local connectivity with marine ecosystems, and secondarily due to local habitat size and suitability (Vorwerk et al. 2003; Vasconcelos et al. 2015; Henriques et al. 2017b). Fish species richness tends to decline in estuaries towards the poles compared to estuaries at lower latitudes (Blaber 2008). This pattern is likely related to rates of speciation and extinction that vary along latitudinal gradients and holds true for marine (Tittensor et al. 2010), freshwater (Tisseuil et al. 2013), and estuarine fish assemblages (Vasconcelos et al. 2015). The spatial boundaries of species’ physiological tolerances are highly changeable because estuaries are prone to changes in conditions at various spatial and temporal scales. Environmental instability does not favor the evolution of specialists (Whitfield 1994a), which are relatively lacking in estuaries compared to more stable marine environments. It follows that estuaries favor robust euryoecious species with wide environmental tolerances that can exploit a wide range of conditions; species richness in estuaries is, therefore, lower than in adjacent marine habitats (Wallace et al. 1984; Martino and Able 2003). As a result, estuarine fish communities are dominated by few but persistent “core species” with robust physiological tolerances (Schulte 2007) that should cope with climate change. The species composition of estuarine fish assemblages is primarily driven by physiological constraints associated with salinity (Kültz 2015; Whitfield 2015). Connectivity between estuaries and marine ecosystems has a major influence on the estuarine salinity regime and is an important driver of changes in species richness and composition, and can be measured via proxies such as estuary mouth morphology, open versus closed, including frequency, timing, and duration of opening (Whitfield 1999), mouth width (Nicolas et al. 2010), tidal range, and catchment

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run-off (Tweedley et al. 2016). Generally speaking, permanently open estuaries have greater species richness than temporarily closed estuaries due to their greater connectivity with adjacent habitats and less disrupted salinity gradients (Vorwerk et al. 2003; Vasconcelos et al. 2011). Estuaries prone to mouth closure (temporarily closed estuaries), usually occur in temperate semi-arid regions such as southern Australia and South Africa (Hoeksema et al. 2006; James et al. 2016) but may also be present in other regions (e.g., Behrens et al. 2013). Such estuaries typically have smaller surface areas, intermittent river flow, narrower mouth widths, and smaller drainage basins than permanently open estuaries (Vorwerk et al. 2003; Vasconcelos et al. 2015). Species richness and the representation of functional traits in faunal communities can impart resilience and stability on ecosystems, buffering them against stress such as climate change, although the resistance and resilience of constituent species and biotic interactions also contribute (Mouillot et al. 2013). Although redundancy in fish assemblages appears to be well established in species-rich tropical and warm-temperate estuarine systems, this may not be the case in cool-temperate estuaries (Whitfield and Harrison 2020), thus making these assemblages more vulnerable to climate and global change effects. By understanding the drivers behind taxonomic diversity and functional structure of ecosystems, and predicting how changes are likely to affect them, it may be possible to enhance the resilience of ecosystems to climate change and associated stressors, or to attempt to offset the impacts of particular stressors by reducing the impacts of stressors that are more easily controlled (Gillanders et al. 2021). Conservation management should therefore take a tiered approach to estuary management (Elliott et al. 2021a) that is alert to the different drivers of biodiversity and functional structure at different spatial scales and which, if addressed appropriately, may offer assistance in maintaining the resilience of estuarine habitats in the face of human-induced change such as climate change or other anthropogenic change such as pollution or habitat destruction (Moore and Jarvis 2008).

24.3 CLIMATE CHANGE DRIVERS FOR FISHES IN ESTUARIES The major climate change drivers need to be understood in order to predict how climate-induced changes may impact these ecosystems and their fish communities. In particular, the effects of climate change on fish in estuaries will likely stem mainly from changes to five main drivers, namely river flow, salinity regime, temperature increases, sea-level rise, and estuary connectivity. Other possible impacts on fishes include alterations to dissolved O2, CO2, pH, and an increase in diseases and parasitism in estuaries. Some of these changes are likely to also interact directly with anthropogenic perturbations as outlined in Table 24.1.

Climate Change and Estuaries

24.3.1 River Flow The functioning of estuaries relies on natural dynamics imposed on these systems by riverine and marine influences. The increasing human use of fresh water from both large and small river catchments in arid and semi-arid areas has had the effect of forcing some estuaries into artificial cycles (Tweedley et al. 2019; Moyle and Stompe 2021), that is, natural successions now have human imposed trajectories that are changing estuarine variability and forcing some systems into extreme states (Whitfield and Bruton 1989). Climate change may either exacerbate or reduce the above trends due to decreased or increased rainfall regimes in the different catchments (Gudmundsson et al. 2021). Droughts and reduced river flows into estuaries, particularly in semi-arid regions of the world, are increasing, and salinities beyond the tolerance of most estuary-associated species are becoming more common (Krispyn et al. 2021). Only two fish species (elongate hardyhead Atherinosoma elongata and flathead mullet Mugil cephalus) in the Beaufort Inlet (Australia) were able to withstand salinities above 80 and all species dependent on the zoobenthos within that estuary disappeared with the progressive loss of food resources and increasing salinity brought about by climate change and reduced river flow (Krispyn et al. 2021). The fish fauna of estuaries can also be influenced by reduced river flow in other ways, especially a decrease in fish recruitment, particularly by the juveniles of estuaryassociated marine species (Whitfield et al. 1994). The collapse in planktonic productivity brought about by reduced nutrient inputs from rivers negatively affects resident zooplanktivorous fishes and the survival of their larvae (Grange et al. 2000; Costalago et al. 2015), and the decreased olfactory cues entering the sea may no longer attract larval and juvenile marine fishes into adjacent estuaries also reducing the numbers of marine species entering these systems (Whitfield 1994b; Sullivan et al. 2006). Hypersaline conditions can result in both reduced species diversity and fish abundance (Zampatti et al. 2010). However, where estuaries lose their normal estuarine salinity gradient and become “arms” of the sea, there is often an increase in fish species diversity due to stenohaline marine taxa then entering and moving up these systems. Unfortunately, the gain in small numbers of marine stragglers is insufficient to compensate for the decline in estuarine-dependent fish that usually dominate these systems (Figure 24.3), and also the decline in primary and secondary productivity that accompanies the loss of nutrient and sestonic riverine inputs (Grange et al. 2000). Conversely, episodic river flooding often causes temporary decreases in both fish species diversity and abundance in estuaries due to a rapid decline in salinity, increased suspended sediments, reduced dissolved oxygen levels, and a collapse in the availability of pelagic and benthic food resources. However, the “resetting” of estuaries by episodic events is part of the essential cycle that maintains and enhances estuarine productivity and habitat diversity, as

479

Climate Change Effects on Fish Populations

TABLE 24.1 Table showing the ecological and functional effects that climate change, directly and synergistically with anthropogenic perturbations, will have on estuaries and their associated biota. The confidence level associated with each effect is provided. Driver Sea-level rises increase the occurrence and extent of seawater intrusions. Exacerbated by drought and anthropogenic modifications of drainage areas

Perturbation Increased salinisation

Effect

Confidence

Salinisation of estuaries projected to continue in response to sea-level rise, warmer temperatures and droughts under global warming greater than 1.5 ºC

High

Upstream expansion of brackish and marine communities, and a reduction in the diversity and richness of freshwater fauna. However, as the distribution of some species is related to habitat, a lack of such habitat can be a barrier to upstream shifts. This can cause a reduction in species richness in mid- to upper-estuarine areas and alter trophic dynamics

Medium

Increased salinisation and coastal inundation

Reduced extent and productivity of estuarine wetlands and tidal flats due to increased salinity, inundation and wave exposure. More marked in areas with limited capacity for soil accretion or inland migration due to coastal squeezing

High

Anthropogenic activities that inhibit sediment movement and deposition in coastal deltas increase the likelihood of their shrinking as a result of sea-level rise

Coastal inundation and reduce sediment deposition

Decrease in extent of coastal deltas

Medium

Increased air temperatures

Warming waters

Poleward migration of tropical and sub-tropical biota between estuaries. Loss of temperate species

Medium

Poleward expansion of tropical mangroves into saltmarsh

High

Increased air temperatures and altered freshwater discharge

Compounding effects of heat waves, hypersaline conditions, and increased turbidity and nutrient levels associated with floods

Negative changes in composition and biomass of seagrass species

High

Increased water temperature

Increased herbivory of seagrasses due to poleward migration of tropical herbivores

Negative changes in composition and biomass of seagrass species

Medium

Intensive anthropogenic activities have substantially increased nutrient and organic matter inputs into estuaries. Increased organic matter accumulation and warming lead to intensification of bacterial degradation and eutrophication

Warmer, eutrophic, and hypoxic waters

Increased occurrences of harmful algal blooms

High

Increased abundances of pathogenic bacteria (e.g. Vibrio spp.)

Low

Warmer, eutrophic, and hypoxic waters with increased stratification

Expansion of suboxic and anoxic areas

High

Mass mortality of fish and invertebrate communities

Medium

Large-scale climatic variations and extreme events

Fluctuations in salinity, turbidity, and nutrient gradients

Change in plankton phenology and composition

High

Low-frequency, high-intensity weather events (cyclones) and climatic extremes (droughts and heat waves)

Physical damage and temperature extremes

Loss of mangroves

High

Spatial variation in effects Climate effects more pronounced in high-latitude and temperate estuaries with limited exchange with the ocean

Medium

Impact of sea-level rise and precipitation more severe in shallow estuaries (< 10 m)

Medium

Estuaries with low tidal exchange and sediment supply more vulnerable

Medium

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Climate Change and Estuaries

FIGURE 24.3  Bubble plots showing the catch per unit effort (0.3–23 fish per haul) of the 10 most abundant fish species caught at 15 sites during hypersaline (autumn 2004) and estuarine (autumn 2014) conditions in the Kariega Estuary, South Africa (after Nodo et al. 2018). Benthic species are shaded white, benthopelagic species black, and pelagic species grey. Site 1 was near the mouth and Site 15 was near the head of the estuary.

for hurricanes (Valenti et al. 2020 and references therein). Recovery by estuary-associated fishes from such events is usually rapid and linked to a variety of factors, especially estuary morphometry which has a direct influence on the flushing or retention of estuarine biota (Marais 1982). However, a major increase in flash flooding of estuaries in a particular region due to climate change could result in declines in both fish species diversity and abundance in these systems. In conclusion, freshwater flows interact directly and indirectly with the fishes that inhabit estuaries (Vorwerk et al. 2008; Williams et al. 2017), e.g., river floods directly influence estuarine morphometry, water temperature, salinity, pH, turbidity, nutrient status, organic inputs, dissolved oxygen concentrations and olfactory cues; and indirectly affect mouth status, tidal prism, habitat diversity, primary and secondary productivity, fish recruitment, food availability, and competition. Depletion or removal of components of river flow to an estuary has major short- and long-term negative impacts on the ichthyofauna and associated fisheries (Loneragan 1999), some of which can be ameliorated by the provision of an environmental freshwater allocation that is appropriate to that particular system. It is also anticipated that changing river flow patterns as a result of climate change will affect estuary-associated fish species differently (Williams et al. 2017).

24.3.2 Salinity Regime Salinity is a key factor determining euryhaline fish species distribution and occurrence within an estuary (Whitfield et al. 2012). At the individual level, salinity influences osmoregulation, impacts metabolism, and triggers physiological and behavioral responses in fish. Therefore, abrupt changes in salinity may lead to relocation, but also to mortality if the magnitude of change and osmoregulatory stress surpasses physiological tolerance levels, and fish are unable to move out of an area quickly enough (Whitfield 2005). This is particularly the case in certain temporarily closed estuaries in semi-arid regions where salinities rise due to evaporation and the fish are prevented from escaping due to the lack of connectivity with either the river or sea (Krispyn et al. 2021). Indications are that marine fish species in estuaries are generally tolerant of salinities up to 60 and that most mass mortalities in estuaries occur in estuarine lakes and lagoons when escape from even higher hyperhaline conditions is not possible (Whitfield et al. 2006). Conversely, many of these marine fish taxa are intolerant of prolonged salinities below 2, especially when these oligohaline salinities are accompanied by very low or high water temperatures that compromise the metabolic processes of the trapped fishes. Small estuarine resident species appear to be more tolerant

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of low (oligohaline) rather than high (hyperhaline) conditions (Whitfield et al. 2006). Although freshwater fish species entering estuaries are tolerant of oligohaline and some mesohaline conditions, almost all species belonging to this guild are intolerant of polyhaline and euyhaline conditions (Whitfield 2015). Clearly changing estuarine salinity brought about by climate change is going to have a major impact on all three of the above fish guilds (Garcia-Seoane et al. 2016). In estuaries located in biogeographic regions expected to become warmer and drier, rising salinities will displace fish species with freshwater associations farther upstream and may also serve to interfere with the life cycles of diadromous species (Gillanders et al. 2021). Reduced freshwater inputs and increased evaporation may result in parts of estuaries drying out (Figure 24.4). Hypersalinity is overwhelmingly associated with negative impacts on estuarine fish communities and their food resources, with dramatic decreases in community structure and composition, potentially leading to species extirpation (Krispyn et al. 2021), as very few species can tolerate harsh osmoregulatory conditions (Whitfield et al. 2006). In contrast, in permanently open estuaries, marine stragglers will probably be able to penetrate farther into the affected estuaries, with several studies worldwide documenting the depletion of freshwater fish and a surge in marine species in estuaries associated with decreases in freshwater flow and increased salinity, particularly under drought conditions (Baptista et al. 2010). For larvae and juveniles of marine migrant species, freshwater is an important cue driving recruitment to estuarine nurseries (Sullivan et al. 2006; James et al. 2008a). Both extreme high and low flows may impose physical and chemical barriers that hinder the ingress of larval and juvenile stages, and therefore reduce estuarine colonization and connectivity. For instance, whilst a decrease in flow may reduce larval attraction, or create temperature barriers that inhibit estuarine colonization, increases in flow, linked to increased precipitation, may increase osmoregulatory stress, as well as the risk of larval and juvenile stages being flushed out to sea (Strydom et al. 2002). Moreover, variations in flow and estuarine salinity gradient may limit the extent of suitable habitat for juveniles and prompt area avoidance or fish emigration. Ultimately, changes in flow triggered by climate change may result in habitat compression, in particular along mesohaline habitats, which are widely recognized as key nursery areas in estuaries (Strydom et al. 2003). As changes to hydrological factors will necessarily affect habitats and food sources within estuaries, they are likely to have a range of indirect effects on estuarine fish. Variations in river flow may either disrupt or promote spatial connectivity in estuarine food webs, with the importance of autochthonous and allochthonous sources likely to change under climate change (Gillanders et al. 2021). Furthermore, estuarine fish communities may rely more heavily on nonterrestrial sources of carbon in food webs during drought conditions (Paterson and Whitfield 1997) that are likely to become more widespread in semi-arid areas under future climate change scenarios.

24.3.3 Temperature Changes Climate change effects on fishes are often interpreted relative to temperature change (Costa 1990; James et al. 2008b; Able and Fahay 2010). It is anticipated that estuaries and estuary-associated fishes will be affected by changes in both estuarine and ocean temperatures. A time series analysis of ocean heat content showed that the global trend is one of warming, with an increase of 0.1°C estimated to have occurred in the 0–700 m ocean layer for the period 1961 to 2003 (Solomon et al. 2007). Water temperature increases in shallow estuaries are likely to be much higher due to significantly higher atmospheric temperatures and increased heating of river water entering estuaries. There are also indications that estuary-associated tropical fish species such as Terapon jarbua may be better adapted to further

FIGURE 24.4  (a) A dead sparid Rhabdosargus holubi in a portion of the Seekoei Estuary (South Africa) that completely evaporated during a particularly prolonged drought. Salinities in excess of 90 prior to a drying out of most of the estuary caused mass fish mortalities to 10 other estuary-associated fish species in this system (Photo: Alan Whitfield). (b) An aerial drone view of the Hamersley Inlet (Western Australia) during December 2020 following extreme freshwater deprivation and the build-up of extensive evaporative salt structures across the lower reaches of the estuary. No fish were documented within the system under these conditions (Photo: James Tweedley).

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water temperature increases in littoral waters than temperate species (Eme et al. 2011). Temperature affects traits such as metabolic rates, spawning, migration, and development, and is regarded as the most important factor controlling the distribution and climate vulnerability of marine and estuarine fish species (Attrill and Power 2004; Morson et al. 2019). Temperaturedependent processes vary over a species’ latitudinal distribution (Henriques et al. 2017a), such that fish populations living at the edge of their species distribution may be more influenced by changes in temperatures than those living at the center of their distribution (Hampe and Petit 2005). In addition, it has been predicted that temperate fish species at the limit of their distribution will decrease in abundance due to global warming, and that tropical species will increase in abundance (James et al. 2016). For tropical species, winter survival is often the bottleneck in the establishment of populations in temperate areas (Henriques et al. 2007; Able and Fahay 2010). Elevated winter temperatures associated with climate change may allow these fish to overwinter and become established in previous temperate estuarine systems (Gillanders et al. 2021). As climate change accelerates and temperatures rise in estuaries, it can be expected that there will be marked changes in the composition of associated fish communities, resulting in new mixes of predators, prey, and competitors (Roessig et al. 2004; Howell and Auster 2012). Indications are that these changes can occur very quickly as evidenced by the arrival of new tropical species in the warm-temperate East Kleinemonde Estuary (South Africa) over a period of only one decade (Figure 24.5, James et al. 2008b) and in seagrass beds in the Gulf of Mexico over three decades (Fodrie et al. 2010; Gericke et al. 2014). In addition, it is very difficult to predict how fish communities will change in response to climate change, as each species responds

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differently to warming and assemblages are unlikely to shift their distribution as a unit (James et al. 2013). Further, temperature effects can be combined with changes in wind speed and direction to influence larval dispersal and estuarine recruitment (Lacroix et al. 2017).

24.3.4 Sea-Level Rise Sea-level rise will play a dominant role in shaping estuarine environments, not only because it will change seawater intrusion and drive changes to the tidal prism, but it will also be the main climatic factor dictating increases to estuarine depth and thus intertidal habitat gain or loss (Yang et al. 2015; Crosby et al. 2016). Sea-level rise will modify estuarine topography and, as water levels rise, marine water will naturally penetrate farther into estuaries (Figure 24.6). However, the direct impact of sea-level rise on estuarine fish is hard to quantify. Overall, numerous physical and chemical mechanisms in estuaries, including erosion/deposition cycles, sediment composition, changes to the extent of the photic zone, and biogeochemical cycles will be affected (Gillanders et al. 2021). Several studies have documented changes in estuarine geomorphology, salinity intrusion, habitat loss, habitat migration, and changes in animal and plant communities in estuaries linked to sea-level rise and associated coastal squeeze. A number have reported chronic decreases in productivity and loss of nursery habitat, though others show increases in seagrass beds and nursery function due to advancing marinization, or improved productivity and habitat quality linked to greater tidal flushing (Gillanders et al. 2021). For fishes in estuaries, the impact of sea-level rise will be underpinned by the capability of estuarine systems to continue to deliver key ecosystem functions. One key concern

FIGURE 24.5  A trend of increasing numbers of fish species recorded in the East Kleinemonde Estuary between 1996 and 2006, primarily due to the arrival of certain tropical taxa in this system (after James et al. 2008b).

Climate Change Effects on Fish Populations

FIGURE 24.6  An aerial view of the lower Columbia Estuary (USA). Increasing sea-level rise due to global warming will have a major impact on estuary morphometry, water depth, marine penetration up the estuary, and the locality and extent of intertidal areas and habitats (Photograph: Si Simenstad).

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regulate sand-bar formation and thus the frequency, timing, and duration of mouth opening (Figure 24.7). For instance, declining rainfall in some regions will likely promote extended closed periods between mouth openings which will significantly impact estuarine fish (Gillanders et al. 2021). Conversely, increased precipitation in other regions will lead to a shortening of the closed periods and facilitate greater connectivity between riverine, estuarine, and marine environments. A similar prerequisite exists for diadromous species to perform their migratory runs and complete their life cycles. Because of their dependence on freshwater-marine transitions at different stages of ontogenetic development, diadromous species face an uncertain future worldwide as climate change compounds a number of past, present, and future anthropogenic impacts (Lassalle and Rochard 2009; Limburg and Waldman 2009). Overall, increases in estuarine mouth closures will in general restrict connectivity and fish movement between fresh, estuarine, and marine environments, contributing to greater proportions of resident species, and to declines in the presence

regarding sea-level rise in estuaries is the potential alteration or loss of important habitats for fishes such as mangroves, saltmarshes, seagrasses, and oyster reefs (Kearney and Turner 2016; Borchert et al. 2018; Woodroffe 2018). This is especially true in systems where coastal development and urbanization impede the natural relocation of habitats in parallel with the displacement of optimum or suitable conditions along the estuarine gradient (Kirwan and Megonigal 2013; Raw et al. 2020). In other systems, such as in the Gulf of Mexico, increasing winter temperatures and sea-level rise are correlated with mangrove expansion and increasing fish diversity by tropical species.

24.3.5 Estuary Connectivity Fish responses to climate change are highly contextdependent, as they are influenced at large biogeographical scales, such as connections between estuaries and marine ecosystems, but also by local estuarine conditions, the fish guilds, or even the species being considered, and will differ depending on estuarine hydrogeomorphology (Hoeksema et al. 2009; Able et al. 2021). Estuaries can be generically classified as permanently or temporarily open, with the former having continuous access to the marine, estuarine, and freshwater environments. However, variations in physicochemical conditions have the potential to curb or enhance the ingress, permanence, or egress of different species and life stages, and create range shifts linked primarily to river flow, habitat availability, and species affinities to environmental conditions (Lauchlan and Nagelkerken 2019). The responses of fishes to climate-induced changes in temporarily open/closed estuaries will be less predictable, and will depend on the timing and duration of the mouth openings, and the effects of altered river flow on salinities. Changes to rainfall patterns, as well as sea-level rise, and stronger wave actions via increased storm surges, will all

FIGURE 24.7  The mouth region of temporarily open/closed estuaries in South Africa. (a) Increasing sea-level rise due to global warming and changing catchment rainfall patterns due to climate change will have a major impact on estuary mouth morphometry and the duration that the system is linked to the sea on an annual basis. (b) Increased coastal wave action associated with sea-level rise and storm activity will impact both small and large estuaries (Photographs: Alan Whitfield and Paul Cowley).

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of estuary-associated marine and diadromous taxa, linked in part to reduced larval recruitment and loss of estuarine nursery function (Gillanders et al. 2021).

24.3.6 Declining Dissolved Oxygen, Increasing Carbon Dioxide, and Lower pH Values Dissolved oxygen is a critical parameter that can vary on multiple spatial and temporal scales and manifest impacts on fishes from individuals to populations to ecosystems (Breitburg et al. 2009), from the tropics to temperate regions (Roessig et al. 2004). The solubility of oxygen in water is mediated by temperature, with elevated temperatures contributing to declines in oxygen storage capacity in estuarine waters (Doney et al. 2012). The interactive or synergistic effects of these parameters have direct impacts on fish reproduction, growth, and survival, and can act as water quality barriers for estuarine fish species (Pomfret et al. 1991) and alter the availability of invertebrate food resources. In general, the loss of oxygen from aquatic systems is referred to as “deoxygenation,” whereas waters with oxygen concentrations below specified thresholds are known as “hypoxic” (Breitburg et al. 2018). In the context of climate change, because water temperature controls oxygen solubility but also the metabolic demand of organisms, variations in temperature across ecosystems will drive the severity of hypoxia impacts on estuarine fish assemblages. Sensitivity of fish to low dissolved oxygen can also be influenced by species, water pH, diet quality, consumption rates, dissolved oxygen acclimation histories, and toxic metal exposure (Gillanders et al. 2021). Physiological processes, including metabolism, development, and growth, can all change in response to reduced oxygen, and many of these responses are additionally temperature-dependent (Marcek et al. 2019; Targett et al. 2019). Beyond those processes directly linked to metabolism, fish may be impaired in other fitness-relevant traits such as fecundity (Hassell et al. 2008). For instance, hypoxia exposure has been shown to interfere with normal reproductive development in estuarine-dependent fishes (Thomas et al. 2007). Episodic river flooding that brings anoxic bottom sediments into suspension, and results in the development of hypoxic conditions within the estuarine water column, can cause mortalities of fishes in estuaries. The impact of hypoxia on fish under such conditions is exacerbated by high silt loads that can cause clogging of the gills, therefore making the uptake of dissolved oxygen from the water column even more difficult (Whitfield and Paterson 1995). Estuarine systems may experience substantial changes in O2 and CO2 between day and night. This variation can lead to high levels of oxygen during the day associated with photosynthesis and low levels at night associated with respiration. This process is referred to as diel-cycling hypoxia (Lifavi et al. 2017) and can lead to direct mass mortalities or indirect mortality through enhanced predation on fishes in estuaries (Figure 24.8). The diel cycles in oxygen also cause considerable changes in CO2 leading

FIGURE 24.8  A fish kill in the temporarily closed Mdloti Estuary (South Africa) caused by high biological oxygen demand due to allochthonous inputs and a smothering of open water areas by the alien floating pond weed Salvinia molesta and water hyacinth Pontederia crassipes that prevented surface oxygen diffusion (Photograph: Nicolette Forbes).

to increased pCO2 and decreased pH (Miller et al. 2016). Eutrophication and ocean acidification exacerbate these patterns (Gillanders et al. 2021). Behavioral experiments suggest that ocean acidification may affect the sensory functions of fish, thereby potentially affecting a fish’s ability to orientate, including the detection of estuarine cues (Leis 2018). Interestingly, it was suggested that fish may then conform to a passive-dispersal paradigm and that less self-recruitment, increased dispersal differences, and decreased settlement rates may occur (Leis 2018). Only a single study has demonstrated that preference for different temperature and salinity cues (associated with estuarine waters) may change associated with ocean acidification (Pistevos et al. 2017). If fish recruiting to estuaries cannot sense the temperature and salinity gradients associated with an estuary this is likely to impact population replenishment and alter connectivity patterns (Gillanders et al. 2021). In addition, Nilsson et al. (2012) have indicated that changes in carbon dioxide levels brought about by climate change can also alter fish behavior by interfering with neurotransmitter function.

24.3.7 Spreading Diseases and Parasites As water provides an ideal medium for the transmission of disease and parasites, it is not surprising that these interactions are central to estuarine fish growth and survival, especially as a result of climate change effects (Burge et al. 2014; Woo et al. 2020). Temperature is likely to have the greatest effect on disease and parasite infections in fish due to its effects on the distribution and physiology of pathogens, parasites, and fishes (Marcogliese 2008). For example, during long-term monitoring of the mouth of the Rhine Estuary (Europe) from 1961–2005, Vibrio spp., which prefers warm

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temperatures and low salinities, were in greatest abundance during warm years (Vezzulli et al. 2013). Similarly, the parasite Glugea stephani increased its infection rate on English sole Parophrys vetulus in the temperate Yaquina Bay (USA) in association with warmer estuarine temperatures during El Niño years (Olson et al. 2004). Temperature increases enhance dermal myco-bacteriosis in Morone saxalitis in Chesapeake Bay (Groner et al. 2018) and can increase the ectoparasite sea louse (Lepiophthei salmonis) in salmonids based on modeling approaches (Groner et al. 2014). Further, warming water temperatures during the winter may increase the severity of parasite-inflicted damage in some fishes (Klemme et al. 2021). The bacterial pathogen Aeromonas salmonicida, which typically infects salmonids and other freshwater-estuarine species with ulcerative furunculosis, made its first appearance in the James Bay region (Canada) when water temperatures exceeded 15°C in the 1990s (Tam et al. 2011). However, the influences of warming are unlikely to be uniform. In cooler regions, warming may lead to influxes of new species and increased rates of growth and transmission. However, in warm regions, further warming may lead to decreases in host lifespan and biodiversity, thus leading to reductions in pathogen and parasite communities (Lafferty 2009). When eutrophication exceeds certain limits, resulting in algal proliferation, cyanobacteria blooms, and hypoxia (Paerl and Huisman 2009), it becomes a stressor that increases fish susceptibility to diseases and parasite infection. For example, in Chesapeake Bay, infection of mummichogs Fundulus heteroclitus by nematodes of the genus Eustrongylides was highest in eutrophic areas of the system and more prevalent as salinity decreased (Weisberg et al. 1986). This further illustrates the often complex interplay between environmental variables that are likely to be influenced by climate change and the capacity of disease or parasites to exploit favorable conditions that may arise in estuarine fish communities.

24.4 GLOBAL EXAMPLES OF CHANGING ESTUARINE FISH POPULATIONS The impacts of climate and global change in any given estuary will vary with context because the responses are linked to overarching biogeographic realms and how large-scale patterns drive changes to local physical, chemical, and ecological processes (Able et al. 2021). Just as environmental filtering governs the patterns of taxonomic diversity in estuarine fish communities (Henriques et al. 2017b), the relative impacts of global changes on estuarine fishes will not be identical or equally distributed worldwide. Moreover, continental, regional, and local features and processes can impact what may be similar estuaries differently. But one key aspect to consider is that the responses of local fish communities will be determined by the sensitivity/tolerances of individual species and the estuarine habitats in which they occur, as well as their adaptability and scope to respond and persist within each estuary (Gillanders et al. 2021).

24.4.1 Temperate Northern Atlantic and Northern Pacific Estuaries The northeastern Atlantic, from the British Isles to the Iberian Peninsula, encompasses a wide temperature gradient, from cool-temperate to warm-temperate biogeographic regions, with areas in the southern Iberian Peninsula presenting a semi-arid, Mediterranean-type climate with reduced precipitation and hot summers. Throughout the region, sea shelf warming is expected to continue to rise, with climate projections predicting changes in regional hydrological processes, such as increasingly clustered rainfall events in the United Kingdom, or increasing drought frequency in the Iberian Peninsula, modulated by variations in large-scale climatic patterns such as the North Atlantic Oscillation (NAO) (Trigo et al. 2004; Robins et al. 2016). Indeed, the NAO in particular has been shown to be the most important climatic forcing parameter explaining variations in fish assemblage composition, as well as juvenile marine fish abundance and growth during estuarine residency (Attrill and Power 2002). Over recent decades, rising water temperatures have led to increasing occurrences of subtropical ichthyofauna along the latitudinal gradient of the North Atlantic coast. As these fish species overcome overwintering bottlenecks (Figueira and Booth 2010), estuarine communities are mirroring adjacent marine environments and gradually showing increased numbers of traditionally lower latitude species, whilst other cool-water species are displaced poleward (e.g., Cabral et al. 2001; Genner et al. 2004, Hermant et al. 2010; Baptista et al. 2015). A broad range of climate-induced biogeographical shifts in estuarine fish assemblages along the northeastern Atlantic coast from Portugal to Scotland is evident (Nicolas et al. 2011). Comparisons of mean latitudinal species distributions supported the northward migration of estuarine species over 30 years (1970s–2010s) (e.g., common eel Anguilla anguilla), with trends further corroborated by the occurrence of numerous subtropical species above historical latitudinal limits (e.g., round sardinella, Sardinella aurita). This increase in warm-temperate and tropical species, and retreat of cool-temperate species has been evident along the Portuguese coast for species that use estuaries opportunistically, as well as among estuarine residents and species that use estuaries as nursery areas (Cabral et al. 2001; Vinagre et al. 2009). The same is occurring in the Dutch Wadden Sea for European sea bass Dicentrarchus labrax (Cardoso et al. 2015). It is also been documented for the salmon Salmo salar along the Spanish Atlantic coast (Nicola et al. 2018). In addition to temperature, changes to river flow will play a pivotal role in shaping fish assemblages in temperate Atlantic estuaries, as in other regions around the globe. For the Gironde Estuary (France), Nicolas et al. (2010) highlighted how shifts in environmental conditions significantly affected the abundance of a variety of species, with the enhanced presence of marine stragglers in periods of

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reduced river flow. In contrast, in the outer Severn Estuary (UK), total fish species declined with increasing salinity (Henderson 2007). Overall, numerous studies have shown how the occurrence and abundance of fishes in estuaries is related to environmental variables, including river flow and regional scale climatic patterns (Chícharo et al. 2006; Costa et al. 2007; Pasquaud et al. 2012). These studies provide a clear indication of the potential responses of North Atlantic estuarine fish assemblages to environmental change. Whilst general trends follow Henriques et al. (2017a), these trends are still subject to species- and site-specific responses according to physiological scope, estuarine conditions, habitat availability, or estuarine geomorphology (Le Pape et al. 2003; Dolbeth et al. 2010; Vasconcelos et al. 2011). Ultimately, recruitment and species persistence can differ among estuaries and are dependent on the interplay between influencing factors. A growing concern is how climate change, and in particular alterations brought by extreme events (e.g., droughts, floods), may affect the nursery-role function and the contribution of European estuaries to coastal areas where many key commercially important marine species are sustained by juveniles that utilize estuarine habitats (Reis-Santos et al. 2013; Tanner et al. 2013). Studies in numerous estuaries have shown how drought and reductions in river flow impact estuarine fish assemblages, and can compromise the estuarine nursery role via a decrease in the abundance of juveniles (Chícharo et al. 2006; Costa et al. 2007; Martinho et al. 2007; Gillson et al. 2009; Baptista et al. 2010). This reduction is often linked to reduced river plumes and cues for larval recruitment, as well as sub-optimal environmental conditions or food availability (Vinagre et al. 2009). Overall, production in drought-affected estuaries decreases with decreasing river flow but patterns are not always consistent (Dolbeth et al. 2010). In the Mondego Estuary, severe drought reduced juvenile densities and thus the nursery role of the system, but, as generally expected, overall species richness increased. However, in the Tagus Estuary, (Portugal), Costa et al. (2007) recorded increased densities of estuarine species as well as of those that use the estuary as a nursery habitat in drier years. This resilience is likely due to the size and geomorphology of the larger Tagus, which enables species to move elsewhere compared to the smaller Mondego Estuary (Portugal). Similarly, in the large Gironde Estuary (France), a combination of temperature-induced shifts in species distributions, associated with decreases in river flow and raised salinities, increased juvenile abundances, and therefore Pasquaud et al. (2012) argued that climate change may positively influence this estuary’s overall nursery function. Robins et al. (2016) summarised the main climatic drivers and expected impacts of climate change in UK estuaries. The anticipated increase in river flows in winter may enhance high turbidity and hypoxia events, potentially deterring larval recruitment or leading to the flushing out of fishes from within these estuaries. Linked to changes in river flow and consequent variations in the inputs of

Climate Change and Estuaries

allochthonous materials and indirect effects on prey species, estuarine food webs that can play determinant roles in shaping the persistence of many fishes within estuaries are also likely to be affected. One key issue in UK estuaries, and other regions of the North Atlantic, is the concomitant effects of climate change–related sea rise with land sinking, which results in a greater relative sea-level rise in the region, and therefore increased estuarine inundation risk together with changes to geomorphology and consequently habitat availability. Conversely, in other areas (e.g., Scotland, Sweden) isostatic rebound dampens this effect (Shennan and Horton 2002) and could even lead to a shallowing and loss of water area in some estuaries. The temperate continental and estuarine waters of the western North Atlantic, from Cape Cod to Cape Hatteras, share some of the same responses to a changing climate as those in the eastern North Atlantic. This area, often referred to as the mid-Atlantic Bight, has some of the widest seasonal temperature ranges of anywhere in the world (Parr 1933; Grosslein and Azarovitz 1982; Cook 1988). As a result, the fish fauna, both estuarine-dependent and estuarine-associated species, have pronounced seasonality in their use of estuaries (Able and Fahay 2010). An overarching effect is that the waters of the region, including the midAtlantic Bight, are warming (Shearman and Lentz 2010; Pinsky et al. 2013; Forsyth et al. 2015) and some analyses suggest that this rate may be three times as fast as the global average (Saba et al. 2015) and is having a major impact on fisheries in coastal waters (Pershing et al. 2015). Fish populations on the continental shelf in the region, including the juveniles and adults of estuarine fishes, are exhibiting poleward shifts with an expansion of the range of some species (Nye et al. 2009, 2011; McBride et al. 2018; Morley et al. 2018; Kleisner et al. 2016). This pattern is also true for larval fishes in the same region (Walsh et al. 2015). As expected, this kind of response is also evident in the long-term pattern in adjacent estuaries. For example, larval ingress at a single estuary in southern New Jersey has demonstrated poleward shifts in recent decades (Morson et al. 2019). In this study, over a 26-year period of weekly sampling, all five species with northern affinities declined in abundance while 18 of 21 species with southern affinities increased. This same pattern is evident for the distribution of some species with northern affinities such as rainbow smelt Osmerus mordax, Atlantic tomcod Microgadus tomcod, and winter flounder Pseudopleuronectes americanus which are disappearing from some estuaries in the southern part of their range (Able and Fahay 2010; Able et al. 2014). Others are becoming less abundant such as pollock Pollachius virens and threespine stickleback Gasterosteus aculeatus, i.e., the “rear edge” of their distribution is changing as well, which is just as important (Hampe and Petit 2005). For some species, these range expansions and increases in population size are resulting in new fisheries such as that for Atlantic croaker Micropogonias undulatus in the mid-Atlantic Bight in response to warming winter water

Climate Change Effects on Fish Populations

temperatures (Able and Fahay 2010), and reduced winter mortality in estuaries (Hare and Able 2007) even when the effects of the fishery are taken into consideration (Hare et al. 2010). The decreasing role of winter mortality with warmer temperatures in estuaries is reflected in other estuaries in the region as well (Hurst 2007; Wuenschel et al. 2012). The adults of some taxonomically diverse southern species are also responding to warming temperatures in the region as they become more frequently collected and abundant, e.g., the rhinopterid (Rhinoptera bonasus), clupeid (Opisthonema oglinum), cottid (Myoxocephalus aenaeus), sciaenid (Pogonias cromis), and a tetradontid (Sphoeroides maculatus) (Able and Fahay 2010). Other southern species have become residents such as a gobiesocid (Gobiesox strumosus) and several gobiids (e.g., Ctenogobius boleosoma, Microgobius thalassinus) (Able and Fahay 2010). The indirect effects of changing temperatures associated with a changing climate might be expressed in several ways (Able and Fahay 2010). First, milder fall and early winter temperatures may provide additional time for the juveniles of southern species to migrate to a thermal refuge. Milder temperatures could also provide a thermal refuge of greater spatial extent and thus be easier to reach. This may allow a shorter retreat in the fall and winter and allow them to migrate farther north for spawning in the spring. Other climate-associated factors may also influence estuarine fishes. On the east coast of the USA, winters characterized by high levels of precipitation typically lead to elevated levels of freshwater runoff and thus river flow (Schmidt and Luther 2002) which in turn is known to influence the ingress of glass eels that follow chemical odors (McCleave and Jellyman 2002; Sullivan et al. 2006). Furthermore, models of climateinduced changes in the location of the Gulf Stream suggest that the abundance of Paralichthys dentatus in this region will also be influenced (O’Leary et al. 2019). Another of the potential climate change–induced changes is that evidence for pronounced relative sea-level rise in the western North Atlantic has been occurring over at least three centuries (Gehrels et al. 2020) and is generally expected to continue into the future (Schuerch et al. 2018). Some studies also suggest that other climate-related variables, such as salinity, sea surface height, and seabed characteristics, are also responsible (McHenry et al. 2019). The rate of sea-level rise is evident at local levels as well. Tide gauge data from southern New Jersey beginning in 1912 indicates that sea-level rise has been occurring since that time and, more importantly, that it is about twice the global average because of the additive effects of global sea-level and land subsidence in the region (Sallenger et al. 2012; Miller et al. 2013). As a result, estuarine intertidal habitats, such as salt marshes, are particularly susceptible to sea-level rise (Gedan et al. 2009; Crosby et al. 2016). This is important to fishes because many species rely on these productive marshes for food, growth, and survival to maturity (Able and Fahay 2010; Able et al. 2021), and they, in turn, provide food for fishes or serve as nurseries that are the basis for fisheries (Baker et al. 2020). Future projections

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suggest that relative sea-level rise in the region will result in widespread estuarine salt marsh loss (Watson et al. 2017a, 2017b; Able 2020). Pacific coast estuaries respond to both inland and oceanic climate variability (Reum et al. 2011; Feyrer et al. 2015). The three largest estuaries – Puget Sound, San Francisco, and Columbia – have large inland catchments that encompass and integrate a broad expression of climate patterns. There is also extensive human alteration and degradation of the estuaries and their catchments, which can force systems into novel ecological regimes and impact their capacity to respond naturally to climate variability or climate change (Borde et al. 2003). The San Francisco Estuary is the largest (by historical surface area) of the temperate Pacific Coast systems and represents the southern range limit of multiple marine and anadromous species that use the estuary. The system has been highly altered by human activities associated with agricultural and municipal development (Nichols et al. 1986). Native fish populations have plummeted as these changes have occurred, resulting in 14 migratory or resident fishes now being protected under state or federal endangered species laws. These declines have stimulated rigorous scientific investigations to understand the underlying mechanisms, including the effects of climate (Sommer et al. 2007; Cloern et al. 2011). Notably, the effects of climate variability on San Francisco Bay Estuary fishes and primary productivity are apparent despite the presence of such strong local internal drivers and perturbations that could confound them (Lehman 2004; Cloern et al. 2010; Feyrer et al. 2015). Projections from down-scaled global climate models suggest that the San Francisco Estuary will likely experience increased water temperature, sea-level rise, salinity intrusion, changes in the amount and timing of freshwater inputs, and increased average annual water-year outflows (Knowles and Cayan 2002, 2004; Cloern et al. 2011). These changes will potentially affect fishes in many indirect and direct ways. Anticipated rates of sea-level rise will likely outpace the rate at which extant wetlands can adjust and may contribute to further loss of wetland habitat to support fishes (Thorne et al. 2018). Changes in the timing and magnitude of freshwater inflows can affect the inundation of floodplains important to the reproduction of some fishes (Cloern et al. 2011). Changes in water temperature may change the suitability of habitat, the timing of critical life history stages, and fish behavior (Pankhurst and Munday 2011; Brown et al. 2013, 2016; Davis et al. 2019). Generally, these potential broad-scale changes to the system are likely to compress in time and space the types of physico-chemical habitat favorable for supporting native fish species while promoting conditions that favor non-natives (Moyle 2008; Brown et al. 2013, 2016). Implementing effective climate change adaptation strategies to conserve and manage fishes and fisheries, while also addressing human water needs in the already highly altered San Francisco Estuary, will be a considerable challenge for resource managers.

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24.4.2 Tropical Atlantic and Indo-Pacific Estuaries Climate-related events in South America have become more intense and more frequent over the last few decades. Shifts in the distribution and abundance of shallow water fishes on the coast of southeastern Brazil and Argentina have occurred in response to warming temperatures (Aranjo et al 2018). In northeastern Brazil, 16 out of 25 years between 1991 and 2016 experienced drastic reductions in precipitation (Marengo et al. 2018). The meteorological and oceanic mechanisms that lead to circulation and rainfall changes responsible for drought in northeastern Brazil include the occurrence of El Niño, and/or a latitudinal shift of the Intertropical Convergence Zone (ITCZ) driven by major forces from the Northern Hemisphere climate (Utida et al. 2019). Since 2010, the semi-arid portion of northeastern Brazil has been experiencing one of the most protracted and intense droughts in decades (Pereira et al. 2014; Erfanian et al. 2017). In this region, the climate-physiographic properties of river basins vary greatly from estuarine and adjacent coastal zones to their headwaters. Lower parts of basins are dominated by a humid climate associated with tropical forest ecosystems and habitats, while upper basins are characterized by semi-arid climates and habitats (Costa et al. 2016). For example, two-thirds of the Mamanguape River basin (north-eastern Brazil) is under the influence of a semi-arid climate, which causes an intermittent flow regime in most of the basin; but the 25-km long tidal estuary is perennial and influenced by smaller tributaries from wetter climate areas (da Silva et al. 2018). Relationships between fish abundance and freshwater flow in the Mamanguape Estuary have also been found in adjacent coastal waters, thus giving clues regarding future climate change effects. For example, a large rainfall event in 2011 contributed to the “estuarization” of the adjacent beaches as a result of increased river discharges (Oliveira and Pessanha 2014). This expansion of the estuarine zone into coastal areas promoted a shift in coastal fish assemblages and, during this period, the lowest abundance of resident marine fishes typically found near sandy beaches was recorded. In contrast, low precipitation in 2012 was insufficient to reduce salinity in the lower reaches of the estuary or expand the estuarine area into the coastal area, thus resulting in a higher abundance and intrusion of marine species into this zone. The increase in marine taxa was explained by the more euhaline conditions in 2012, with this process of “marinization” occurring in numerous estuaries globally as climate change began to accelerate (Oliveira and Pessanha 2014). Overall, the influence of rainfall on salinity and its effects on other environmental variables regulate the composition and distribution of the fish assemblages in the Mamanguape Estuary (Brazil). This emphasizes the importance of seasonal changes in freshwater discharge for larvae and juvenile fishes, with salinity acting as the main environmental filter. Predicted changes in global climate for the

Climate Change and Estuaries

region, and the effects of long-term droughts on recruitment processes and the distribution of fish populations have the potential to affect both fish assemblages directly, as well as indirectly via the loss of estuarine habitats, which may ultimately lead to a decrease in estuary-associated species (Dolbeth et al. 2016). St Lucia, a 35,000 ha estuarine lake system, is located in the tropical/subtropical transitional zone on the southeast African coast. This Ramsar and World Heritage Site has come under pressure from both climate and global change impacts over the past century, with the very existence and functioning of the system for fishes coming under real threat in the last few decades (Vivier et al. 2010). Although this dynamic estuarine lake system had a natural variability that ranged from hyperhaline to oligohaline depending mainly on prevailing climatic conditions, human intervention, and especially the separation of the Mfolozi River system from the St Lucia mouth region, has made this variability more extreme (Whitfield et al. 2013). The consequences for the lake biota, especially the fishes, and the ecology of the system have been devastating (Cyrus et al. 2011). An almost two-decade-long drought following the turn of the century, possibly related to climate change, triggered a management-driven reconnection of the Mfolozi River to the St Lucia Estuary and the subsequent refilling of the lake (Figure 24.9). The full reconnection of the Lake St Lucia system with the sea in January 2021 will hopefully

FIGURE 24.9  (a) Landsat photo of the 350 km2 Lake St Lucia system during an almost two-decade-long drought cycle and (b) the same system following reconnection with the Mfolozi River and a breaking of the drought (see text for details).

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Climate Change Effects on Fish Populations

allow the lake to be recolonised by marine fish species that are usually the dominant ichthyofaunal component within this system (Whitfield 2021b). Fortunately, a sea-level increase linked to climate change may compensate, to some degree, for the accelerated shallowing of the system due to increased sedimentation rates from degraded catchments over the past century (Jones et al. 2020). Estuaries form significant ecological zones for many fish species in the northern tropical Indo-Pacific region. Rivers flowing through these zones are characterized by high monsoonal rainfall followed by extended dry seasons with limited outflows (Dudgeon 2005). These systems thus tend to have high monsoonal sediment loads which give rise to extensive delta systems that provide a rich transition from freshwater to seawater, and support productive fisheries comprising species with a diverse array of life-history strategies (Nguyen et al. 2002). Climate change is expected to affect southeast Asia in several ways, including temperature increases, more intense rainfall events (including cyclones), longer drought periods, and sea-level rise. For instance, existing direct threats to anadromous fish in Myanmar include overexploitation of fish stocks and habitat loss, and indirect impacts through climate change (Rao et al. 2013). The proliferation of polders and sluices to combat sea-level rise and protect cropping areas will impede anadromous fish such as the Hilsa shad (Tenualosa ilisha) from being able to enter freshwater habitats to complete their life cycles (Conallin et al. 2019). The Mekong River (Asia) is one of the top 10 river systems in the world in terms of discharge. Among the 1,200 fish species reported for the Mekong is the anadromous Krempfi catfish Pangasius krempfi (Hogan et al. 2007). Each year thousands of adult fish, many more than a meter in length, migrate from the South China Sea some 600 km upstream to spawn. The provision of fish passage in both directions is an important consideration, both to allow upstream adult migrations to occur and for larvae to reach the sea and mature (Dugan et al. 2010). From a climate change perspective, long-term changes to hydrology, catchment run-off, or sea water levels may alter the fundamental dynamics of migration routes for these catfish. Therefore, management strategies to protect Krempfi catfish migrations should consider future climate change scenarios in addition to the obvious challenges posed by river infrastructure (Gillanders et al. 2021). The above three examples from South America, South Africa, and Indonesia highlight the potential challenges in response to climate and global change, with emphasis on maintaining freshwater, estuarine and marine connectivity in estuaries within the Indo-Pacific region. There is presently an unprecedented boom in hydropower, irrigation, dam and barrage construction. However, it is important that these river development projects adequately consider potential impacts on fisheries resources and how development may interact with impacts derived from climate change.

24.4.3 Temperate Southern Atlantic and Southern Indo-Pacific Estuaries Temperate southern hemisphere estuaries tend to have very different climatic and tidal regimes when compared to temperate northern hemisphere estuaries. The extensive (3.5 x 104 km2) Río de la Plata is the largest estuarine system in temperate South America (Barletta et al. 2010). This estuary extends 320 km inland and varies in width from 20 km (head) to 220 km (mouth), with an average depth of