COASTAL WETLANDS AN INTEGRATED ECOSYSTEM APPROACH
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COASTAL WETLANDS AN INTEGRATED ECOSYSTEM APPROACH
Edited by GERARDO M. E. PERILLO ERIC WOLANSKI DONALD R. CAHOON MARK M. BRINSON
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Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, The Netherlands Linacre House, Jordan Hill, Oxford OX2 8DP, UK First edition 2009 Copyright 2009 Elsevier B.V. All rights reserved No part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means electronic, mechanical, photocopying, recording or otherwise without the prior written permission of the publisher Permissions may be sought directly from Elsevier’s Science & Technology Rights Department in Oxford, UK: phone (+44) (0) 1865 843830; fax (+44) (0) 1865 853333; email:
[email protected]. Alternatively you can submit your request online by visiting the Elsevier web site at http://elsevier.com/locate/permissions, and selecting Obtaining permission to use Elsevier material Notice No responsibility is assumed by the publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. British Library Cataloguing in Publication Data A catalogue record for this book is available from the British Library Library of Congress Cataloging-in-Publication Data Coastal wetlands : an integrated ecosystem approach/edited by Gerardo M. E. Perillo . . . [et al.]. — 1st ed. p. cm. Includes bibliographical references and index. ISBN 978-0-444-53103-2 1. Estuarine restoration. 2. Coastal zone management. 3. Wetland ecology. 4. Coasts. I. Perillo, G. M. E. (Gerardo M. E.) QH541.5.E8C63 2009 577.69—dc22 2008046508 ISBN: 978-0-444-53103-2 For information on all Elsevier publications visit our website at elsevierdirect.com Printed and bound in The Netherlands 09 10 11 12 10 9 8 7 6 5 4 3 2 1
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CONTENTS
Preface
xix
List of Contributors
xxi
List of Reviewers
1
Coastal Wetlands: A Synthesis
xxxi
1
Eric Wolanski, Mark M. Brinson, Donald R. Cahoon and Gerardo M.E. Perillo 1 2
Introduction A Synthesis of Coastal Wetlands Science 2.1 Geography 2.2 Geomorphology evolution under climate change 2.3 The influence of vegetation on the geomorphology evolution with climate change 2.4 The stabilizing role of vegetation 2.5 State change and coastal evolution 2.6 The role of physical disturbances 2.7 The role of herbivores 2.8 Observations across ecosystem types 2.9 The human impact 2.10 Modeling and predictions 2.11 Coastal wetland ecosystems as a component of estuaries 2.12 Coastal wetland socioeconomics 2.13 Coastal wetlands are essential for our quality of life 3 Lessons from the Chapters in this Book 3.1 Coastal wetlands as ecosystems 3.2 Physical processes 3.3 Tidal flats 3.4 Marshes and seagrasses 3.5 Mangroves 3.6 Coastal wetland restoration and management 3.7 Coastal wetland sustainability and landscape dynamics References
Part I 2
Coastal Wetlands as Ecosystems
The Morphology and Development of Tropical Coastal Wetlands
1 2 3 4 6 9 13 14 18 21 28 31 32 34 38 43 44 45 47 48 51 53 56 57
63 65
Colin D. Woodroffe and Gareth Davies 1 2
Introduction Mangrove and Associated Wetlands
65 66
v
vi
3
Contents
3 Sedimentation and the Development of Wetlands 4 Sea-Level Controls on Wetland Development 5 Sea-Level Change and the Diversification of West Indian Mangroves 6 Sea-Level Change and the Evolution of Mangrove Habitats in the IWP 7 Impact of Future Climate and Sea-Level Change 8 Summary and Concluding Remarks References
69 71 74 76 80 82 83
Temperate Coastal Wetlands: Morphology, Sediment Processes, and Plant Communities
89
Paula D. Pratolongo, Jason R. Kirby, Andrew Plater and Mark M. Brinson
4
1 2
Introduction Factors Controlling Sediment Dynamics 2.1 The ‘‘ramp’’ model of salt marsh accretion 2.2 The ‘‘creek’’ model of salt marsh accretion 2.3 Storms and salt marsh erosion 3 Factors Controlling Patterns of Vegetation 3.1 Zonation of vegetation 3.2 Ecological development 4 Geographic Variation 4.1 Northern Europe 4.2 Eastern North America 4.3 Western North America 4.4 Mediterranean 4.5 Eastern Asia 4.6 Australasia 4.7 South America 5 Human Impact and Climate Change 5.1 Human impact 5.2 Climate and sea-level change 6 Summary References
89 91 92 93 94 95 95 96 97 97 98 103 104 104 105 107 108 108 110 111 112
Polar Coastal Wetlands: Development, Structure, and Land use
119
I. Peter Martini, Robert L. Jefferies, R. I. Guy Morrison and Kenneth F. Abraham 1 2 3 4 5 6 7
Introduction Geology/Geomorphology Oceanography Climate Structure of Coastal Wetlands Vegetation of Polar Coastal Wetlands Fauna of Polar Coastal Wetlands 7.1 Invertebrate fauna 7.2 Vertebrate fauna using coastal wetlands
119 121 124 125 127 133 136 137 140
Contents
8 Environmental Hazards 9 Conclusions and Research Priorities References
Part II Physical Processes 5
Intertidal Eco-Geomorphological Dynamics and Hydrodynamic Circulation
vii
147 148 149
157 159
Andrea D’Alpaos, Stefano Lanzoni, Andrea Rinaldo and Marco Marani
6
1 2
Introduction Intertidal Eco-Geomorphological Evolution 2.1 Poisson hydrodynamic model 2.2 Model of channel network early development 2.3 Model of marsh platform evolution 3 Results 4 Discussion 5 Conclusions Acknowledgments References
159 163 164 165 168 171 177 178 179 179
Tidal Courses: Classification, Origin, and Functionality
185
Gerardo M.E. Perillo
7
1 Introduction 2 Proposed Tidal Course Classification 3 Geomorphology of Tidal Courses 4 Course Networks and Drainage Systems 5 Origin of Tidal Courses 6 Course Evolution 7 Summary Acknowledgments References
185 187 189 196 198 203 206 206 206
Heat Energy Balance in Coastal Wetlands
211
Marı´ a Cintia Piccolo 1 Introduction 2 Mid-Latitudes 3 Low Latitudes 4 High Latitudes 5 Summary Acknowledgments References
211 215 220 222 225 226 226
viii
8
Contents
Hydrodynamics and Modeling of Water Flow in Mangrove Areas
231
Yoshihiro Mazda and Eric Wolanski 1
9
Introduction 1.1 Peculiar hydrodynamics in mangrove area 1.2 Material dispersion 1.3 Holistic system 2 Physical Characteristics of Mangrove Topography and Vegetation 2.1 Classification of mangrove topography 2.2 Bottom condition of mangrove swamps 2.3 Influence of the vegetation on the hydrodynamics 3 Peculiar Hydrodynamics in Mangrove Areas 3.1 General equations that control water flow 3.2 Timescales of flow system 4 Modeling 4.1 Hydraulic model 4.2 Material dispersion model 4.3 Ecosystem model as the holistic system 5 Summary Acknowledgments References
231 233 233 233 233 234 236 236 240 240 241 245 245 251 255 257 258 258
Mathematical Modeling of Tidal Flow over Salt Marshes and Tidal Flats with Applications to the Venice Lagoon
263
Luigi D’Alpaos, Luca Carniello and Andrea Defina 1 Introduction 2 Wetting and Drying, and the Dynamics of Very Shallow Flows 3 Wind and Wind Waves 4 Salt Marsh Vegetation 5 Salt Marshes and Tidal Flats Morphodynamics 6 Conclusions Acknowledgments References
263 265 272 279 281 285 286 286
Part III Tidal Flats
293
10 Geomorphology and Sedimentology of Tidal Flats
295
Shu Gao 1 2 3
Introduction Basic Conditions for the Formation of Tidal Flats Zonation in Sedimentation and Flat Surface Morphology 3.1 Vertical sediment sequences 3.2 Sediment and morphology on intertidal mud flats 3.3 Sediment and morphology on mixed sand–mud flats 3.4 Sediment and morphology on sand flats
295 297 298 298 300 303 305
Contents
4
11
ix
Factors and Processes 4.1 Influences of quantity and composition of sediment supply 4.2 Sedimentation during tidal cycles 4.3 Long-term accretion–erosion cycles 4.4 Tidal creek systems 5 Summary Acknowledgments References
305 305 308 310 311 311 312 312
Intertidal Flats: Ecosystem Functioning of Soft Sediment Systems
317
David M. Paterson, Rebecca J. Aspden and Kevin S. Black 1 2
Introduction The Depositional Habitat 2.1 The physical background to life in depositional habitats 2.2 The functional difference between mud and sand systems 3 The Functional Role of Biota 3.1 Patterns of life 3.2 Effects of sediment disturbance 3.3 Biodiversity impacts 3.4 Distribution in space and time 3.5 Trophic structure 3.6 New functional groups? 4 Future Shock: Climate Change and Ecosystem Function Acknowledgments References
12 Biogeochemical Dynamics of Coastal Tidal Flats
317 319 321 324 328 328 331 333 334 335 336 337 338 338
345
Samantha B. Joye, Dirk de Beer and Perran L.M. Cook 1 2 3
Introduction Transport Processes on Intertidal Flats Microbial Processes 3.1 Organic matter sources 4 Nitrogen Cycle 4.1 Nitrogen fixation 4.2 Nitrification and nitrate reduction 4.3 Exchange of dissolved nitrogen between the sediment and the water column 4.4 Benthic microalgal N assimilation 5 Phosphorus Cycle 6 Silicon Cycle 7 Concluding Remarks Acknowledgments References
345 347 350 351 352 352 354 357 358 359 362 364 365 365
x
Contents
Part IV Marshes and Seagrasses
375
13 Productivity and Biogeochemical Cycling in Seagrass Ecosystems
377
Marianne Holmer 1
Introduction 1.1 Primary productivity 1.2 Fate of primary productivity – export and burial 2 Sediment Biogeochemistry – Modified by Seagrasses 2.1 Microscale effects 2.2 Nutrient cycling – importance of root uptake 3 Human Pressures and Effects on Biogeochemistry 4 Future Perspectives and Conclusions Acknowledgment References
377 378 380 383 391 392 393 395 396 396
14 Tidal Salt Marshes: Geomorphology and Sedimentology
403
John R.L. Allen 1 2 3 4
Introduction Geographical Distribution Why Salt Marshes Exist? Geomorphology 4.1 Marsh evolution versus inheritance 4.2 Marsh edges and coastal change 4.3 Marsh terraces 4.4 Channels, creeks, and gullies 4.5 Creek networks 5 Morphodynamics 5.1 Tidal regime 5.2 Sediment sources and supply 5.3 Channelized flows 5.4 Platform flows 5.5 Accretion, compaction, and sea level change 6 Sedimentology 6.1 Grain size 6.2 Tidal bedding 6.3 Lithostratigraphic architecture 7 Concluding Discussion References
15 Ecosystem Structure of Tidal Saline Marshes
403 405 405 406 406 406 407 407 408 409 409 409 409 410 410 412 412 413 416 417 417
425
Jenneke M. Visser and Donald M. Baltz 1 2
Introduction Saline Marsh Communities 2.1 Emergent vegetation 2.2 Benthic algae
425 426 426 428
xi
Contents
2.3 Nekton 2.4 Reptiles 2.5 Birds 2.6 Mammals 3 Interaction among Communities 3.1 Effects of animals on emergent vegetation distribution 3.2 Emergent vegetation as animal habitat 3.3 Nursery function 3.4 Saline marsh food webs Acknowledgments References
16 Salt Marsh Biogeochemistry – An Overview
428 429 430 430 430 430 431 432 433 438 438
445
Craig Tobias and Scott C. Neubauer 1 2
17
Introduction Carbon 2.1 Exchanges 2.2 Internal cycling 2.3 Burial 3 Nitrogen 3.1 Exchanges 3.2 Internal cycling 3.3 Burial 4 Iron and Sulfur 4.1 Exchanges 4.2 Internal cycling 4.3 Burial 5 Phosphorus 5.1 Exchanges 5.2 Internal cycling 5.3 Burial 6 Marshes in Transition and Directions for Future Work Acknowledgments References
445 446 446 449 455 455 455 462 464 465 465 468 471 471 472 475 477 477 478 479
The Role of Freshwater Flows on Salt Marsh Growth and Development
493
Laurence A. Boorman 1 2
Introduction Freshwater Routes in Salt Marshes 2.1 Stream flow 2.2 Groundwater flow 2.3 Rainfall 2.4 Surface flow
493 494 495 497 499 500
xii
Contents
3
Associated Processes 3.1 Nutrient transport 3.2 Sediment transport 3.3 Organic matter transport 3.4 Pollutants 3.5 Salinity changes 4 Hydrological Impacts in Salt Marshes 5 Techniques for the Study of Marsh Hydrology 6 Implications of Freshwater Flows for Salt Marsh Management 7 Implications of Freshwater Flows for Salt Marsh Creation 8 The Ecohydrological Approach in Salt Marsh Studies 9 The Way Ahead – Problems and Challenges References
18 Tidal Freshwater Wetlands
500 500 502 504 504 505 507 508 509 509 511 512 512
515
Dennis F. Whigham, Andrew H. Baldwin and Aat Barendregt 1 2 3
Introduction Hydrogeomorphic Setting Biodiversity 3.1 Plants 3.2 Animals 4 Primary Production and Nutrient Cycling 5 Threats and Future Prospectus References
19 Biogeochemistry of Tidal Freshwater Wetlands
515 516 519 519 523 527 528 530
535
J. Patrick Megonigal and Scott C. Neubauer 1 2
Introduction Carbon Biogeochemistry 2.1 Carbon inputs 2.2 Carbon outputs 3 Processes Governing Organic Carbon Metabolism 3.1 Anaerobic respiration 3.2 Processes regulating methane production, oxidation, and emission 4 Nitrogen Biogeochemistry 4.1 Nitrogen exchanges 4.2 Nitrogen transformations 4.3 Nutrient regulation of plant production 5 Phosphorus Biogeochemistry 6 Silicon Biogeochemistry 7 Biogeochemical Effects of Sea-Level Rise 8 Concluding Comments Acknowledgments References
535 536 536 538 542 543 545 546 548 550 551 552 554 554 555 556 556
xiii
Contents
Part V Mangroves
563
20 Geomorphology and Sedimentology of Mangroves
565
Joanna C. Ellison 1 2
Introduction Mangrove Environmental Settings 2.1 Terrigenous settings 2.2 Islands 2.3 Inland mangroves 3 Tidal Range and Sea-Level Control 4 Sedimentation in Mangroves 4.1 Excessive sedimentation 5 Mangroves as Sea-Level Indicators 6 Storms/Tsunamis 7 Inundation Changes Affecting Mangroves 8 Conclusions Acknowledgment References
21 Geomorphology and Sedimentology of Mangroves and Salt Marshes: The Formation of Geobotanical Units
565 566 566 570 570 572 573 575 577 582 583 585 585 585
593
´ n J. Lara, Claudio F. Szlafsztein, Marcelo C.L. Cohen, Julian Oxmann, Rubee Bettina B. Schmitt and Pedro W.M. Souza Filho 1
Driving Forces Determining Main Morphology and Vegetation Types in the Coastal Zone 2 Depositional Environment and Substrate Formation for the Development of Mangroves and Salt Marshes 3 Influence of Sea Level and Climate Oscillations on Local Geobotanical Features 4 Major Factors Leading to the Development of Salt Marshes and Mangroves at the Amazon Coastal Region: An Integrated Analysis 5 Influence of Geomorphology and Inundation Regime of Geobotanical Units on their Sediment Biogeochemistry 5.1 Regularly inundated wetlands 5.2 Rarely inundated wetlands 5.3 Waterlogged wetlands 5.4 Salt marshes and mangroves: Nutrient sources or sinks? References
22 Paradigm Shifts in Mangrove Biology
593 595 597 599 600 601 604 606 607 608
615
Daniel M. Alongi 1 2
Introduction Shifts in Established Paradigms 2.1 Rates of mangrove net primary productivity rival those of other tropical forests
615 616 616
xiv
Contents
2.2
Mangrove forests appear to be architecturally simple, but factors regulating succession and zonation are complex 2.3 Mangrove tree growth is not constant but related to climate patterns 2.4 Tree diversity is low, but faunal and microbial diversity can be high 2.5 Arboreal communities are important in food webs, exhibiting predatory, symbiotic, and mutualistic relations 2.6 Plant–Microbe–Soil Relations are tightly linked and help conserve scarce nutrients 2.7 Crabs are keystone species influencing function and structure in many, but not all, mangrove forests 2.8 Algae, not just detritus, are a significant food resource 2.9 Mangroves are an important link to fisheries 2.10 Mangroves are chemically diverse and a good source of natural products 3 Conclusions References
23 Ecogeomorphic Models of Nutrient Biogeochemistry for Mangrove Wetlands
622 624 624 626 627 628 630 631 632 633 634
641
Robert R. Twilley and Victor H. Rivera-Monroy 1 2 3 4
Introduction Ecogeomorphology of Mangroves (Model 1) A Multigradient Model (Model 2) Geochemical Model (Model 3) 4.1 Redox zones in mangrove soils 4.2 Transition from reduced to oxidized zones 4.3 Hydroperiod effects on transition zones 4.4 Lower oxidation zone 4.5 Linkages in multigradient and geochemical models 5 Soil Biogeochemistry Model (NUMAN, Model 4) 6 Mass Balance Exchange (Model 5) 6.1 CO2 efflux from mangrove sediments and tidal waters 6.2 The balance of N fixation and denitrification 6.3 Tidal exchange 7 Contrasting Coastal Settings and Biogeochemical Models Acknowledgments References
641 648 649 653 653 654 655 656 657 659 662 662 663 671 672 675 675
Part VI Coastal Wetland Restoration and Management
685
24 Seagrass Restoration
687
Eric I. Paling, Mark Fonseca, Marieke M. van Katwijk and Mike van Keulen 1 2
Introduction Regional Activities
687 689
Contents
2.1 Europe 2.2 Australia 2.3 Oceania 2.4 Southeast Asia 2.5 China and Japan 2.6 New Zealand and the Pacific Islands 2.7 United States 3 Policy Issues Relevant to Mitigation 3.1 Costs of restoration 3.2 Valuation of ecosystem services 4 Conclusions Acknowledgments References
25 Tidal Marsh Creation
xv
689 693 695 696 697 698 699 701 702 703 704 704 705
715
Stephen W. Broome and Christopher B. Craft 1 2
Introduction Principles and Techniques of Tidal Marsh Creation 2.1 Site selection 2.2 Conceptual design 2.3 Hydrology 2.4 Soil 2.5 Establishing vegetation 3 Evaluating Functional Equivalence of Created Tidal Marshes 3.1 Biological productivity and food webs 3.2 Biogeochemical cycles 4 Summary References
26 Salt Marsh Restoration
715 716 717 717 718 719 720 723 723 730 733 733
737
Paul Adam 1 2 3 4 5
Introduction Setting Objectives Planning for the Future Addressing Causes and not Symptoms Managing Disturbance 5.1 Restoring hydrology 5.2 Managing weeds 5.3 Introduced fauna 5.4 Grazing 5.5 Pollution 6 Conflicting Priorities 7 Discussion 8 Conclusions References
737 738 740 744 745 745 746 749 750 751 754 755 756 756
xvi
Contents
27 Managed Realignment: Re-creating Intertidal Habitats on Formerly Reclaimed Land
763
Angus Garbutt and Laurence A. Boorman 1 2
Introduction Location, Drivers, and Constraints to Managed Realignment 3 Site Evolution 3.1 Sediments 3.2 Creeks 3.3 Soils 3.4 Nutrient fluxes 3.5 Vegetation 3.6 Fishes 3.7 Spiders 3.8 Benthic invertebrates 3.9 Birds 4 Challenges in Managed Realignment Research Acknowledgments References
28 Methods and Criteria for Successful Mangrove Forest Restoration
763 765 768 768 769 771 772 773 776 777 778 779 780 781 781
787
Roy R. Lewis III 1 Introduction 2 General Site Selection for Restoration 3 Specific Site Selection for Restoration 4 Establishing Success Criteria 5 Monitoring and Reporting Success 6 Functionality of Restored Mangrove Forests 7 Summary References
29 Evaluation of Restored Tidal Freshwater Wetlands
787 788 789 792 793 794 796 798
801
Andrew H. Baldwin, Richard S. Hammerschlag and Donald R. Cahoon 1 2 3 4 5
Introduction Characteristics of Restored TFW Success Evaluation of Restored Wetlands Ecosystem Attributes Measured at Restored TFW Criteria for Successful Restoration of TFW 5.1 Hydrologic criteria 5.2 Geomorphological criteria 5.3 Soil criteria 5.4 Salinity criteria 5.5 Vegetation criteria
801 802 805 806 808 813 813 814 815 815
xvii
Contents
5.6 Seed bank criteria 5.7 Benthic invertebrate criteria 5.8 Fish and wildlife criteria 6 Case Study: Evaluation of Restored TFW of the Anacostia River, Washington, DC, USA 6.1 Characteristics of restored and reference sites 6.2 Evaluation of success of restored TFW 7 Conclusions and Recommendations 7.1 Restoration of TFW in urban landscapes and selection of urban reference sites 7.2 Establishment of vegetation 7.3 Control of nonnative species 7.4 Implications for restoration of TFW Acknowledgements References
816 817 817 817 818 819 826 826 826 827 827 828 828
Part VII Coastal Wetland Sustainability and Landscape Dynamics
833
30 Surface Elevation Models
835
John M. Rybczyk and John C. Callaway 1 2 3
Introduction Measuring Processes that Affect Wetland Elevation Types of Models 3.1 Zero-dimensional mineral sediment models 3.2 Zero-dimensional organic sediment process models 3.3 Geomorphic models 4 Future Directions for Model Improvement 4.1 Data gaps 4.2 Integrating models 4.3 Improved linkage between sediment models and vegetation 4.4 Spatialization 5 Conclusions References
31 Salt Marsh–Mangrove Interactions in Australasia and the Americas
835 838 839 840 842 846 847 848 849 849 850 850 850
855
Neil Saintilan, Kerrylee Rogers and Karen McKee 1 2
3
Introduction Distribution/Geomorphic Settings – Where do Mangrove and Salt Marsh Coexist? 2.1 Mangrove distribution 2.2 Salt marsh distribution 2.3 Coexisting mangrove and salt marsh Long-Term Dynamics
855 856 856 858 859 862
xviii
Contents
3.1 Tropical northern Australia 3.2 Southeastern Australia 3.3 Western Atlantic–Caribbean Region 4 Recent Interactions 4.1 Air photographic evidence of mangrove–salt marsh dynamics in SE Australia 4.2 Saltwater intrusion in Northern Australia 4.3 Western Atlantic–Gulf of Mexico 5 Stressors Controlling Delimitation of Mangrove 5.1 Geomorphic and hydrological controls 5.2 Climatic controls 5.3 Physicochemical factors 5.4 Biotic interactions 6 Conclusions References
32 Wetland Landscape Spatial Models
862 863 865 866 866 867 867 869 870 873 874 874 875 876
885
Enrique Reyes 1 2 3 4
Introduction Physical Models Hydraulic Modeling Hydrodynamic Modeling 4.1 Finite difference solutions 4.2 Finite element solutions 4.3 Finite volume solutions 5 Ecological Models 6 Individual-Based Modeling 7 Eco-Geomorphological Modeling 8 Ecosystem-Level Modeling 9 Desktop Dynamic Modeling 10 Conclusions Acknowledgments References
885 887 887 889 892 893 894 895 895 896 897 900 900 902 902
Subject Index
909
Geographic Index
927
Taxonomic Index
933
PREFACE
Why coastal wetlands? What is so important about them that a whole book is required to try to review and explain their large variety of properties? Of all the coastal habitats, wetlands are the least depicted in the tourist brochures because they lack those paradisiacal long, white sandy beaches backed by palm trees or expensive resort hotels close to transparent blue waters. In fact, most coastal wetlands are quite muddy and are more likely to be inhabited by crabs and worms than by charismatic fish, birds and mammals. Hence, most inhabitants of our world either have never thought about coastal wetlands or may consider them a nuisance, not realizing that their seafood dinner likely had its origin as a detrital food web in a salt marsh or mangrove swamp. Bahı´ a Blanca (Argentina) inhabitants are a classical example: a city of over 300,000 people living within 10 km of a 2,300-km2 wetland, the largest of Argentina, but fewer than 40% have any idea that they are so close to the sea and a short distance of places that are globally unique (Perillo and Iribarne, 2003, in Chapter 6). Similarly, there are many other coastlines dominated by wetlands, yet they are only seen as areas to exploit in an unsustainable fashion. For example, mangroves have served local communities for generations in many Asian tropical countries for harvesting wood and fish in contrast to their wholesale replacement for rice cultivation and shrimp farming. Even though management guidelines have been available for decades, the negative consequences of uninformed exploitation have resulted in poor or even total lack of management criteria by most governments at all levels. Even local stakeholders fail to act in their own best interest without consideration of the ecosystem goods and services that the nearby wetlands provide. Coastal wetlands best develop along passive-margin coasts with low-gradient coastal plains and wide continental shelves. The combination of low hydraulic energy and gentle slope provides an ideal setting for the wetland development. Also passive margins are less prone to receive large episodic events like tsunamis. Tsunamis and storm surges, in particular, are major coast modifiers, but when they act on low coasts their effects are more far reaching than they are on higher relief coasts. For a wetland to form, there is a need for a particular geomorphological setting such as an embayment or estuary providing a relatively low-energy environment favoring sediment settling, deposition and preservation. However, that is only the beginning of a large and complex ‘‘life’’ where many geological (i.e., sediment supply, geological setting and isostasy), physical (i.e., oceanographic, atmospheric, fluvial, groundwater processes and sea level changes), chemical (i.e., nutrients, pollutants), biological (i.e., intervening flora and fauna), and anthropic factors play a wide spectra of roles. Coastal wetlands are areas that have combined physical sources and biological processes to develop structure that continues to take advantage of natural energy inputs. This book has been planned to address in an integrated way all these processes and their consequences on the characterization and evolution of coastal wetlands. It aims to provide an integrated perspective on coastal wetlands as ecosystems for the public, engineers, scientists and resources managers. It is only after acquiring xix
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Preface
this perspective that scientists can confidently propose ecohydrologic solutions for managing these environments in an ecologically sustainable way. This is but one small step toward encouraging humanity to look beyond purely technological, and often failed, solutions to complex environmental problems. This is done by focusing on the principal components considering the full range of environments from freshwater to subtidal and from polar to tropical systems. The book has been divided into seven parts starting from a synthesis chapter that integrates the whole book. Part I covers, in three chapters, the general description of the wetlands structured according to broad climatic regions and introduces the most important physical processes that are common to all coastal wetlands including some geomorphologic and modeling principles. Part II are specific to each particular type of wetland (tidal flats, marshes and seagrasses, and mangroves). Within each part (Parts III to V), there are chapters dealing with their particular geomorphology, sedimentology, biology and biogeochemistry. Finally Parts VI and VII provide insight into the restoration and management and sustainability and landscape dynamics. As editors, our work was greatly facilitated by the tremendous cooperation and enthusiasm from each of the authors to complete this process that began mid 2006. Each author, an authority in his or her specialty, was specifically invited to write a review chapter. Therefore, the challenges were much larger than in the case of typical contributed articles. But the reward, we think, is much more beneficial for the student, professor, or researcher employing this book for his or her particular interest. Readers will not only be able to find a specific topic but will find related information to complement and enhance the understanding of the topic. We are in debt to the more than 50 reviewers (most of them are not authors in the book) who have agreed and provided graciously and unselfishly their valuable time. Some took on the responsibility of two chapters, and their efforts are rewarded with improvements of each contribution. In many cases, reviewers gave us interesting ideas that helped in the general structure of the book. A list of the reviewers is provided. We also thank Elsevier Science and the various Publishing Editors who were in charge of our book along the period since we first proposed our idea to the final result that you are reading now. First of all to Kristien van Lunen who first believed that our proposal was realizable and then to the important contributions and patience of Jennifer Hele, and also to Pauline Riebeek, Linda Versteeg and lastly Sara Pratt. Stalin Viswanathan did an excellent job copyediting the whole book. Gerardo M. E. Perillo Eric Wolanski Donald R. Cahoon Mark M. Brinson July, 2008 This document is based on work partially supported by the U.S. National Science Foundation under Grants No. BSR-8702333-06, DEB-9211772, DEB-9411974, DEB-0080381 and DEB-0621014 and to SCOR under Grant No. OCE-0608600. Any opinions, findings, and conclusions or recommendations expressed in this material are those of the authors and do not necessarily reflect the views of the U.S. National Science Foundation (NSF).
LIST OF CONTRIBUTORS
Kenneth F. Abraham Wildlife Research and Development Section Ontario Ministry of Natural Resources Peterborough, Ontario, Canada Paul Adam School of Biological, Earth and Environmental Sciences University of New South Wales NSW 2052, Australia John R. L. Allen Department of Archaeology School of Human and Environmental Sciences University of Reading PO Box 227 Whiteknights, Reading RG6 6AB, UK Daniel M. Alongi Australian Institute of Marine Science PMB 3 Townsville M.C. Queensland 4810, Australia Rebecca J. Aspden Sediment Ecology Research Group Gatty Marine Laboratory St Andrews University Fife KY16 8LB, UK Andrew H. Baldwin Department of Environmental Science and Technology University of Maryland 1423 Animal Science Building College Park, MD 20742, USA Donald M. Baltz Coastal Fisheries Institute Louisiana State University Baton Rouge, LA 70803-7503, USA xxi
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List of Contributors
Aat Barendregt Department of Environmental Sciences Copernicus Institute for Sustainable Development and Innovation Utrecht University PO Box 80115 3508 TC Utrecht, the Netherlands Kevin S. Black Partrac Ltd. 141 St James Road Glasgow G4 0LT, UK Laurence A. Boorman L.A.B. Coastal Holywell, St. Ives Cambridgeshire PE27 4TQ, UK Mark M. Brinson Biology Department Howell Science Complex East Carolina University Greenville, NC 27858, USA Stephen W. Broome Department of Soil Science North Carolina State University Box 7619 Raleigh, NC 27695-7619, USA Donald R. Cahoon U.S. Geological Survey Patuxent Wildlife Research Center c/o BARC-East, Building 308 10300 Baltimore Avenue Beltsville, MD 20708, USA John C. Callaway Department of Environmental Science University of San Francisco 2130 Fulton St. San Francisco, CA 94117, USA Luca Carniello Dipartimento di Ingegneria Idraulica, Marittima Ambientale e Geotecnica and International Centre for Hydrology ‘‘Dino Tonini’’ Universita` di Padova via Loredan 20 I-35131 Padova, Italy
List of Contributors
Marcelo C. L. Cohen Laboratory of Coastal Dynamics and Centre for Geosciences Federal University of Para´ Avda Perimetral 2651 66077-530 Bele´m (Pa), Brazil Perran L. M. Cook Water Studies Centre Monash University Clayton 3800 Victoria, Australia Christopher B. Craft School of Public and Environmental Affairs Indiana University 1315 E. 10th Street Bloomington, IN 47405, USA Andrea D’Alpaos Dipartimento di Ingegneria Idraulica, Marittima Ambientale e Geotecnica and International Centre for Hydrology ‘‘Dino Tonini’’ Universita` di Padova via Loredan 20 I-35131 Padova, Italy Luigi D’Alpaos Dipartimento di Ingegneria Idraulica, Marittima Ambientale e Geotecnica and International Centre for Hydrology ‘‘Dino Tonini’’ Universita` di Padova via Loredan 20 I-35131 Padova, Italy Gareth Davies School of Earth and Environmental Sciences University of Wollongong Wollongong, NSW 2522, Australia Dirk de Beer Max Planck Institute for Marine Microbiology 28359 Bremen, Germany Andrea Defina Dipartimento di Ingegneria Idraulica, Marittima Ambientale e Geotecnica and International Centre for Hydrology ‘‘Dino Tonini’’ Universita` di Padova via Loredan 20 I-35131 Padova, Italy
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Joanna C. Ellison School of Geography and Environmental Studies University of Tasmania, Locked Bag 1376 Launceston, Tasmania 7250, Australia Mark Fonseca NOAA/NOS Center for Coastal Fisheries and Habitat Research 101 Pivers Island Road Beaufort, NC 28516-9722, USA Shu Gao School of Geographic and Oceanographic Sciences Nanjing University Nanjing 210093, China Angus Garbutt NERC Centre for Ecology and Hydrology Environment Centre Wales Deiniol Road Bangor, LL57 2UP Wales, UK Richard S. Hammerschlag U.S. Geological Survey USGS Patuxent Wildlife Research Center Laurel, MD 20708, USA Marianne Holmer Institute of Biology University of Southern Denmark Campusvej 55 5230 Odense M Denmark Robert L. Jefferies Department of Ecology and Evolutionary Biology University of Toronto Ontario, Canada Samantha B. Joye Department of Marine Sciences University of Georgia Room 220 Marine Sciences Bldg Athens, GA 30602-3636, USA
List of Contributors
List of Contributors
Jason R. Kirby School of Biological and Earth Science Liverpool John Moores University Byrom Street Liverpool L3 3AF, UK Stefano Lanzoni Dipartimento di Ingegneria Idraulica, Marittima Ambientale e Geotecnica and International Centre for Hydrology ‘‘Dino Tonini’’ Universita` di Padova via Loredan 20 I-35131 Padova, Italy Rube´n J. Lara Centre for Tropical Marine Ecology Fahrenheitstrasse 6 28359 Bremen, Germany Roy R. Lewis III Lewis Environmental Services, Inc. P.O. Box 5430 Salt Springs, FL 32134, USA Karen McKee National Wetlands Research Center U. S. Geological Survey Lafayette, LA 70808, USA Marco Marani Dipartimento di Ingegneria Idraulica, Marittima Ambientale e Geotecnica and International Centre for Hydrology ‘‘Dino Tonini’’ Universita` di Padova via Loredan 20 I-35131 Padova, Italy I. Peter Martini Department of Land Resource Science University of Guelph Ontario, Canada Yoshihiro Mazda Department of Marine Science School of Marine Science and Technology Tokai University 3-20-1 Orido Shimizu, Shizuoka 424-8610, Japan
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List of Contributors
J. Patrick Megonigal Smithsonian Environmental Research Center P.O. Box 28 647 Contees Wharf Road Edgewater, MD 21037, USA R. I. Guy Morrison Canadian Wildlife Service National Wildlife Research Centre Hull, Quebec, Canada Scott C. Neubauer Baruch Marine Field Laboratory University of South Carolina PO Box 1630 Georgetown, SC 29442, USA Julian Oxmann Centre for Tropical Marine Ecology Fahrenheitstrasse 6 28359 Bremen, Germany Eric I. Paling Marine and Freshwater Research Laboratory Environmental Science Murdoch University Murdoch 6150, Western Australia David M. Paterson Sediment Ecology Research Group Gatty Marine Laboratory St Andrews University Fife KY16 8LB, UK Gerardo M. E. Perillo CONICET – Instituto Argentino de Oceanografı´ a and Departamento de Geologı´ a Universidad Nacional del Sur CC 804 B8000FWB Bahı´ a Blanca, Argentina Marı´ a Cintia Piccolo CONICET – Instituto Argentino de Oceanografı´ a and Departamento de Geografı´ a Universidad Nacional del Sur CC 804 B8000FWB Bahı´ a Blanca, Argentina
List of Contributors
Andrew Plater Department of Geography University of Liverpool P.O. Box 147 Liverpool L69 7ZT, UK Paula D. Pratolongo CONICET – Instituto Argentino de Oceanografı´ a and Departamento de Biologı´ a, Bioquı´ mica y Farmacia Universidad Nacional del Sur CC 804 B8000FWB Bahı´ a Blanca, Argentina Enrique Reyes Department of Biology Howell Science Complex East Carolina University Greenville, NC 27858, USA Andrea Rinaldo Dipartimento di Ingegneria Idraulica, Marittima Ambientale e Geotecnica and International Centre for Hydrology ‘‘Dino Tonini’’ Universita` di Padova via Loredan 20 I-35131 Padova, Italy Victor H. Rivera-Monroy Wetland Biogeochemistry Institute Department of Oceanography and Coastal Science Louisiana State University Baton Rouge, LA 70803, USA Kerrylee Rogers Rivers and Wetlands Unit Department of Environment and Conservation, NSW PO Box A290 Sydney South NSW 1232, Australia John M. Rybczyk Department of Environmental Science Western Washington University 516 High St. Bellingham, WA 98225, USA
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Neil Saintilan Rivers and Wetlands Unit Department of Environment and Conservation, NSW PO Box A290 Sydney South NSW 1232, Australia Bettina B. Schmitt Centre for Tropical Marine Ecology, Fahrenheitstrasse 6 28359 Bremen, Germany Pedro W.M. Souza Filho Centre for Geosciences Federal University of Para´ Av. Augusto Correa 1 66077-110 Bele´m (Pa), Brazil Claudio F. Szlafsztein Laboratory of Coastal Dynamics and Centre for Geosciences Federal University of Para´ Avda Perimetral 2651 66077-530 Bele´m (Pa), Brazil Craig Tobias Department of Earth Sciences University of North Carolina at Wilmington 601 S. College Rd. Wilmington, NC 28403, USA Robert R. Twilley Wetland Biogeochemistry Institute Department of Oceanography and Coastal Science Louisiana State University Baton Rouge, LA 70803, USA Marieke M. van Katwijk Environmental Science Radboud University Nijmegen Postbus 9010 6500 GL Nijmegen, the Netherlands Mike van Keulen School of Biological Sciences and Biotechnology Murdoch University Murdoch 6150, Western Australia
List of Contributors
List of Contributors
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Jenneke M. Visser Department of Oceanography and Coastal Sciences School of the Coast and Environment Louisiana State University Baton Rouge, LA 70803, USA Present address: Institute for Coastal Ecology and Engineering University of Louisiana at Lafayette PO Box 44650 Lafayette, LA 70504-4650, USA Dennis F. Whigham Smithsonian Environmental Research Center P.O. Box 28 647 Contees Wharf Road Edgewater, MD 21037, USA Eric Wolanski, DSC (Hon. Causa), FTSE, FIE Aust Australian Centre for Tropical Freshwater Research and Department of Marine Biology and Aquaculture James Cook University and Australian Institute of Marine Science Townsville, Australia Colin D. Woodroffe School of Earth and Environmental Sciences University of Wollongong Wollongong, NSW 2522, Australia
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LIST OF REVIEWERS
Susan Adamowicz Carl L. Amos Shimon C. Anisfeld Edward Anthony Andrew H. Baldwin Thomas S. Bianchi Henry Bokuniewicz Henricus T.S. Boschker Mark M. Brinson David M. Burdick Jaye Cable Donald R. Cahoon John Calloway Marcelo Cohen Christopher Craft Steve Crooks Carolyn A. Currin Stephen E. Davis, III Donald D. DeAngelis Monique Delafontaine Norman C Duke Keith Dyer Susana Enrı´ quez Stuart Findlay Jon French Carl Friedrichs Keita Furukawa
Carl Hershner John Jeglum Samantha Joye Ruben Lara Mary A Leck Jorge Marcovecchio James Morris David Osgood Morten Pejrup Gerardo M. E. Perillo James Perry John Portnoy Denise J. Reed Peter Ridd Ken Ross Lawrence Rozas Neil Saintilan T.J. Smith Charles Tarnocai Stijn Temmerman Robert Twilley Reginald Uncles Steven Victor Melanie Vile Robert Whitlatch Rob Williams Eric Wolanski
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C H A P T E R
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C OASTAL W ETLANDS : A S YNTHESIS Eric Wolanski, Mark M. Brinson, Donald R. Cahoon and Gerardo M.E. Perillo
Contents 1. Introduction 2. A Synthesis of Coastal Wetlands Science 2.1. Geography 2.2. Geomorphology evolution under climate change 2.3. The influence of vegetation on the geomorphology evolution with climate change 2.4. The stabilizing role of vegetation 2.5. State change and coastal evolution 2.6. The role of physical disturbances 2.7. The role of herbivores 2.8. Observations across ecosystem types 2.9. The human impact 2.10. Modeling and predictions 2.11. Coastal wetland ecosystems as a component of estuaries 2.12. Coastal wetland socioeconomics 2.13. Coastal wetlands are essential for our quality of life 3. Lessons from the Chapters in This Book 3.1. Coastal wetlands as ecosystems 3.2. Physical processes 3.3. Tidal flats 3.4. Marshes and seagrasses 3.5. Mangroves 3.6. Coastal wetland restoration and management 3.7. Coastal wetland sustainability and landscape dynamics References
1 2 3 4 6 9 13 14 18 21 28 31 32 34 38 43 44 45 47 48 51 53 56 57
1. INTRODUCTION Our understanding of the functioning of coastal wetland ecosystems has grown rapidly over the past decade. We have gained insight into the roles of geomorphic processes, hydrologic dynamics, biotic feedbacks, and disturbance agents in creating and molding a variety of coastal wetland ecosystems across climatic gradients. The variety is expressed not so much in the more obvious differences in vegetation cover, Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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but rather how physical processes both facilitate and constrain a diversity of plant and animal communities. At one level, coastal wetlands are the product of tidal forces and freshwater inputs at the margin of continents. At another level, biotic factors exert feedback controls through biofilms, bioturbation of sediments, the buffeting of currents and waves, organic enrichment of sediments, and the closing of nutrient cycles. Few ecosystems provide us with such clear examples of feedback controls. Still, much remains to be learned and understood. What we do understand about the structure and functioning of coastal wetlands should provide the theoretical underpinnings for effective management in protecting them for their many contributions to ecosystem goods and services. And what we do not understand should compel us to focus our attention and energies toward seeking optimal solutions to some of the most perplexing and urgent problems facing natural resource management. What are coastal wetland ecosystems and what are their limits of distribution? There are several general definitions for wetlands, but the Ramsar definition is likely the most broadly encompassing (http://www.ramsar.org/), while others are more focused definitions tailored to country-specific protection and management policies (Mitsch and Gosselink, 2000). We offer a very general approach rather than a precise definition: coastal wetlands are ecosystems that are found within an elevation gradient that ranges between subtidal depths to which light penetrates to support photosynthesis of benthic plants and the landward edge where the sea passes its hydrologic influence to groundwater and atmospheric processes. At the seaward margin, biofilms, benthic algae, and seagrasses are representative biotic components. At the landward margin, vegetation boundaries range from those located on groundwater seeps or fens in humid climates to relatively barren salt flats in arid climates. This book focuses on commonly recognized ecosystems along this hydrologic gradient: seagrasses, tidal flats, tidal salt and freshwater marshes, and mangrove and tidal freshwater forests. Coral reefs are not covered at all because they are so physically and biologically distinct from the foregoing list, and in part because they have received research attention equivalent to the totality of all of the wetlands covered in this book (Duarte et al., 2008). Little direct reference is made in this book to lagoons that are intermittently connected to the sea; regardless, all of the ecosystem types (seagrass meadows, mudflats, marshes, and mangroves) can and do occur in these and other more specialized geomorphic settings. They would all comprise the array that we recognize as coastal wetlands. This book addresses the pressing need to quantify the ecological services provided by coastal wetlands as a tool to press for better management worldwide, because coastal wetlands are disappearing worldwide at an alarming rate; in some countries, the loss is 70–80% in the last 50 years (Frayer et al., 1983; Duarte, 2002; Hily et al., 2003; Bernier et al., 2006; Duke et al., 2007; Wolanski, 2007a).
2. A SYNTHESIS OF C OASTAL W ETLANDS SCIENCE In this section, we provide an overview of the structure and functioning of coastal wetlands with emphasis on key forces and processes that interact with their coastal geographic location. It is difficult to discuss these forces in isolation because
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climate change influences sea level, sea level forces changes in vegetation structure, disturbance and herbivory affect vegetation, and so forth. As such, comparison across major ecosystem types can provide insight into the differences and the relative importance of both extrinsic forces and intrinsic structure. We also discuss the role of modeling in elucidating the relative importance of key processes and in predicting the effects of alteration by humans. This last effect presents us with challenges of how to best protect and manage coastal wetlands for their attributes of life support, and importantly for the natural beauty and their contribution to cultural values. To this end, we outline key research needs that recognize the usefulness of working beyond “single factor cause and effect.”
2.1. Geography Coastal wetlands include salt marshes, mangroves, tidal flats, and seagrasses. They are found in all continents and at all latitudes. Cliffs and rocky shores are probably the only coasts with minimal wetlands. Mangroves cover about 230,000 km2 worldwide (Diop, 2003; Duke et al., 2007), and are restricted to a mangrove belt at low and mid-latitudes (Figure 1) as controlled by the combination of continental (no frosts or only very rarely, typically less than one frost every 10 years; Lugo and Patterson Zucca, 1977) and oceanic climates (warm waters; Duke et al., 2007). Salt marshes cover a larger area worldwide; in North America alone this area is 300,000 km2 mainly in Canada and Alaska (Mitsch and Gosselink, 2000), and possibly a similar large area of salt marshes exists in northern Russia. Salt marshes are also found scattered in the mangrove belt, usually in the upper intertidal areas landward of mangroves. Seagrasses cover at least 165,000 km2 worldwide, possibly up to 500,000 km2 (Green and Short, 2003). The worldwide area of tidal flats and freshwater coastal wetlands seems unknown though order-of-magnitude estimates suggest that they may reach 300,000 km2.
A
B Mangrove belt C
A
B
Figure 1 An approximate outline of the mangrove belt distribution around the world (- - - -), and the approximate distribution of the coasts that followed one of the three typical relative mean sea-level curve types A, B and C as shown in Figure 4. Redrawn from Ellison (2009).
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2.2. Geomorphology evolution under climate change Wetlands continuously evolved in time and space. The story of the present-day coastal wetlands starts 120,000 years ago, which is an interglacial period that lasted about 15,000 years and when the mean sea level was at a level comparable to the present one (Figure 1). Where natural conditions allowed, coastal wetlands probably flourished with a somewhat similar flora and fauna as today, though the macrofauna was very different. The human population was tiny at that time and had a negligible impact on the evolution of the coastal wetlands. Then came a 100,000-year-long ice age that buried much of the present temperate coasts and continental shelves under hundreds of meters, sometimes several kilometers, of ice. Water was taken away from the ocean and stored on the continents. As a result, the mean sea level decreased and was at its lowest, about 120 m below the present mean sea level, about 20,000 years ago (Figure 2). At that time, coastal wetlands would only have existed along those coasts that were not buried by permanent ice, mainly along the then “tropical” belt. Even there, the coastal wetlands could only have existed on the upper continental slope, which is where at that time the sea met the land (Figure 3). Because the inclination of the continental slope is much steeper than the continental shelf where the coastal wetlands are now located, the space for accommodating coastal wetlands would have been very limited. At that time, the area occupied by coastal wetlands would have been much smaller and in many cases restricted to estuaries that developed along the present-day submarine canyons (Perillo, 1995). Taking all of this into account, coastal wetlands would have occupied perhaps as little as 5% of the area covered just prior to human alterations. Then, 20,000 years ago came a period of high disturbance for coastal wetlands. The mean sea level rose rapidly, much faster than it fell in the previous 100,000
Figure 2 Time-series plot of the evolution of the absolute mean sea level during the last 140,000 years. Reproduced fromWolanski (2007a).
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River
High tide at present Low tide Δh 120 m
Δh
Present estuary and coastal wetland High tide 20,000 years BP
Continental shelf Continental shelf break
Shelf slope
Estuary and coastal wetland 20,000 years BP
Figure 3 Sketch of the continental shelf and continental slope, and the location of the estuary and the coastal wetlands at present and 20,000 years BP (years before present).
years (Figure 2). The mean sea level reached its present, world-averaged value about 6,000 years ago. At that time, the bulk of the continental ice outside of polar and near-polar regions had melted, which released a huge weight from those continental shelves that were burdened by ice. In those areas, the continental shelf rose by isostatic rebound of the land, with the largest rebound occurring where the ice burden was the greatest. In those areas (zone A in Figure 1), the land rose faster than the sea, and the relative mean sea level decreased over the last 10,000 years, the coast prograded, and new land emerged (Figure 4). In other areas,
Relative mean sea level (m)
80 A
40
C 0 B
–10,000
–8,000
–6,000 –4,000 Years BP
–2,000
0
Figure 4 The three dominant relative mean sea-level curves for the coasts in areas A, B, and C shown in Figure 1. In addition, there may be local tectonic motions such as those due to continental plate motions that can locally drown areas or uplift others (e.g., the southeast coast of Papua New Guinea). Redrawn from Ellison (2009).
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the relative sea level rose about 20 m during the last 10,000 years. In some areas (zone B in Figure 1), this increase was asymptotic, being largest 10,000 years ago and minimal at present (Figure 4); in such areas, the coast retreated from an advancing sea. In the other areas (zone C in Figure 1), the relative sea level reached a maximum about 5,000 years ago, typically 3 m higher and up to 7 m higher than at present, and it decreased smoothly to its present elevation until about 2,000 years ago. In those areas, the coast initially receded from an advancing sea and then prograded slightly during the last 5,000 years. Within each of these zones, there were local exceptions, especially in delta regions where sediment supply was sufficient to compensate for coastal retreat. Worldwide, the estuaries responded to these changing conditions of sea level, river and sediment discharge, changing sea ice, water currents, storms, and waves brought about by a new climate. The elevation of the mean sea level relative to land level determined at any time the location of the estuary. The net sediment budget, the balance between the sediment inflow (from the river and an eventual import from the sea) and the sediment outflow from the estuary to the sea, determined the evolution of the estuary. Where the relative sea fell rapidly (zone A), new land emerged constantly and the estuary migrated seaward, continually reinventing itself; the “old” estuary rose above the sea level and became part of the landscape; this evolution is still proceeding in zone A (Figure 5a). Where sea level rose, “old” estuaries were drowned and new estuaries formed landward where the land met the sea (Figure 5b). In zone A, no steady-state has yet been reached. The estuary is still steadily moving seaward as new land emerges. In zone B, a quasi-steady state is being reached only recently as the relative mean sea level stabilized only in the last 1,000 years or so. In zone C, the estuary is still evolving, albeit much slower than in the past because the changes in the relative mean sea level slowed down considerably during the last 6,000 years; estuaries are, however, still not at a steady state and coastal wetlands continue to migrate.
2.3. The influence of vegetation on the geomorphology evolution with climate change Estuaries trap some portion of the riverine sediment input (Perillo, 1995; Wolanski, 2007a). Estuaries in zone A are thus silting as soon as they form in a new position; small estuaries not flushed by the spring snow melt of large rivers are rapidly turning into mud flats that are stabilized by biofilms and they themselves can rapidly be colonized by salt marsh vegetation, which creates new high latitude salt marshes (Martini et al., 2009). The evolution of estuaries in zones B and C was controlled by the geomorphology when the rising seas flooded new land and by the tides, river runoff, and oceanographic conditions at the coast. Because these conditions varied from site to site around the world, a large number of estuarine evolution pathways occurred (Ellison, 2009; Woodroffe and Davies, 2009). At some locations (Figure 5c), a wide, shallow estuary formed; this enabled a vast coastal wetland to form that fringed the main channel as the estuary filled with sediment from both the river and the sea (through processes such as tidal pumping). Because of vegetation and adequate sediment supplies, this wetland type accelerated siltation and enabled the estuarine
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(a)
“Old” estuary is uplifted
Falling sea level “New” estuary is formed
(b) Rising sea level
“New” estuary is formed
“Old” estuary is drowned Recent inorganic riverine sediment cap
(c)
Riverine sediment
Present mean sea level Organic-rich wetland sediment Coastal sediment Holocene base
(d) Present mean sea level
Recent organic-rich wetland mud < 4,000 years Inorganic sediment Old organic-rich wetland mud > 6,000 years Holocene base
Figure 5 Sketch of the migration of an estuary with (a) falling and (b) rising sea level; (c) in macrotidal areas with large riverine sediment inflow, a vast coastal wetland often formed in the estuary that kept up with sea level rise until the last few thousand years when the wetland died as it was capped by inorganic riverine sediment; (d) in other open estuaries with a smaller riverine sediment inflow, rising seas drowned the wetlands, leaving behind organic-rich wetland mud. This mud was capped by inorganic sediment of oceanic and riverine origin, until, a few thousand years ago, it was close enough to the water surface that the coastal wetland could expand on it, creating a near-surface organic-rich mud layer; (e) in basins that were submerged by rising seas to depths too large to support a wetland, riverine mud accumulated in the basin until depths became small enough that a costal wetland was established and accumulated organic-rich mud; (f ) in lagoons with little riverine sediment inflow, a coastal wetland formed originally as the area was flooded by rising seas; the wetland could not keep up with rising seas; the wetland died; and was capped by calcareous sediment of oceanic origin that can support seagrass; (g) in other lagoons, a coastal wetland was formed as the area was flooded by rising seas; it died when it could not keep up with sea-level rise and it was capped by inorganic sediment of terrestrial and oceanic origin. When this sediment surface neared the water surface, a coastal wetland developed over this sediment, creating a near-surface layer of organic-rich mud. [Examples taken from Ellison (2009); Lara et al. (2009);Woodroffe and Davies (2009).]
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(e) Present mean sea level Organic-rich wetland mud Inorganic sediment wedge
Holocene base (f) Present sea level Seagrass Calcareous sediment Old wetland mud
Holocene base
(g) Present sea level New wetland mud Inorganic sediment Old wetland mud Riverine mud
Holocene base
Figure 5
(Continued)
system to keep up with sea level rise. Later on this estuarine, saline sediment was capped by freshwater sediment; the estuary had by then reached “old age” because it had used all the accommodating space on land and it can then only evolve further by moving offshore to form a delta; or it can occasionally be rejuvenated by opening a new channel or by isostatic or structural movements. In other cases (Figure 5d), the estuary moved upland with a rising sea level until about 6,000 years ago when the sea level stabilized; as the estuary migrated, so did the coastal wetlands, leaving behind their signature in the form of a well-preserved organic-rich mud layer. This mud was capped by inorganic sediment, which provided a substrate on which the wetland prograded, and in the process forming a near-surface, organic-rich mud layer. Still in other cases (Figure 5e), principally in microtidal areas and in areas where the geomorphology formed semienclosed coastal waters as the sea level rose, riverine sediment filled a lagoon-type environment, and this area was colonized by wetlands, leaving behind an organic-rich mud layer. The relative sea-level evolution over the last few thousand years has created in sediment-starved estuaries of zone C a varied succession of wetland habitats, including (Figure 5f) the die-off of mangroves and their capping by calcareous sediment later colonized by seagrass, and (Figure 5g) the die-off of mangroves, their capping by calcareous sediment with an even later capping by a new mangrove swamp with accreting, organic-rich mud.
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Thus, coastal wetlands have survived, adapted, and greatly expanded their distribution area during the last 10,000 years.
2.4. The stabilizing role of vegetation Several authors in this book (D’Alpaos et al., 2009; Ellison, 2009; Gao, 2009; Lara et al., 2009; Woodroffe and Davies, 2009) argue that these coastal wetlands are not just opportunistic colonizers of intertidal areas formed by the accumulation of mud in sheltered estuaries and coastal waters. As long as the mud flat elevation remains below the mean sea level, generally the mud flat remains fairly flat and uneventful and no tidal creeks may form (Figure 6a). Once the mud flat rises above mean sea level, the tidal hydrodynamics change profoundly. A large volume of water floods and drains the mud flat in a short time compared to the entire tidal period. This flow becomes funneled in a natural depression; this generates large currents, which in turn quickly erode the banks to form a tidal course (Figure 6b). The tidal course grows by the tidal asymmetry in it, with larger ebb than flood tidal currents, a (a)
(b)
(c)
(d)
Figure 6 Aerial photographs of (a) the 6 -km wide mud flat at the mouth of the Mary River, Australia’s Northern Territory, showing the river channel and no tidal creeks, and (b) a dendritical tidal creek in an unvegetated tidal flat (King Sound,Western Australia; mangrove trees are only present in the lower reaches of the creek where they form a one-tree wide vegetation strip). (c) The tip of a tidal creek in an unvegetated mud bank stops at the mangrove vegetation (Darwin Harbour, Australia). (d) Numerical prediction of the distribution of frictional stresses in a dendritical tidal creek in an unvegetated mud flat (D’Alpaos et al., 2009). (e) A dense mangrove root network locks the soil together to a depth of several meters and inhibits wave erosion of the bank. (f ^ h) High near-bottom vegetation density due to tree trunks, prop roots, buttresses, and pneumatophores in mangroves.
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(e)
(f)
(g)
(h)
Figure 6
(Continued)
situation characteristic of shallow water hydrodynamics (D’Alpaos et al., 2009; Mazda and Wolanski, 2009). In addition, intense erosion at the head of the tidal course is driven by small waterfalls that form there at ebb tide although other forms of erosion also help in the growth of the tidal creek over unvegetated mud flats (Perillo, 2009). At ebb tide, frictional forces retard the flow over the mud flat more than the ebb flow is retarded in the tidal channel (Figure 6c). A tidal creek forming in an unvegetated mud flat develops a dendritic pattern with several erosional “hot spots,” each one located at the tip of a creek (Figure 6d). Before a mud flat becomes vegetated, it can be stabilized by biofilms or destabilized by bioturbation. Once the mud flat is vegetated by pioneer species of mangroves or salt marsh vegetation, a quasi-steady state develops characterized by a sharp discontinuity separating the vegetated wetland from the unvegetated tidal channel (Figure 7; Perillo, 2009); the exception is Arctic coastal wetlands because scouring by ice reshapes the bathymetry yearly (Figure 7d). The patterns of erosion and siltation change profoundly once the vegetation is established for three main reasons. First, the roots of the trees or the grass lock the soil together and this slows or even stops bank erosion by tidal currents and waves to a depth of several meters in mangroves in well-aerated soils near the banks and several tens of centimeters deeper in the forest (Figure 6e), typically a few tens of centimeters in salt marshes, and 15 cm in seagrass (Holmer, 2009). This process is apparent in Figure 6c that
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(b) (a)
(c)
(d)
Figure 7 Tidal channels through coastal wetlands are generally unvegetated and can persist for long periods (years to hundreds of years) such as in these photographs at low tide of (a) a salt marsh channel (Mont Saint Michel, France; courtesy of E. Langlois-Saliou) and (b) a mangrove channel (Darwin, Australia). (c) Tidal channels start to form as soon as pioneer vegetation is established, even sparsely (Mont Saint Michel, France; courtesy of E. LangloisSaliou. (d) In the Arctic coastal wetlands, general mesotidal conditions and yearly scouring by ice tend to impede the formation of long-term tidal channels, especially on relatively flat coasts, though ice does lead to the formation of sculptured marshes (“jigsaw” marshes) ( James Bay, Canada; courtesy of I.P. Martini).
shows the tidal creek in the muddy boundary where it meets the vegetation. Second, the wave erosion of the vegetated substrate is decreased because hydraulic energy is dissipated by vegetation (Figure 6f–h for the case of mangroves; salt marsh vegetation is less tall, but in the lower 30 cm, the vegetation density can be higher than that of mangroves) and effectively absorbed by wave energy (Figure 8a–c for mangroves, seagrass, and salt marshes, respectively). Thirdly, the tidal currents are slowed down by the vegetation to values equal to 5% than in the tidal creek; also the currents around the vegetation generate eddies and stagnation zones where the suspended sediment brought in at rising tide will settle; the tidal currents are too small at falling tide to resuspend all that sediment; thus the vegetated tidal flat silts (D’Alpaos et al., 2009; Lara, 2009). From the wetland edge to the interior, waves are attenuated, thus reducing their capability to erode and transport sediments. To quantify the evolution of an estuary, models have been developed to compute the sediment dynamics over vegetated and unvegetated mud flats. The aim of these models is to simulate the evolving interaction between
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(a)
1.2
Substrate elevation (m)
Wave height (m)
No vegetation, high tide 0.8 Kandelia mangroves, high tide 0.4
Sonnertia mangroves, high tide
0 0 –0.2 –0.4 –0.6 –0.8 0
200
400 600 Distance (m)
800
1000
Figure 8 (a) Swell waves intruding in mangroves along the coast of the Gulf of Tonkin, Vietnam, decrease in amplitude with distance into the mangroves and this decrease is a function of the tidal height and the mangrove species [drawn from data in Mazda and Wolanski (2009)]. Hydrodynamics model predictions of the attenuation of swell waves (b) propagating over a seagrass meadow at low tide and (c) intruding in a salt marsh at high tide, together with the predicted wave height if there were no vegetation. In all cases (a ^ c), the vegetation significantly reduces the wave height.
currents and sediment transport, since this interaction determines the evolution of the bathymetry (D’Alpaos et al., 2009; Lara et al., 2009). Although most modelers rely on engineering formulae to calculate sediment transport from water currents and waves, in practice these formulae neglect the biology. Consequently, this approach is of little predictive value in vegetated coastal wetlands where the sediment dynamics are highly influenced by the inadequately quantified influences of biota and biotic processes. Even in the simplest case of unvegetated mud flats, a variety of biological mechanisms can result in either destabilizing the mud (e.g., by burrowing fauna) or in stabilizing the mud (e.g., by an algal mat on the substrate; Le Hir et al., 2007; Wolanski, 2007a). Further, there are no models to predict which flora and fauna will ultimately dominate (but see Reyes (2009) for “habitat switching” models). Thus, for vegetated wetlands (salt marshes and mangroves), the ability of models to predict the future evolution of the linked sedimentology–ecology remains an active and challenging area of research.
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(b)
1.2
No vegetation, low tide
Wave height (m)
1
0.8
0.6 Seagrass, low tide 0.4
Substrate elevation (m)
0.2 –1.8 –2 –2.2 –2.4 –2.6 0
Figure 8
100
200 Distance (m)
300
400
(Continued)
2.5. State change and coastal evolution A fluctuating sea level, and the geomorphologic evolution that it induces, generates changes of the ecological state of coastal wetlands. These changes modulate the transgression upslope from upland forest to subtidal ecosystems (Figure 9a) as a surface expression of the scenario shown in Figure 5c. During this coming century, the historical evolution of coastal wetlands described above may be accelerated, or reactivated for systems that have reached steady state at present, by the predicted increased rate of sea level rise due to global warming. Ecological succession of vegetation occurs within each of the states as a result of such disturbances as long as hydrology and sediments do not change substantially. For example, freshwater wetland forest landward of high marsh may die back after a fire, but unless salinity has increased, ecological succession will return the vegetation to forest. Alternatively, an increase in salinity will prevent forest re-establishment and subsequently change the state to emergent salt marsh. Often, the vegetation is out of phase with physical conditions so that disturbance such as fire or wrack deposition becomes the trigger to initiate change (Brinson et al., 1995).
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Eric Wolanski et al.
Wave height (m)
(c) 1.2
No vegetation, high tide
0.8
Saltmarsh, high tide 0.4
Substrate elevation (m)
0 3 2 1 0 0
Figure 8
200
400 Distance (m)
600
(Continued)
Over larger timescales and more divergent conditions, ecosystem types may change from one to another (Figure 9a). These coastal ecosystems have different flooding and aeration regimes that control plant dynamics (Figure 9b). Submerged aquatic vegetation (SAV, such as seagrass) has persistent flooding, mud/sand flats are intermittently exposed, and wetlands have intermittent aeration of the root zone. In contrast, uplands are not under the influence of sea level and rarely experience soil saturation. SAVs differ from flats by the presence of roots in the sediment and a requirement for transparency in the water column. State change among ecosystem types normally occurs over much longer timescales than ecological succession and may be better considered under the rubric of coastal evolution rather than disturbance.
2.6. The role of physical disturbances The role of physical disturbances in coastal wetland evolution is recognized but seldom quantified as yet. At high latitudes, salt marshes are yearly scoured by sea ice, creating a network of patches of rich vegetation, degraded vegetation, and unvegetated mud flats (Figure 7d; Martini et al., 2009). Further, ice erodes by floating imbedded sediments and transports them elsewhere for deposition. At temperate
15
Coastal Wetlands
(a)
Ag
ent
Salt water stress
s fo
rcin
gc
han
Wrack, root collapse
Dire “Sta
ctio
te”
n of
Sediment redistribution
High marsh
Forest
Low marsh
cha
nge
in e
cos
ge
yste
m
(b) Soil or sediment
Subtidal sediment
Roots
Water level fluctuation Water table Soil/sediment surface
SAV
Persistent
Mud/sand flat
Wetland
Flooding duration
Upland
None
Figure 9 (a) Predicted state change model in response to rising sea level that results in transgression upslope from upland forest to subtidal ecosystems. Modified from Brinson et al. (1995). (b) Differences in flooding duration among three coastal ecosystem types and uplands. SAV is an abbreviation for submerged aquatic vegetation, such as seagrass.
latitudes, crabs in salt marshes can excavate local depressions that can develop into tidal creeks or become unvegetated muddy patches (Perillo, 2009). In temperate and tropical climates, however, the most common natural disturbances are storms. In mangrove areas, hurricanes can destroy the vegetation facing the ocean by the force of the wind (Figure 10a). People living along the edge of the coastal wetlands are particularly at threat from storm surges (Figure 11) because storm surge flood waves occur at long timescales (hours to days) and are therefore little attenuated by vegetation. Tsunamis are rare but generate a major disturbance over much shorter periods of time. The presence of
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Eric Wolanski et al.
(a)
(b)
Figure 10 (a) Mangroves defoliated by a hurricane in Florida (courtesy of USGS). (b) A 5- to 10 -m wide strip of mangroves along the estuary banks was destroyed by the 2004 Indian Ocean tsunami, and the remaining mangrove forest remained intact (Khao Lak, Thailand).
extensive areas of shallow waters, such as intertidal flats, is important because it presents a wide area for dissipation of energy before the wave hits the vegetation. In mangroves, for small (<6 m) tsunamis waves the damaged area may be limited (Figure 10b) to a 5- to 10-m wide strip along the tidal creek and another, much wider, strip along the coast (Mazda and Wolanski, 2009). This occurs because the (a)
(b)
(c)
Figure 11 People living along the edges of coastal wetlands such as in (a) the Fly River estuary, Papua New Guinea, (b) Chumphon mangroves, Thailand, and (c) the mangrove inlets at Cedar Key, Florida, USA, are particularly at threat from natural hazards including tsunamis and storm surges.
17
Coastal Wetlands
vegetation is an obstacle to swift flows and thus steers the tsunami wave along the tidal creek. Thus, people (fishermen) living along the tidal creek – with no vegetation between them and the creek – are not protected against a small tsunami; however, people living behind the mangrove belt receive some protection (Wolanski, 2007b; Alongi, 2008). Provided that the tsunami wave is less than 6 m and the trees are fully developed, the vegetation can survive and the wave at 500 m inland is transformed from rising in 10 seconds to rising in 3 minutes; this gives a chance for people to take shelter or flee the area, and the attenuation considerably reduces damage to property and infrastructure (Wolanski, 2007b). A large tsunami (wave >6 m) flattens and uproots the vegetation. In some cases, the energy extracted from the wave and used for this destruction reduces the wave energy sufficiently to provide some protection to people living in the coastal zone inland from the mangroves (Figure 12). However, for a large tsunami wave, mangroves and coastal forests are known to significantly contribute to human mortality from the physical damage of water-borne debris (Cochard et al., 2008). In climates with distinct wet and dry seasons, the evolution of the bathymetry of a tidal creek may be determined by seasonal bank slumping principally at the end of the wet season, which is a large-scale process that vegetation cannot stop (Figure 13a) and one which all estuarine geomorphology evolution models ignore so far. In other cases, (a)
(b)
Figure 12 Satellite images of Kitchall Island (Nicobars, Indian Ocean) (a) before and (b) after the 26 December 2004 tsunami. The likely tsunami wave direction is shown by arrows. The white dotted line shows the bay. Before the tsunami, the bay was covered by mangroves. After the tsunami, no mangroves survived as the tsunami uprooted or snapped the trees. The vegetation and the villages in the area marked “B” behind the mangroves were not badly impacted by the tsunami disaster because their high elevation together with the tsunami wave attenuation by the vegetation prevented flooding; villages and the agricultural areas marked “A”were not in the lee of mangroves and were destroyed or severely damaged by the tsunami. Thus, the land behind the mangroves was partially protected by the sacrificial belt of mangroves. At longer timescales, the surviving mangroves died because tectonic movements changed the substrate elevation and the tidal hydrology (courtesy of Y. Mazda).
18
(a)
Eric Wolanski et al.
(b)
Figure 13 (a) Seasonal estuarine bank slumping at the end of the wet season (Daly Estuary, NT, Australia). (b) The growth of a tidal creek in freshwater flood plains is facilitated by the destruction of the protective vegetation by salt water intruding in the dry season and creating bare mud flats that may be colonized by mangroves years later (Mary Estuary, NT, Australia).
the tidal creek develops from headward erosion at the tip of the creek during floods; during the subsequent dry season, saline water floods the eroded area, kills the freshwater vegetation, and creates bare soils that are unprotected by vegetation and thus readily erode; this allows the creek to grow in the tidal freshwater flood plains (Figure 13b). This process is also still largely ignored by estuarine geomorphology evolution models. A rapid sediment deposit of more than 5 cm (during a storm) can kill mangroves (Ellison, 1998); by contrast, salt marsh plants seem able to cope well and plant growth is even accelerated following such events (Boorman, 2009; Garbutt and Boorman, 2009). Both salt marshes and mangroves suffer mortality through hydrologic regime change, particularly from sustained flooding or loss of structural support through sediment erosion. The below-ground biomass of mangroves, seagrass meadows, and salt marshes offers enormous protection against erosion, thus providing a more stable medium for plant growth. The soft sediments of unvegetated intertidal flats are highly mobile and generally inhospitable to emergent plant growth. The establishment of emergent vegetation is rapid once the right hydraulic conditions occur, and from that time onward, the vegetation modifies the water and sediment circulation and promotes further biomass accumulation and the creation of a new ecosystem. What parameters control the threshold conditions for the establishment of pioneer vegetation over a soft sediment system still remain unclear. Once initial colonization has been achieved, vegetation may develop rapidly, being possibly a seagrass meadow, a mangrove swamp, or a salt marsh. This is commonly achieved by the arrival of species of fauna and flora with strong dispersal characteristics, followed by corresponding changes in food web dynamics, sedimentology, and hydrology. Humanity has accelerated this process by inadvertently or purposefully encouraging the spread of exotic invasive species and by the nutritive effects of eutrophication.
2.7. The role of herbivores In mangroves, crabs dig burrows often at high densities (Figure 14a). These burrows circulate tidal waters and permit salt to be flushed from sediments, thus
19
Coastal Wetlands
(a)
(b)
(c)
(d)
(e)
(f)
Figure 14 (a) Crab burrows are numerous in healthy mangrove soils (Congo River Delta). (b) Mangroves permanently destroyed by groundwater hypersalinity and acid sulfate leaching as a result of poor land use practices in reclaimed mangrove soils (Konkoure River delta, Guinea). (c) Nearly all mangrove seedlings planted in this abandoned shrimp pond died due to drowning as the tidal hydrology was not properly restored (Surat Thani, Thailand). (d) This mangrove seedling was destroyed by being physically brought down by the weight of algae due to local eutrophication from cattle dung (Iriomote Island, Japan). A bamboo curtain (e) and a seawall (f ) were constructed as a wave barrier to protect mangrove seedlings in a mangrove restoration site in the upper Gulf of Thailand.
preventing hypersalinity (Alongi, 2009; Mazda and Wolanski, 2009) that would otherwise occur from mangroves that selectively avoid salt uptake and increase pore water salinity of the sediments. When crabs are destroyed, the groundwater becomes hypersaline and the vegetation is either destroyed or it becomes stunted. Similarly, farming on sulfur-rich soils acidifies the soil through the oxidation of
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Eric Wolanski et al.
sulfides; the vegetation is destroyed and cannot be recovered (Figure 14b) without massive additions of lime. Coastal wetlands with an aerial canopy of vegetation generate a microclimate at the sediment surface, which in turn affects the fauna and flora. For instance at mid-latitudes, the daily temperature variation in stagnant pools of water in wetlands exceeds 5C if the pools are shaded by mangroves and 15C if the pools are not shaded; this in turn affects the population dynamics of fish and mosquito in these pools (Knight, 2008). The fate of dead organic matter is important in determining the evolution of a coastal wetland. There are large differences between the fate of above-ground plant biomass and the below-ground plant biomass. Live above-ground biomass can be consumed by crabs, snails, shrimps, and cone shells (Figure 15a–b); for instance, shrimps can remove up to 50% of seagrass leaf production (Vonk et al., 2008) and crabs and snails climb trees to harvest mangrove leaves (Vannini and Ruwa, 1994; Vannini et al., 2008). Another fraction can be exported as particulate organic matter outwelling that in turn enhances detrital food webs in the adjoining estuarine and coastal waters; the remaining fraction may be accumulated on the substrate for eventual burial by sediment deposition. Below-ground organic biomass is generally not exported in particulate form but rather is lost through decomposition by microbes as carbon dioxide and methane, and dissolved nutrients. Decomposition takes several months to years for labile organic matter, while refractory organic matter may take hundreds of years to decompose or become permanently buried through peat accumulation. The fate of plant litter in coastal wetlands is largely controlled by the activities of benthic fauna and by the circulation of water. Crabs, bivalves, and snails can remove and process up to 80% of mangrove leaf litter (Alongi, 2009; Lara et al, 2009; Smith et al., 1991). This enables the nutrients to be recycled and less likely to be lost by export with the next outgoing tide. In turn, the distribution of these animals is linked to the sedimentology, topography, and hydrology. The fate of dead, above-ground organic matter differs widely between the types of coastal wetlands. Much of the (a)
(b)
Figure 15 (a) Cone shells eating leaves in mangroves (Iriomote Island, Japan; courtesy of K. Furukawa). (b) Crabs consume mangrove litter that falls on the ground; some crabs also climb on trees, principally at night, and eat the young leaves that have little tannin in them (Australia mangroves; courtesy of N. Duke).
Coastal Wetlands
21
Figure 16 Awrack of dead salt marsh plants and some plastic litter after a storm (Bah´ıa Blanca Estuary, Argentina). The plant litter is composed of dried stems of Spartina alterniflora. Sarcocornia perennis plants die when covered by the dead plants.
litterfall is consumed in mangroves. In some salt marshes, a large portion of live plant biomass may be consumed and subsequently exported as birds migrate away from sites of grazing (Canada geese in North America). In other salt marshes, plant litter is not consumed by herbivores or exported to the estuary; instead large racks of dead grass form in winter as the above-ground component dies (Figure 16). The amount of that organic matter is enormous adding up to tons for a wetland area of a few square kilometers and export from seagrass meadows during storms can result in massive accumulations on beaches (Holmer, 2009). Some of that material remains trapped in the wetland, redistributed during extreme tides, or exported as an outwelling event during a single storm. The fate of this material once it arrives in the estuary and coastal water remains largely unknown. Fish and crabs can also consume biomass on wetland surfaces at high tide and then facilitate export as they move toward the estuary with ebbing tides. Thus, mangroves, salt marshes, and seagrass meadows can be net exporters of organic matter depending on local physical conditions. This link between flora and fauna in the energy flows and nutrient fluxes between ecosystem types is a rich, largely unexplored, field of research. Coastal wetlands are constantly evolving by feedback mechanisms, especially with respect to rising sea level and to disturbances. The state of a particular wetland at any one time affects its future and is determined by the accumulated history of previous states. What we see in a wetland is a snapshot in time.
2.8. Observations across ecosystem types Surface and underground water circulation are vital to the sedimentation processes, chemistry, and biology of coastal wetlands. There is a strong feedback between water circulation and sediment dynamics and the constantly changing geomorphology of coastal wetlands. The circulation of surface and sediment porewaters in
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coastal wetlands also plays a key role in the adjoining estuarine and coastal ecosystems and their biology and biodiversity. The processes vary spatially from small to large scales, and temporally at timescales from that of individual mixing events to tidal to geomorphological scales. This results in a complex interplay between the hydrodynamics, sedimentology, and biology of the ecosystem. Almost every day of field work in coastal wetlands reveals new aspects of the inter-relationships among water flow, sediments, flora, and fauna. To cite a few, we compare in Table 1 the attributes of the five major ecosystem types in terms of their geographic distribution, physical limits, food web dynamics, and responses to human activities. In making such comparisons, there is a tendency to overgeneralize and to gloss over the diversity of patterns found within each of the types. We make these comparisons in the spirit that they may lead to insight not obvious by working within or examining a single ecosystem type (the remainder of this section refers to the contents of Table 1). Some patterns are obvious, such as climatic differences between the latitudinal ranges of salt marshes and mangroves. While the geographic distribution is reasonably well established, inventories of the area covered by each ecosystem type are less reliable. For example, tidal freshwater wetlands (TFWs) have not been much studied outside of the North American and European continents, so their distribution has not been mapped to our knowledge (Whigham et al., 2009; Conner et al., 2007). In fact, most of these ecosystems are associated with the large deltas of the world where dominance by freshwater discharge results in salinities lower than is often associated with the term “coastal.” At a particular site, physical limits and factors (water depth, wave energy, salinity range, rate of rising sea level, etc.) determine the type of coastal wetland. Because of the high diversity in life forms across ecosystem types, ranging from diatoms to trees for just the primary producers, there is great variability in the capacity of biotic structure to influence physical processes. For tidal flats and seagrasses, both of which are dominated by obligate aquatic taxa, maximum depth distribution (or lowest elevation) is limited by light availability when flooded. The upper elevation of disturbance is limited by desiccation when the ecosystem is exposed. For emergent life forms (marsh grasses, shrubs, and trees), the vertical range is highly influenced by tidal amplitude, that is, greater tidal amplitudes allow a greater elevational range of distribution (McKee and Patrick, 1988). Variations in intertidal soil salinity are largely a consequence of climate except in deltas where freshwater discharge dominates hydrology. In humid climates, groundwater discharge establishes the upper boundary of mangroves and marshes (Gardner et al., 2002; Plater and Kirby, 2006), while arid climates and associated high soil salinities restrict the landward extent of coastal wetlands (Pratolongo et al., 2009). With changes in the relative position of sea level, the boundary at which coastal effects are no longer apparent changes over time as wetlands migrate landward in response to rising sea level (most regions; see Figure 1) and regress as sea level drops (mainly at high latitudes) (Figures 4 and 5.). Virtually all coastal wetlands respond to and rely on sediment sources and exchanges. The effects of suspended sediment on light transmission in the water column are critical for survival of seagrasses and likely have short-term effects on
Characteristics
Intertidal flats
Seagrass meadows
Saline marshes
Freshwater marshes
Mangrove forests
Same as saline marshes, forested wetlands occur also in midlatitudes Grossly underestimated
Mean monthly temps >20C
Geographic distribution Geographic rangea
Global abundanceb
Unlimited
Absent in polar regions
Mid to high latitudes, replaced by mangroves in subtropics
Widespread, undocumented
18 106 ha
140 106 ha
Coastal Wetlands
Table 1 Comparison of coastal wetland ecosystem characteristics for major types
15 106 ha
Physical limits Maximum water depth limitation/minimum elevationc
Approx 20% incident light
Varies with amplitude and temperature
Unresolved
Unresolved
Maximum elevation/ minimum hydroperiodd
Light limitation of net primary production where P/R <1 Supratidal during storms
Intertidal conditions (exposure to some drying)
No information
Salinity rangee
Variable
Source of material for vertical accretionf
Allochthonous inorganic only (biofilm stabilizing effect) Not applicable
Euryhaline to polyhaline Also organic accretion
Groundwater discharge zone, salt accumulation in arid climates Hyperhaline to oligohaline Also organic accretion
Also organic accretion
Groundwater discharge zone, salt accumulation in arid climates Hyperhaline to oligohaline Also organic accretion
Not known
Fish nursery variation
Surge may affect soil salinity
Decades-long effects after blowdown of trees
Interannual variation in sea level; ENSO cyclesg Hurricanes and cyclonesh
Likely short-term effects
Not applicable Temporary unless burial by sediments or severe erosion
Demonstrated effect on primary production Local shore erosion and sediment redeposition
Fresh to oligohaline
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Table 1
(Continued )
Characteristics
Intertidal flats
Seagrass meadows
Saline marshes
Freshwater marshes
Mangrove forests
Food web dynamics Typical dominant primary producersi Grazing food websj Detrital food websk Openness of organic matter fluxesl
Epipelagic microalgae (esp. diatoms) and macroalgae Infauna, epifauna Indistinguishable from grazing No information; presumed open
Submersed vascular plants
Emergent grasses and forbs
Ranges from herbaceous to forest dominance
Trees and shrubs
Large mammals and turtles Invertebrate dominated Exports and imports well documented
Snails and insects
Rodents
Tree crabs
Invertebrate dominated
Invertebrate dominated Few studies
Burrowing crabs
Evidence of net exports
Evidence of net exports
Response to human activities – global change No information
Thresholds not established
Possible distribution to higher latitudes
No information
No information
Thresholds not established
Strong sediment sources needed for survival (little information)
Altered salinities in response to climate drying or wettingo
No information
Hypersalinity in seasonally isolated lagoons; freshening in others
Strong sediment sources needed for survival (glacial rebound areas excluded) Expansion or reduction of salt flats
Transformation to greater or lesser salinity tolerant species
Possible distribution to higher latitudes lacking frost Strong sediment sources needed for survival Expansion or contraction of salt flats
Eric Wolanski et al.
Response to climate warming: ambient temperaturesm Response to climate warming: acceleration of rising sea leveln
Increased salinity from reduced freshwater flows (upstream withdrawals, irrigation, etc.)p Reduced sediment supply due to reduced freshwater flows (dams, etc.)q
Increasingly stressful
Hyper salinity in seasonally isolated lagoons
Expansion of salt flats
Transformation to salinity tolerant species
Expansion of salt flats
Reduced development of tidal flats
Erosion
Erosion
Erosion, hypersaline conditions
Enhanced or excessive sediment supplyr
Accretion with shift to marsh and mangrove colonization
No information (reduced suspended sediments beneficial to water clarity) Reduced water clarity stressful to seagrasses
Accretion self-limiting
Sensitive to burial of pneumatophores
Tidal barriers – dyking and bulkheadss
Not applicable
Not applicable
Eliminates physical and biotic exchanges
Eutrophicationt
Proliferation of macroalgae
Increased herbivory
Bottom disturbance: trawlingu Bottom disturbance: channel dredging, fish/shrimp ponds, etc.v Fragmentation within habitatw Loss of connectivity with other habitatsx
No information
High sensitivity to nutrients; Macroalgal smothering Extremely destructive
Massive development of marshes historically (Chesapeake Bay, USA) Eliminates physical and biotic exchanges No information
Not applicable
Not applicable
Deepening beyond euphotic zone No information
Extremely destructive
Extremely destructive
Extremely destructive
Extremely destructive
No information
No information
No information
No information
No information
No information
No information
No information
No information
Coastal Wetlands
Response to human activities – local and upstream
Eliminates physical and biotic exchanges Altered root growth and surface elevation trajectories Not applicable
25
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Table 1
(Continued )
Characteristics
Intertidal flats
Seagrass meadows
Saline marshes
Freshwater marshes
Mangrove forests
Phragmites throughout No information
No information
Response to human activities – biotic y
Invasive species
No information
Diseasesz
No information
Caulerpa taxifolia aggressive clones Wasting disease historically
Phragmites in spring tidal and fresh zones No information
No information
a
Martini et al. (2009); Pratolongo et al. (2009); Woodroffe and Davies (2009); Conner et al. (2007). Duarte et al. (2008) for all but intertidal flats and freshwater tidal wetlands, the latter grossly underestimated (Whigham et al., 2009). Holmer (2009); McKee and Patrick (1988) for saline marshes, Gallegos and Kenworthy (1996) for seagrass. d Martini et al. (2009); Pratolongo et al. (2009); Woodroffe and Davies (2009). e Holmer (2009); Martini et al. (2009); Pratolongo et al. (2009); Visser and Baltz (2009); Whigham et al. (2009); Woodroffe and Davies (2009). f Allen (2009); Megonigal and Neubauer (2009); Lara et al. (2009); Rybczyk and Callaway (2009); Tobias and Neubauer (2009). g For saline marshes, Morris et al. (1990); for mangroves, Rehage and Loftus (2007). h For mangroves, Smith et al. (1994); for marshes and mangroves, Cahoon (2006); for marshes van de Plassche et al. (2004). i Holmer (2009); Martini et al. (2009); Pratolongo et al. (2009); Saintilan et al. (2009); Visser and Baltz (2009); Whigham et al. (2009); Woodroffe and Davies (2009). j Martini et al. (2009); Paterson et al. (2009); Pratolongo et al. (2009); Woodroffe and Davies (2009) for tidal flats; Visser and Baltz (2009) for saline marshes; and Whigham et al. (2009) for tidal freshwater marshes. k Martini et al. (2009); Paterson et al. (2009); Pratolongo et al. (2009); Woodroffe and Davies (2009) for tidal flats; Visser and Baltz (2009) for saline marshes; and Whigham et al. (2009) for tidal freshwater marshes. l Alongi (2009); Holmer (2009); Megonigal and Neubauer (2009); Paterson et al. (2009); Visser and Baltz (2009). m Alongi (2009); Ellison (2009); Woodroffe and Davies (2009). n Ellison (2009); Martini et al. (2009); Pratolongo et al. (2009); Woodroffe and Davies (2009). o Predictions depend upon local effects of precipitation/evapotranspiration change as they may affect continental sediment supplies. p No documented examples in chapters. q Pasternack et al. (2001). r Pasternack et al. (2001); Ellison, (2009). s Garbutt and Boorman (2009). t Seagrasses particularly vulnerable; Holmer (2009), mangroves – McKee et al. (2007). u Paling et al. (2009). v Paling et al. (2009). w No information. x No information. y Baldwin et al. (2009); Holmer (2009); Whigham et al. (2009). z Holmer (2009). b c
Eric Wolanski et al.
Coastal Wetlands
27
the primary productivity of tidal flats. For marsh and forest ecosystems, sediment accumulation through deposition is a critical process because it maintains the relationship between a wetland surface and sea-level change. This dynamic interaction is to some degree self-maintaining because too much accretion places the sediment surface too high for flooding and sediment deposition, while too little accretion has the opposite effect (Rybczyk and Callaway, 2009). In quiescent environments where sediments supplies are low or negligible, organic matter accretion responds to similar feedbacks in all coastal ecosystem types but tidal flats. Interannual variation in sea-level stand adds another dimension of complexity, but it has been demonstrated only for primary production rates in salt marshes (Morris et al., 1990). Of course, hurricanes, cyclones, and tsunamis, where they occur, can produce short-term disruptions of tidal flats, but highly variable effects locally for other types (Cahoon, 2006). For salt marshes, sediment erosion is a shore phenomenon that in extreme cases can remove large areas in a single storm. In an example studied by van de Plassche et al. (2004) in Connecticut (USA), massive removal of salt marsh sediments could be returned to intertidal status only with substantial infilling and regrowth of vegetation. In contrast, mangroves can experience massive blowdown with return to full growth forests only on decadal timescales (Cahoon et al., 2003). The effects of soil salinity on tidal freshwater swamps in temperate zones have been shown to cause tree mortality and replacement with marsh vegetation (Conner et al., 1997). Fire is likely a disturbance only in the upper portions of tidal wetlands where the mortality of trees can accelerate the landward movement of marsh vegetation (Poulter, 2005). Food web dynamics vary greatly among ecosystem types. Tidal flats represent the largest departure from other ecosystem types in that the grazing food web is dominated by deposit feeders rather than the more typical herbivory of higher plants. This is somewhat deceiving because marshes and swamps also have epiphytic communities that support substantial food webs. Algal production in both seagrasses and salt marshes has been demonstrated repeatedly to support substantial flows of energy. This is not to diminish the role of living plant tissue in supporting grazing food webs because there are examples of substantial herbivory, ranging from sea turtles and dugongs for seagrasses and invertebrates for mangroves and salt marshes. Regardless, plant tissue is often unpalatable to many potential consumers resulting in the majority of primary production being entrained in the detrital food webs (Tobias and Neubauer, 2009; Visser and Baltz, 2009). In spite of the openness of most tidal ecosystems to organic matter exchanges, both particulate and dissolved, there is general agreement that greater tidal amplitudes not only facilitate imports and exports but that associated currents serve to amplify nutrient cycling and associated primary productivity. Even seagrasses, which are completely open to exchange, have the capacity through their baffling effect to trap much more particulate organic matter than bare sediments (Holmer, 2009). Nevertheless, it should be pointed out that shallow bare sediments used for comparison are similar to the intertidal flats that fully contribute to the habitat complexity of coastal wetlands. Migrating birds also provide a connection, through biomass transfer, between coastal wetlands that can be adjoining or widely separated, even in different hemispheres.
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2.9. The human impact Starting about 7,760 years ago in China (Zong et al., 2007) and typically a few hundreds years ago in most other coastal areas, humanity has profoundly impacted, degraded, or destroyed many coastal wetlands worldwide by direct physical degradation and pollution. Ironically, reduced coastal wetland area increases the threat to human safety at the same time that shoreline development exposes populations to coastal hazards such as tsunamis, erosion, flooding, storm waves, and surges. As a result, nearly all salt marshes have been destroyed by land reclamation in several highly populated temperate countries, including Japan, China, and the Netherlands (Wolanski, 2007a). Recent wetland destruction has been phenomenal in the United States where over 9 million hectares of wetlands were lost between 1950 and 1970, the greatest loss occurring in coastal wetlands of the Gulf of Mexico (Frayer et al., 1983). Despite the “swampbuster” 1985 Food Securities Act and the “no net loss” of wetlands policy that emerged in 1989, wetland loss along that coast continues, though at a slower rate (Streever, 2001). This loss in turn is economically significant because coastal wetlands are critical habitats for many fishery species in the Gulf of Mexico (Heck et al., 2003; O’Connor and Matlock, 2005). The remaining wetlands are sinking and shrinking as they are not replenished by sediment because rivers are diverted elsewhere and this exacerbates flooding by river floods and storm surges (Streever, 2001). This march of humanity now threatens also tropical coastal wetlands; indeed we face the prospect of a world without mangroves this century (Diop, 2003; Duke et al., 2007). The main threats are urbanization, clear cutting mangroves for charcoal, and conversion of the land for urbanization, salt ponds, rice farms, and shrimp ponds (Figure 17). Arctic coastal wetlands remain the least impacted, simply because there are fewer people living in these areas. Geomorphology teaches us that all coastal wetlands suffer mortality if mineral and organic sediment accumulation is unable to keep pace with sea level rise. This is important to consider when planning the future of coastal wetlands in view of a predicted sea level rise during this century because many rivers are dammed worldwide resulting in reduced sediment supplies to estuaries and coastal wetlands. Impacts from human activities are invariably destructive to coastal ecosystems as a general principle, whether inadvertent or in the course of ecosystem management. However, they often provide insight into the workings of relatively unaltered or pristine ecosystems. (The term “pristine” in this context is a relative term, and global change would argue that there are no remaining examples that warrant this status). We summarize effects under three categories: those resulting from global change, effects from more localized physical and chemical changes, and effects from biotic sources (Table 1). Global change has three related effects: warming that may expand the range of frost-intolerant species (i.e., mangroves), accelerating rise in sea level that may exceed the capacity of some wetlands to keep pace, and the drying or wetting of climates that may alter coastal salinity and fluvial sediment delivery patterns (Kundzewics et al., 2008). In comparison with more local types of human impacts, global changes tend to be longer term and more subtle, but this feature is offset by
29
Coastal Wetlands
(a)
(b)
(c)
(d)
Figure 17 The on-going worldwide destruction or degradation of mangroves is mainly due to (a) urbanization (Mai Po, Hong Kong), (b) shrimp ponds (Surat Thani, Thailand), (c) salt ponds (Wami Estuary, Tanzania), (d) rice farms (Konkoure Estuary, Guinea), as well as (not shown) clear cutting for charcoal production.
the large geographic reach of such effects. Further, there is great uncertainty in actual rates of change. Some reports have suggested that increased interannual variation in precipitation and temperatures, as well as increased “storminess,” should be considered to have ecological consequences beyond more linear, timeaveraged projections (Michener et al., 1997). For terrestrial ecosystems, reconstruction of past climate changes indicate that forests did not shift latitudinally as intact communities, but rather individual species responded independently resulting in community mixtures not recognizable in today’s assemblages (Clark et al., 1998). It is unlikely that this will have significant consequences for coastal wetlands because of their relatively low plant diversity. Regardless, examples of asynchrony in migratory bird activity relative to the supply of seasonal foods are examples of global change that warrant scrutiny in coastal wetlands (Michener et al., 1997). We are in our infancy in the understanding of possible interacting effects of changing salinity, temperature, and diseases for wetlands. Reduced freshwater flows resulting from domestic and agricultural extractions of water from rivers have the potential to increase coastal salinities. This could and has increased the salinity in already hypersaline lagoons that have seagrasses and marshes (Buskey et al., 1997). Further, associated irrigation return flows and
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domestic wastewater discharges have additional influences in changing the nutrient loading and introducing pesticides and other toxicants. Seagrasses are the most likely to be affected by nutrient enrichment (Holmer, 2009; Broome and Craft, 2009). Land clearing activities can increase sediment delivery to coastal wetlands, while impoundments have the opposite effect; both have the potential of changing sediment balances or imbalances locally. Seagrasses are particularly susceptible to effects of suspended sediments on light attenuation in the water column. Mangrove mortality has resulted from excessive sedimentation (Ellison, 1998). Salt marshes and tidal freshwater marshes along the east coast of United States, however, owe their current existence to past increases in sediment supply (Pasternack et al., 2001). The effects of sediment starvation have been well documented as the cause of tidal marsh losses (Streever, 2001; Bernier et al., 2006). Barriers to tidal exchange interfere with the most fundamental processes in coastal ecosystems dependent on sediment dynamics. Barriers include dykes and various forms of seawalls (Adam, 2009). For example, tidal marsh areas isolated from sediment supplies and astronomic tides often undergo major subsidence. Seagrasses generally may be immune to these effects. Eutrophication affects seagrasses meadows in multiple ways, including accumulations of excessive macroalgal growth that depletes nighttime oxygen to the detriment of the plants, stimulation of sulfide production through organic enrichment resulting in toxicity to roots, and excessive periphyton growth having shading effects on plant photosynthesis (Holmer, 2009). Experimental fertilization of salt marshes has been shown to increase insect herbivory (Vince et al., 1981) although the more subtle effects of lower levels of nutrient loading have not been shown to cause changes. Experimental fertilization has been shown to differentially affect biogenic soil formation, and consequently soil elevation trajectories, across geomorphic settings of mangrove forests in mineral sediment-poor settings (McKee et al., 2007). Bottom disturbance from trawling activities is particularly destructive to seagrass beds, resulting in scars that can take decades to heal; these are the object of some restoration projects (Paling et al., 2009). While salt marshes and mangroves would be exempt from this type of disturbance, more intense bottom disturbances involving canal dredging, creation of aquaculture ponds, and so on completely alter the structure and function of these coastal wetlands. Effects often extend beyond the boundaries of these activities as a result of altering tidal reaches and organic matter fluxes, not to mention possible nutrient enrichment and pesticide loading often associated with discharges from aquaculture practices. Fragmentation within a particular habitat has been little studied in coastal ecosystems. Fragmentation is a component of the placement of tidal barriers already discussed. The loss of “between-habitat” connections may have far-reaching effects for the exchange of larvae and the other nursery functions that many coastal wetland habitats provide. Biotic effects include invasive species and diseases. While not all examples are documented to have been initiated by human activities, human agents may be responsible for accelerating their introduction, and stressful conditions created by eutrophication may magnify their effects. The overharvesting of oyster in
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Chesapeake Bay, for example, is believed to have completely changed the trophic dynamics of the estuary. Whether or not this has interfered with oyster recovery, now hampered by diseases (Harvell et al., 1999), is unresolved. While oyster habitat would be classified more as reef structures than a coastal wetland type, influence of the species on sediment dynamics, biogeochemistry, and species composition make it a keystone species. Both native species and exotics can be aggressive invaders that rapidly change the structure and dynamics of coastal wetlands. Phragmites australis, for example, is capable of overtaking the spring tidal portion of salt marshes by outcompeting the shorter Spartina patens, the original dominant (Windham and Lathrop, 1999). The aggressiveness of Phragmites communis is attributed to the development of an ecotype genetically dissimilar to the tamer native type (Vasquez et al., 2005). In the case of Caulerpa taxifolia, a submerged macrophyte, clones in the Mediterranean and California (USA) change the structure of benthic communities for both seagrasses and tidal flats. This not only interferes with fishing practices and recreation, but fundamentally alters food webs of lobsters (in California) and displaces eelgrass beds (Williams and Grosholz, 2002). The aquarium trade provides a fertile pathway for the dispersal of exotics. As with oyster diseases, the role of human influence on the incidence of other diseases is uncertain. Nevertheless, it is worthwhile to reveal examples that may be facilitated by human activity. A major loss of eelgrass beds that occurred in the 1930s has been attributed to a fungal pathogen that caused “wasting disease” (Orth et al., 2006). This had effects over large areas along the coasts of Europe and North America (Holmer, 2009). Eelgrass can be considered another keystone species because of its role in the life cycle of the scallop fisheries in coastal lagoons, its protective habitat role for blue crab in the vulnerable stages of molting, and the nursery function for juvenile fishes.
2.10. Modeling and predictions In view of the perennial changes in wetlands, and the acceleration of these changes because of human impacts and climate change, there is a need to obtain models that can predict the future evolution of coastal wetlands. The models must be able to distinguish between two evolution pathways, namely a “state change” in contrast to “ecological succession.” Succession repeats vegetation patterns after disturbance, while state change leads to new patterns normally as a result of some underlying physical change in the ecosystem (Hayden, 1991). Brinson et al. (1995) for instance showed that disturbances including sea-level changes have generated state changes in coastal ecosystems, resulting in a zonation of adjoining upland forest, high marsh, low marsh, and intertidal flats. Such models need elaboration in order to transition from being qualitative to being quantitative. The result should allow prediction in changes of coastal wetlands to climate change so that adaptation strategies can be put in place (Reyes, 2009). Predictive models of the evolution of coastal wetlands are needed and remain in early stages of meeting goals for management; this is particularly true for landscape scale models. Such models need the input of oceanographers, sedimentologists,
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chemists, biologists, ecologists, hydrologists, pedologists, dendrologists, entomologists, and geneticists. Clearly models to be successful and practical must avoid getting bogged down in details while recognizing that the system is intrinsically biogeomorphic with equally important and connected biological, chemical, and physical components. Understanding the functioning of coastal wetland ecosystems depends both on the collection of good long-term sets of real-time data on the levels and fluxes of each of the significant plant nutrients and on the development of functional models of the magnitude and direction of all major organic and inorganic material fluxes. Cumulative effects on coastal wetlands are very difficult to predict. Quite a lot has already been revealed by various models; these have, amongst many findings, revealed the seasonal and interannual variations in these fluxes. In view of global climatic change, the baseline of environmental parameters (i.e., extreme climatic conditions) is changing. As a result, the adequacy of the existing data sets on coastal wetlands to address climate change needs to be re-examined. Thus, the problem of a shifting data base of climate, a key problem for humanity in planning a reliable water supply (Kundzewicsz et al., 2008; Milly et al., 2008), applies equally forcefully for the survival of coastal ecosystems. Models should also consider the interaction between adjoining coastal wetlands, such as between mangroves and seagrass, or salt marshes and seagrass, or mangroves and salt marshes, as these frequently coexist side by side though very rarely intermixed. Scientists often choose to study pristine coastal wetlands to understand and quantify the dominant processes shaping the ecosystem. They then hope to apply this knowledge to quantify the human impact in other wetlands. However, we are rapidly losing our scientific control sites because pristine coastal wetlands are increasingly a rarity worldwide.
2.11. Coastal wetland ecosystems as a component of estuaries There is often a sharp transition, within a few meters distance, from one ecosystem type to the other (Figure 18), showing that ecosystem types can coexist in the same estuary. The components of a coastal ecosystem (Figure 19) include salt marshes, mangroves, mud flats and seagrass, the river, and the estuarine waters and the coastal waters. These components exchange mass and energy with each other. This exchange is believed to be synergistic, but this effect has been little quantified yet. However, numerical models suggest that because the coastal ecosystem has that degree of robustness, it may be able to withstand some degree of human impact with minimal degradation (Jorgensen and Bendoricchio, 2001). However, numerous examples of the connection between two components of a coastal ecosystem have been studied. The earliest known connection is the nutrient outwelling from coastal wetlands that enhances the primary production in estuaries (Odum and Heald, 1972, for mangroves; Odum, 1980, for salt marshes). Other connections that have been documented include biomass transfer by water currents of plant litter between different components of the coastal ecosystem (de Boer, 2000), the acceleration of this mass transfer by migrating fish and shrimp that use different habitats in different stages of their life (Mumby et al., 2004), and the protection that coastal wetlands offer to seagrass and coral reefs by trapping riverine nutrients and
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Figure 18 Sharp transitions generally exist between various coastal wetlands. (a) Spring tidal salt marsh, neap tidal salt marsh, and upland forest (Virginia Coast Reserve, USA), and (b) seagrass and mangroves (Palau).
Tidal wetlands
River
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Figure 19
Sketch of energy and mass flows in a coastal ecosystem.
sediments (Wolanski, 2007a). Coastal freshwater wetlands and saline wetlands are also connected and each of these ecosystems is made more robust because of this connection. This is the case, for instance, of freshwater tidal wetlands adjoining mangroves in Micronesia where the mangroves were spared the stress of hypersalinity during a severe El Nin˜o drought because the supply of freshwater (through groundwater flow) from TFWs continued 6 months into the drought even after the
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Figure 20 (a) Invasion of ferns in the Mai Po mangroves, Hong Kong, and (b) accumulation of plant litter on the substrate. Compare the accumulation of plant litter in (b) with the absence of plant litter in healthy mangroves (Figures 6f ^ h and 14a).
water courses had dried out (Drexler and Ewel, 2001; Drexler and de Carlo, 2002). On the negative side, this connectivity can lead to the degradation of the whole ecosystem if one component has been severely impacted by human activities. For instance, the destruction of coastal wetlands leads to a degradation of the seagrass and coral reefs in coastal waters (Duke and Wolanski, 2001). As another example, a degraded estuary can in turn degrade adjoining coastal wetlands. In this case, the discharge of treated sewage from more than 1 million people in the small bay draining the Mai Po mangrove reserve in Hong Kong has resulted in an excess of nutrients in the bay waters and sediment. Ferns and weeds have invaded the substrate of these mangroves (Figure 20a). There is so much plant litter that the detritivores cannot consume it all (Figure 20b); the thick fern vegetation along the banks prevents the flushing of this plant litter; the plant litter accumulates and decays, releasing H2S, which in turn further reduces the crab population, which stresses the mangrove trees by inhibiting the aeration of the soil and the flushing of excess salt from the soil. The stressed trees generate less tannin, and borers use this weakness to attack the trees, resulting in severely stunted tree growth.
2.12. Coastal wetland socioeconomics Coastal wetlands provide numerous ecological services to humanity. They protect the coast against erosion and guard against loss of capital infrastructure and human lives. They are habitats that support seasonal or perennial fisheries and are vital for migratory and resident birds. Also, they provide ecological services that have socioeconomic benefits to the human population, including, according to location, fuel, forage, building material, timber, fisheries, and protection of commercial, recreational, and naval vessels (Williams et al., 2007). Mangroves provide another important ecosystem service to the population living in the hinterland, by sheltering it from storm winds and capturing salt spray (Figure 21) and thus improving crop production in arid coastal areas (Wolanski, 2007b). Coastal wetland management has typically been a failure so far in human history in most countries of the world, with enormous destruction and degradation of these
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z
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Figure 21 Sketch of the trajectory of the landward wind facing a mangrove belt and the fate of salt particles in suspension. The wind slows down as the air flows through the mangroves, the bulk of the flow is deflected upward over the vegetation, and a turbulent wake forms behind the trees; SSC, suspended salt concentration. Down arrows, deposition of salt particles. Redrawn fromWolanski (2007b).
ecosystems, resulting in the loss for humanity of the ecosystems services that they used to provide and the loss of the quality of life (Millennium Ecosystem Assessment reports; http://www.millenniumassessment.org/en/index.aspx; Frayer et al., 1983; Duarte et al., 2008; Hily et al., 2003; Bernier et al., 2006; Duke et al., 2007; Wolanski, 2007a). Scientists must find ways to more effectively communicate with stakeholders, managers, and policy-makers. Communication is difficult due to the multitude of stakeholders in coastal wetlands and fragmented governance. Possibly, communication of the coastal wetlands science results and their implications for humanity could be improved by quantifying the economic value of ecosystem services. This is very much an emerging science and there have been very few attempts to do that (Costanza et al., 1997 for wetland ecosystems in general; Balmford et al., 2002, for the case of tropical rainforests, mangroves, wetlands, and coral reefs; Barbier et al., 2008, for the case of mangroves). Wetland ecosystems are undervalued because the market system consistently ignores externalities of higher economic returns than those gained from preserving the coastal wetlands especially if they are already degraded and thus have lost their capacity to generate goods and services (Valiela and Fox, 2008). Thus, especially when the coastal wetlands are degraded and even if their socioeconomic services were fully quantified and communicated, it does not mean that the wetlands will be preserved when competing job-producing development opportunities arise. This and economic forces operating at global scales are probably the principal causes for the on-going wetland loss worldwide. There is an urgent need to quantify the socioeconomic value of coastal wetlands because they are too easily perceived as a breeding ground for mosquitoes that transmit diseases to humans (Dale and Knight, 2008). Some such diseases can be life threatening, such as malaria, yellow fever, dengue, and forms of encephalitis. Other
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(b)
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Figure 22 Disease-carrying mosquito species include (a) Aedes vigilax, (b) A. Coquillettidia linealis, and (c) Aedes Culex annulirostris (courtesy of Stephen Doggett, Westmead Hospital, Australia).
diseases are debilitating, such as Ross River virus and West Nile virus, though they can be occasionally also lethal. The link between coastal wetlands (both fresh and saline) and mosquitoes is undeniable. The saltwater mosquito Aedes vigilax (Figure 22a) breeds in mangroves (Knight, 2008). Eggs are laid on most substrates, they are nurtured under dry condition (no tidal flooding) in as few as 3 days, they are hatched when flooded by the tides, the larvae develop in small water holes that occur in localized areas of poorly flushed mangroves. When the mosquitoes emerge, they may fly many kilometers. For combating this mosquito, it may be sufficient to increase the drainage pattern to avoid slow-flowing or stagnant waters (Dale and Knight, 2006); this requires careful engineering to avoid creating acid sulfate soils. There are a number of disease-carrying mosquito species (Figure 22), and the management strategy to combat them in coastal wetlands varies from species to species because of their different breeding strategies. For instance, to control the salt marsh mosquitoes Ochlerotatus taeniorrhynchus and Ochlerotatus sollicitans, which do not lay their eggs upon standing water, impoundments can be built with earthen dikes in salt marsh or mangroves and kept flooded during the breeding season to present ovipostion. The rest of the time, the impoundment should be open through culverts in the dikes to allow access by estuarine plankton communities and fish (Rey et al., 1991; Brockmeyer et al., 1997). Mosquitoes can thus be managed without having to use pesticides.
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There is clearly a need for mosquito control managers and wetland managers to communicate and integrate their efforts to sustain both wetland and human health. A variety of chemical, biological, and physical methods is possible (Dale, 1994; White and Beumer, 1997; Dale and Knight, 2008). Physical control methods that minimize the conditions favorable for breeding include modifying the hydrology by constructing runnels to connect depressions with tidal exchange, and this has been successful in salt marshes. Historically, physical methods to control mosquitoes were very destructive of the coastal wetlands (i.e., filling or impounding) but they are increasingly discarded because of their destructive impact on flora and fauna. Biological control methods include introducing fish and larval predators (White and Beumer, 1997). Historically, DDT was the preferred chemical to control mosquito; because of its impact on the environment, it has been replaced in many (but not all) countries by mosquito larvicides, including methoprene and temephos (EPA, 2002). Wetlands have often been blamed by the public for the proliferation of Vibrio cholerae, the etiological agent of cholera that, despite progress in medicine, still exists in over 90 countries (WHO, 1998). The combination of a brackish wetland environment, rich in organic matter and with a high density of human population, represents ideal conditions for V. cholerae. Although the high human density near some wetlands amplifies the transmission, it may not be its primary cause because V. cholerae is part of the autochthonous flora of brackish and estuarine environments. Cholera epidemics can also be linked to plankton blooms, rise in temperature, and El Nin˜o southern oscillation. Outbreaks can occur after natural disasters (Colwell, 1996). These additional links may explain why the disease is established along the Bay of Bengal and not along the Amazonia coast even though human and wetland settings in these two regions are similar (OCHA, 2007; R. Lara, personal communication). Climate change may exacerbate the problem of mosquito-borne diseases (WHO, 2000). Increased temperature may speed up the development time for mosquitoes and pathogen cycles. This may extend their distribution to higher latitudes so that presently cooler wetlands may become disease vector mosquito habitats. Some infectious diseases, such as malaria, yellow fever, and dengue, are believed to be associated mainly with wetland conditions. Dengue and yellow fever were common from the 17th century onward in the United States, with yellow fever killing tens of thousands of people as far north as New York City (Reiter, 1996). Temperature and favorable habitat for vectors may be a necessary but not sufficient condition for disease prevalence in human populations. Public health practices and lifestyles have a greater influence on the spread of diseases than the temperature tolerance of their vectors (Marshall, 1997). Increased incidence of West Nile virus in recent years (Bourgeade and Marchou, 2003) is apparently unrelated to changes in temperature. In a worst case scenario, the melting of permafrost in northern latitudes would be likely to create larger areas suitable for mosquitoes. Coastal wetland ecosystems should not be viewed only for their ecosystem values and services but also as a key component of the total ecosystem composed of the river basin, the river, and the estuary and coastal waters. Thus, for human needs, an ecohydrology approach is needed for integrated management across ecosystem types (Wolanski, 2007a). There are virtually no examples in the world of such an approach being implemented, or if they do occur, outcomes of their
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success or failure have not been widely communicated. Instead, management is still regulated by social and political boundaries that ignore the reality of the watershed geography and hydrology or by humanity’s historical habits of managing individual users (water resources, water supply, irrigation, hydroelectricity, forestry, farming, and fisheries) without integration. This invariably leads to environmental degradation and the loss of ecosystem services. A new ecohydrology (i.e., holistic) management strategy is clearly needed. The restoration of functional coastal wetlands is a combination of evolving multidisciplinary science and engineering practices based on analysis of past failures and success. The aim is to create coastal wetlands that have the same appearance and provide many, if not all, of the ecosystem services of natural wetlands. Success is improving for seagrass, but the size and location of projects are greatly constrained by high costs and appropriate environmental conditions (Paling et al., 2009). Success is the norm for salt marsh and TFWs, in part because hydrology (i.e., sea level and tides) is predictable and practices have been ongoing for decades (Baldwin et al., 2009; Broome and Craft, 2009; Garbutt and Boorman, 2009). Success for mangroves is improving, but nevertheless is costly (Lewis, 2009). The reasons for failure vary from site to site. Failures can be as simple as restoring incorrectly the hydrology (Figure 14c) or the weight of mats of algae physically bringing down emerging plants and seedlings (Figure 14d), or simply failing to learn from nature and planting the wrong species in the wrong places (Primavera and Estaban, 2008). They can be more complex and involve the interaction between a number of ecological processes and/or invasive species. Restoring coastal wetlands in exposed sites remains a challenge; the initial goal is to ensure through engineering means (using bamboo curtains as in Figure 14e or seawalls as in Figure 14f) the survival of the emerging vegetation against wave and erosion until it is large enough to be able to survive and attenuate waves.
2.13. Coastal wetlands are essential for our quality of life It may be unrealistic to imagine that humanity would act to preserve and cherish coastal wetlands simply because they are beautiful. As editors, we recognize that science alone, regardless of how compelling the evidence, is unlikely to motivate societies to protect coastal wetlands at the expense of numerous competing shortterm gains. In our experience, science does not provide the political will for such action. In contrast, aesthetic appreciation and spiritual values are often strong motivators for action. The environmental movement is not just facilitated by necessity based on the science but also by a sense of beauty that is vital for maintaining the quality of life of the human population. The human need for beauty should not be underestimated as a prime factor in the preservation of coastal wetlands (Dorst, 1965; translated from French): Nature will only be preserved if man loves it a little, simply because it is beautiful, and because we need beauty whatever is the form we are sensitive
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(a)
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Figure 23 Coastal wetlands offer spectacular landscapes such as (a) the mangrove-fringed Hinchinbrook channel, Australia (courtesy of H. Yorkston), and (b) the salt marshes of the Virginia Coast Reserve, USA.
to as a result of our culture and intellectual formation. Indeed this is an integral part of the human soul. Coastal wetlands offer spectacular landscapes for humanity to enjoy (Figure 23). Intangibles such as existence value relate to some of the more popular images of coastal wetlands including birds and other charismatic wildlife. Coastal wetlands are becoming increasingly important as last refuges to wildlife (Figures 24 and 25) faced with the human coastal squeeze. Thus, coastal wetlands are increasingly becoming the last remaining coastal biodiversity hot spots worldwide. At less disturbed sites, large, charismatic animals still freely migrate back and forth between the land and the coastal wetlands, such as along the Arctic coast (Figure 26).
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Figure 24 Charismatic wildlife found in coastal wetlands include (a) chimpanzee (Conkouati mangrove reserve, Congo-Brazzaville; courtesy of S. Thomas), (b) crab-eating Macaque in Rhizophora apiculata mangroves (Ranong, Thailand; courtesy of N. Duke), (c) dugong in seagrass along the Queensland coast (courtesy of J. Freund and S. Freund), (d) mudskipper and saltwater crocodile (Queensland mangroves; courtesy of M. Read), (e) red colobus (Chwaka Bay, Zanzibar; courtesy of Farhat Mbarouk), (f ) spotted deer (Sundarban, Bangladesh; courtesy of P. Dyas), (g) Sundarban tiger (Bangladesh; courtesy of P. Dyas), and (h) hippopotamus (Wami Estuary, Tanzania; courtesy of H. Kiwango).
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Figure 24
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Figure 25 Coastal wetlands support an enormous diversity of resident and migratory birds including (a) snow geese, often found in large flocks on coastal wetlands where they raise their broods (Alaska, USA; courtesy of W. Streever), (b) heron (Cedar Key, Florida, USA), (c) yellow-billed storks (Wami Estuary, Tanzania; courtesy of H. Kiwango), (d) fish eagle (Congo Estuary, Angola), (e) lesser flamingoes (Wami Estuary, Tanzania), (f ) diverse shorebirds (Mai Po, Hong Kong), (g) 20,000 waders (mainly knot, dunlin, short-billed dowitcher, and semipalmated sandpiper in the foreground with a few laughing gulls in the background) feeding on horseshoe crab eggs (Delaware Bay, USA; courtesy of N. Clark), (h) a flock of knots and oystercatchers overTheWash (UK; courtesy of N. Davidson).
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Figure 25
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Figure 26 American and Canadian Arctic coastal wetlands are probably amongst the least human-impacted coastal wetland areas of the world. Large charismatic wildlife such as (a) caribou (Alaska; courtesy of W. Streever) can move back and forth between inland and coastal wetlands, and (b) polar bears ( James Bay, Canada; courtesy of R.I.G. Morrison) use coastal wetlands during the summer when the sea ice has melted.
3. LESSONS FROM THE C HAPTERS IN T HIS B OOK The following chapters in this book provide in-depth reviews of the state of knowledge of coastal wetlands by leading scientists, organized into seven main topics: Coastal Wetlands as Ecosystems, Physical Processes, Tidal Flats, Marshes and Seagrasses, Mangroves, Coastal Wetland Restoration and Management, and Coastal Wetland Sustainability and Landscape Dynamics.
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3.1. Coastal wetlands as ecosystems This section provides a global perspective on the morphology, development, and distribution of tropical, temperate, and polar coastal wetlands. Tropical coastal wetlands, comprising mangrove forests and associated freshwater wetlands, are introduced by Woodroffe and Davies who indicate that these wetlands are most extensive in low-lying coastal plain regions. The relatively species-poor communities of the Atlantic East Pacific Province contrast with the diverse forests associated with more complex habitats in the Indo-West Pacific Province, centered on southeast Asia and northern Australia. Sedimentation in these wetlands is a function of the type of substrate and the extent to which sediment sources are derived from inland catchments (e.g., terrestrial mud), the near-shore (e.g., marine carbonate), or in situ production of organic accumulation (e.g., fibrous peat). Stratigraphical and chronological records of mangrove sediments have shown that sea level is a primary control on the way in which the wetlands have developed. In the West Indies, the pattern of decelerating sea level rise has resulted in a transgressive sequence of sediments that has only recently changed to progradation of mangrove wetlands in a few locations. In southeast Asia and northern Australia, the change from transgression to regression occurred around 6,000–7,000 years ago when the sea level stabilized around its present level. Since that time, extensive coastal plains have been built and the nature of the wetlands that have developed on the landward margin of mangrove forests is climatically determined; extensive salt flats occur in arid and semiarid areas, seasonally flooded grass and sedgelands dominate the wet–dry tropics, and peat swamp forest occurs in the wettest areas. Understanding the response of these wetlands to past sea-level changes, and their association with distinct depositional environments, can yield important insights into how they may respond in the face of future climate change. Pratalongo et al. demonstrate that temperate coastal wetlands, in addition to their ecological, cultural, and aesthetic significance, play an important role in determining the sustainability of coastal populations through their dynamic interaction with physical and chemical properties of the environment including, as for mangroves, the climatic range from arid to humid. The interaction of biota, hydrology, and sediments is clearly evident in their ecological and geomorphologic characteristics. These wetlands offer perhaps the clearest illustration of the extent to which coastal habitats respond to abiotic factors such as climate, groundwater, sea level, accommodation space, sediment supply, water quality, and water and sediment dynamics – as well as human activities. The long-term resilience of temperate coastal wetlands is very much determined by their capacity to respond to
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environmental change, and it is this capacity that is heavily modified by humans. The construction of sea defenses, embankments, and other engineered structures has introduced a degree of stability for many coasts, which is in direct contrast to the dynamic response of coastal wetlands to changing sea level. Martini et al. describe polar coastal wetlands as consisting mostly of salt- and brackish-/fresh-water marshes, coastal tundra inundated by seawater during storm surges and by freshwater at the time of snow and ground-ice melt, and coastal tundra plains with numerous ponds and shallow lakes. Permafrost controls the structure and the functioning of these Arctic and sub-Arctic ecosystems. During the quaternary era, continent-wide glaciers developed and waned several times in response to alternating cold and warm periods. The biota, including human populations, adapted to these changes by migrating and recolonizing the land affected by ice. These environments and their associated biota are particularly vulnerable to global warming. Permafrost reduces the significance of groundwater activity to overall hydrology, and seasonal freezing eliminates freshwater inflows to estuaries over extended periods. The majority of human settlements in the Arctic are located on coasts where they will also be adversely affected by storms and tidal surges. The melting of the sea ice and the opening of Arctic sea routes present further possible hazards for low-lying coastal regions and their biota. In particular, a new threat that is affecting most coasts facing the Arctic is wave erosion from reduced coverage of sea ice. Wetlands should play an important role in helping to attenuate wave activity by reducing their effect but will themselves likewise be exposed to coastal erosion.
3.2. Physical processes Hydrology is a controlling factor in the development and sustainability of coastal wetlands. The dynamics of water circulation and sediment distribution are reviewed in this section of the book. D’Alpaos et al. indicate that a more comprehensive understanding of such dynamics and their predictive modeling can only be achieved through the description of the feedbacks coupling hydrodynamics and sediment transport, on one hand, and ecological processes on the other hand. They provide a comprehensive theoretical framework describing the large-scale, long-term evolution of the intertidal landscape based on an ecomorphodynamic model that allows investigation of the long-term evolution of tidal networks, the adjacent marsh platforms, and the vegetation colonizing them. Furthermore, they employ the model to explore the response of tidal ecogeomorphologies to different scenarios of changing sea level, incoming sediment concentrations, and halophytic vegetation characteristics.
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Tidal courses represent the basic circulatory system through which water, sediments, organic matter, nutrients as well as pollutants are transported in and out of these wetlands. Considering their importance, Perillo reviews the geomorphology of tidal courses through coastal wetlands suggesting a new denomination and classification. Tidal courses are commonly one of the first features that appear upon the formation of a coastal wetland. Rills, grooves, and gullies are the first to develop, and their persistence is dependent mostly on local factors such as some particular irregularity on the sediment surface or the depth of the incision. Pioneer plants help stabilize the soils and may lead to the formation of salt marshes and/or mangroves. Levees tend to form even on tidal flats where preferential sedimentation along courses stabilizes the channels. Levees may act as dams allowing the formation of ponds where burrowing activities by crabs in turn modify the geomorphology. Windwave erosion enlarges ponds and the local water circulation over tidal flats; the effect on salt marshes and mangroves is smaller. When levees are breached by tidal creeks and gullies, tidal circulation is restored. The formation and evolution of meanders in tidal creeks remains poorly understood. Piccolo reviews the studies of heat energy balance, most of which have been carried out in temperate environments and mainly in the open waters of estuaries and tidal flats. Much less work has been done in vegetated coastal wetlands. Latent heat flux is a major component of energy balance and contributes to water balance through evapotranspiration, usually the most troublesome flux to estimate reliably. Precipitation and tidal flooding replace this water in a spatially and temporally varied pattern with the net result contributing not only to the heat flux but also to salinity patterns. Latent heat flux is important in regulating leaf temperatures and becomes constrained when water potential decreases, with a shift to radiative and sensible heat losses. Weather that brings changes in the temperature of air masses also can reverse the balance between latent and sensible heat flux. The chapter reviews studies conducted from low to high latitudes and points to differences along the temperature gradient. Even though there are some studies of the heat-driven interaction between the biological and physical processes, an exhaustive and conclusive analysis of the basic interactions between flora, fauna, and heat stored in coastal wetlands is still lacking. Mazda and Wolanski apply oceanographic modeling approaches to mangroves. Mangroves often constitute a very wide intertidal area, which is flooded at high tide and exposed at low tide; thus, water movement in such an area cannot be treated as a continuous flow throughout the cycle of a tidal period. The system is drained by tidal creeks. The swamps themselves present a large number of obstacles to water flow, including trees trunks, pneumatophores and prop roots, and leaves that affect the horizontal and vertical hydrodynamics. Below
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ground, there are many macropores created by animal burrows and decayed roots, through which surficial ground water circulates and facilitates an interaction between the swamp, the estuary, and the open sea. The periodicity of the dominant water motions through mangroves includes seasonal (months), storm surges (days), tides (12 and 24 h), sea waves (less than 20 s), tsunami waves (10–30 min), and turbulence (1–5 s). The mangrove ecosystem is sustained through feedback processes among biotic actions surrounding mangrove trees, landform with peculiar three dimensional topographies, water flows over and under the ground, the atmosphere, and the vegetation sheltering the fauna from the wind and solar radiation. Further modeling of the geomorphologic evolution of coastal wetlands through feedbacks coupling hydrodynamics and sediment transport, on the one hand, and ecological processes on the other hand are described by D’Alpaos et al. The model decouples the initial rapid formation of tidal channels in a tidal flat, and their subsequent slower evolution by meandering, from the even slower evolution of intertidal platforms. The latter is strongly influenced by the vegetation. The model effectively reproduces qualitative observations of complex channel network structures, tidal circulation patterns, and general patterns of vegetation distribution in salt marshes. Model validation for the vegetation biomass remains elusive and the reasons for that are explained and solutions are proposed. The biogeomorphological modeling of coastal wetlands is still in its infancy but looks promising.
3.3. Tidal flats Tidal flats are distributed widely along the world’s coastlines and accumulate fine-grained sediments on small bed slopes. In essence, they are the basic structure upon which most coastal wetlands build. Gao reviews their hydrodynamics and sediment dynamics. He identifies three classes based on flooding and wave action: spring tidal and neap tidal, and a supratidal class that is flooded only by storms. Tidal currents are strong on the tidal flat, resulting in high mobility of bed materials that is facilitated also by the dissipation of wave energy as long as the waves do not break. Gao also describes the progress in understanding the characteristics, processes, and system evolution of tidal flats, focusing on sedimentation processes and associated physically and biologically induced sedimentary structures. Soft sediment systems, basically tidal flats, are extreme habitats for the organisms that inhabit the system where the moving substrate, surging waves and tides, and extreme temperature and salinity fluctuations impose considerable physical challenges. Paterson et al. review the functioning of these ecosystems by showing how organisms
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inhabiting depositional environments are successful by using a number of strategies to resist, tolerate, and exploit, and ultimately to stabilize these harsh environments. They compare sandy and muddy sediments with regard to in situ productivity and dependence on allochthonous sources for organic matter. Ecosystem processes and functions of soft sediments are framed in the context of ecosystem services that relate to mankind. The temporal and spatial heterogeneity and strong forcing by tidal and wind dynamics generates extreme biogeochemical variability in coastal tidal flats analyzed by Joye et al. Benthic intertidal flats play important roles in cycling of carbon, nitrogen, phosphorus, and silica in coastal ecosystems, and this cycling is modulated by physical (tides and storm) forcing and the position of redox gradients in the sediments. Enhanced fluid exchange at low tide may stimulate N, P, and Si exchange and return of these nutrients to the overlying water column. The authors identify the need for focused research in a broad array of intertidal flat habitats on a number of important processes that control biological dynamics.
3.4. Marshes and seagrasses Holmer reviews biogeochemical cycling in seagrass meadows. The roots and rhizomes of seagrasses typically penetrate at least 15 cm into sediments. The active surfaces of roots contribute to a dynamic sediment layer where biogeochemical processes are highly modified by the activities of plants. Organic matter from the seagrasses stimulates microbial processes that lead to the development of highly reduced sediments. A major adaptation to reduced environments is the release of oxygen from root tips that is rendered harmless through oxidation of toxic sulfides in sediment porewaters. Human pressures on the coastal zone, especially eutrophication, have exceeded the capacity of the affected plants to adapt by increasing the stress of greater production of sulfide and simultaneously reducing water column transparency to light through increased phytoplankton production. Seagrass decline has been observed along most coastlines with increasing human activities. Seagrass ecosystems face a global crisis due to human pressures in the coastal zone, and using her vast experience, Holmer provides suggestions for future researchers to address this issue.
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Earlier classifications of the geomorphology and sedimentology of tidal salt marshes are the first basis of Allen’s review. He also describes the geographical distribution and reasons for the existence of salt marshes; evolution versus inheritance in salt marshes; salt marsh edges, terraces, creeks, and drainage networks; channel and platform flows; the response of salt marshes to tidal regime, sediment supply, compaction, and water-level change; accretion over salt marshes; and grainsize, bedding, and lithostratigraphic architecture in salt-marsh deposits. Allen shows that most salt marshes arose during the Holocene, an epoch of changing and fluctuating climate and sea level, and that change on decadal to millennial scales continues. He argues that more attention should be given to the developmental history of contemporary marshes to emphasize the importance of relative maturity and of inheritance versus evolution. This would provide a detailed picture of the circumstances under which salt marshes have arisen in the recent past and reveal how they have responded to conditions that have changed on subdecadal to millennial time scales. It would provide clues on how salt marshes will change as a result of climate change. Salt marshes have a complex ecosystem structure as presented by Visser and Baltz who describe fauna and flora, and the interaction among trophic levels. They reveal a multitude of bottom-up and top-down ecological controls. For example, besides predation, many factors acting alone or in concert may influence the distribution and abundance of invertebrates and their predators, primarily crustaceans, fishes, and birds. These include density-dependent processes, selective larval settlement or mortality, physical gradients that influence habitat selection, and both unpredictable and cyclical physical disturbances. Spartina detritus and phytoplankton are important carbon sources for macroconsumers. Killifish and mud snails rely largely on Spartina while filter feeders such as oysters and mussels consume plankton. Nitrogen fixation by the epiphytic community on standing dead stems of Spartina alterniflora contributes significantly to the nitrogen supply in Spartina marshes at rates comparable to cyanobacterial mats. For saline marshes, Tobias and Neubauer focus on the exchanges of elements with the estuary as well as pathways of internal cycling. In particular, the authors review the nutrient demands of competing processes by, for example, contrasting plant uptake with estimates of elemental import. They point out the importance of
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benthic microalgae and macroalgae in contributing to total autotrophic production, especially in arid climates. Chemoautotrophy is quite prevalent in these wetlands because of the abundance of reduced compounds, such as Fe(II) and sulfides, that themselves are a product of the decomposition of photoautotrophic production. Although allochthonous sources of particulate organic carbon to marshes have been well documented, the long-term fate of this material is poorly known. As for freshwater tidal marshes, the potential role of humic acids, interacting with alternate electron acceptors, is a rather new area of interest. They point out that export of dissolved inorganic carbon and CO2 can have significance not only for saline wetlands but also for balancing the carbon budget of estuarine waters. And as for freshwater tidal marshes, nitrogen cycling is a relatively closed process in spite of frequent tidal exchanges. Depending on the adjacent landscape, some marshes intercept flows of nitrate in groundwater that would otherwise contribute to estuarine eutrophication. As indicated in the physical chapters at the beginning of this book, plants stabilize the marsh platform, a biotic feedback that is fundamental to providing an environment for biogeochemical processes to occur. Boorman reviews the ecohydrology of salt marshes with emphasis on British coastlines. He shows that the primary productivity of salt marshes contributes to food chains both in the marshes themselves and in adjacent marine habitats. He provides an overview of the methods for measuring freshwater flows, quantifies their importance on salt marsh growth and development, and describes the implications for the management both of salt marshes and adjacent estuaries. Seepages of fresh groundwater control plant zonation. While eutrophication of estuaries and coastal waters is usually due to excessive nutrient loading of surface waters, groundwater flows can also transport these nutrients over considerable distances with long-term effects. Generally, however, salt marshes control the eutrophication of coastal waters by removing excessive nutrients. Groundwater movements can also contribute to fluxes of potentially damaging pollutants in the form of agricultural and industrial chemicals, notably pesticide and herbicide residues and various heavy metals. Significant levels of these pollutants have been shown to occur in salt marsh sediments and vegetation with the possibility of damaging effects along food chains. The implications of fresh water pathways and their effects in humid climates must be included to improve the performance of salt marsh models. TFWs are abundant in many estuarine systems, particularly in the northern hemisphere. Whigham et al. point out the absence of a global inventory of the distribution and abundance of TFWs, a clear sign that these ecosystems have been largely overlooked by the scientific community. TFWs have been eliminated or degraded over much of their historical distribution in northwestern Europe and they
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have suffered the same fate in many areas along the east and west coasts of the United States. Most of the destruction and alteration occurred gradually over long periods of time as coastal cities and ports developed in Europe and North America. Only in recent decades have ecological studies of TFWs been conducted. Because they are influenced by tides but do not have salt stress, they are rich in species, especially plants. They also are important nursery areas for aquatic organisms and they provide important water quality benefits within the tidal freshwater zone of coastal estuaries. As a result of recent studies demonstrating the importance of TFWs, attempts are underway in Europe and North America to preserve and restore them. In addition to continued threats associated with coastal development, TFWs in more remote areas are likely to be impacted by factors associated with global climate change. For the biogeochemistry of TFWs, Megonigal and Neubauer use their own results from a site in Virginia, USA, as a basis for elaborating on the carbon budget. Their data on repeated measures of carbon dioxide and methane exchange with the atmosphere provide a temporal scale similar to that of measurements of allochthonous exchanges with the estuary. In this way, a relatively complete carbon budget can be constructed that allows estimation of exchanges across the soil–atmosphere interface, between the sediments and the surface water, and below the active root zone as burial. This partitioning places in perspective the relative importance of exchanges with the atmosphere and with the estuary. The literature regarding the regulation of carbon metabolism is reviewed, and in particular the roles of alternate electron acceptors. Plants play a major role in microbial metabolism through the transport of oxygen to sediments. For nitrogen, internal cycling in sediments and in organic turnover of plant biomass indicates that these wetlands rely little on external supplies to drive primary production. As with other wetlands, phosphorus behavior is controlled more by chemical and physical mechanisms than is the case for nitrogen. There are major differences in elemental cycling between low and high marshes because of differences in tidal flushing, exposure of surface sediments to the atmosphere, and plant species composition. The relative importance of freshwater tidal wetlands as sinks for watershed-derived inputs varies greatly because of large differences in their extent of coverage geographically.
3.5. Mangroves Ellison describes the historical development of mangroves. Intertidal microtopographic gradients control mangrove distributions primarily through frequency of tidal inundation, along with variation in wave energy and salinity, which results in zonation of mangrove species that are controlled by the
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sea-level position. Ellison demonstrates that mangroves play a geomorphological role by promoting sediment accumulation and so influence the topography of their intertidal settings, which are controlled by a range of factors such as tidal range, wave energy, and fluvial influences, as well as sediment sources, storm or tsunami influences, and sea-level trends. Sedimentation rates in mangroves vary as a result of these influences that ultimately affect their resilience to sea level rise during the Holocene. She shows that when the sedimentation rates under mangroves did not keep up with sea level rise rates, old mangroves drowned and new mangrove forests colonized higher elevations. Mangroves are thus relatively sensitive to changes in sea level and frequency of inundation. Mangroves as coastal biogenic sedimentary systems stabilize massive amounts of sediment to benefit near-shore coral reefs. Lara et al. describe the major factors leading to the development of salt marshes and mangroves, depending on inundation frequency. They show under what settings a coastal wetland can be a sink for one nutrient and a source of another, a sink for an inorganic form of a nutrient and source for an organic form of the same nutrient. They discuss seasonal and interannual variations in the direction and strength of nutrient flows. Import and export can also be achieved by other routes than tidal currents, such as N2-fixation (gain) or denitrification (loss), and the role of the benthic fauna such as crabs. Crabs can process up to 80% of leaf litter in mangroves. In turn, the distribution of these crabs is often related to sediment softness and presence of channels and thus depends on the geomorphology, topography, and hydrology of these ecosystems. Thus, depending on the local settings, mangroves and salt marshes can be net exporters of organic matter and, simultaneously, net consumers of N and P. Alongi reviews a number of ecosystem properties of mangroves. He finds that (1) rates of mangrove primary productivity rival those of other tropical forests; (2) mangrove forests are architecturally simple, but factors regulating their structure, recruitment, and growth are complex; (3) mangrove growth is not constant, but related to climate patterns; (4) tree diversity is low, but faunal and microbial diversity can be high.; (5) plant–microbe–soil relations are highly evolved and efficient; (6) crabs are keystone species influencing mangrove function and structure; (7) arboreal communities are important players in animal–plant relations; (8) algae, and not just detritus, are a significant food resource; (9) mangroves are an important link to fisheries; and (10) mangroves are chemically diverse and a good source of natural products.
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Twilley and Rivera-Monroy review nutrient biogeochemistry of mangrove wetlands as either a nutrient source, sink, or transformer depending on the net flux across two mangrove boundaries, the atmosphere, and tidal exchange. To do that, they use five models (conceptual and numerical) to compare and contrast the biogeochemistry of mangroves across muddy, estuarine, and carbonate coasts of tropical and subtropical regions. They develop a hierarchical classification system to describe patterns of mangrove structure and function based on global (temperature), geomorphological (regional), and ecological (local) factors that control the concentration of nutrient resources and regulators in soil along gradients from the estuarine fringe to more interior locations from shore.
3.6. Coastal wetland restoration and management As the world’s coastal wetlands become increasingly degraded through both natural and human activities, it is essential that we develop and expand successful wetland restoration practices. The principles of wetland restoration and criteria for restoration success are reviewed in this section of the book. Paling et al. review the practices, successes, and failures in seagrass restoration. Worldwide, there are many possible causes, both human and natural, for the losses of seagrasses. Until these reasons are understood and dealt with, seagrass restoration efforts may often be futile. Restoration experiments and projects using different seagrass species have been attempted with varying degrees of success mainly in the United States, Europe, and Australia. While there have been many failed projects, some successes have been reported in restoring small areas of lost or damaged seagrasses, particularly with faster growing species such as Zostera marina (eelgrass), Halodule wrightii (shoal grass), and Syringodium filiforme (manatee grass). In areas that have lost seagrasses due to eutrophication, restoration should not be attempted until the source of the decline, namely the reduction in water column transparency, is addressed. Valuation approaches have been attempted in the United States to quantify habitat gains and losses over space and time as part of programs to offset damaged resources. The temporal component deals with the losses of goods and services that occur while functional replacement is being achieved as the seagrass meadows develop to maturity. Seagrass restoration, however, is still an evolving technology that, globally, remains quite difficult and challenging.
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Broome and Craft provide a state of the art review of salt marsh creation. Loss of salt marshes occurs as a result of dredging, filling, tidal restrictions, subsidence, and erosion. To mitigate those losses, techniques have been developed to create marshes on sites where they did not previously exist. The goal of tidal marsh creation is to provide habitats similar in structure and function to natural marshes. Since tides are the controlling abiotic factor of tidal marshes, the most critical requirement for creating new marshes is constructing a site at the correct elevation relative to the local tidal regime. Other key site-related factors include slope, drainage, wave climate, currents, salinity, and soil physicochemical properties. Cultural practices include options for selection of native plant species, seed collection and storage, seedling production, site preparation, soil testing, fertilization, handling of transplants, timing of planting, plant spacing, control of undesirable invasive plants, and maintenance until the marsh is self-sustaining. Criteria used to define successful tidal marsh creation are often controversial. Plant communities may be equivalent to natural reference marshes in a few years, while other characteristics, such as soil organic matter and numbers and species of benthic invertebrates, require much longer to reach equivalence. When marsh creation technology is properly applied, tidal marshes can be created with many of the same attributes as natural systems. To achieve functional equivalence in a reasonable length of time, care must be taken to establish the target plant communities. Important steps are site selection and preparation, assessment of a reference marsh, obtaining plant propagules, proper planting, plant spacing, providing adequate nutrients from fertilizers if natural nutrient supplies are low, and maintenance. When sound principles of ecological engineering are applied, tidal marshes can be created that have the same appearance and, with time, provide many of the functions and values of natural marshes. Adam reviews engineering and biological techniques used in salt marsh restoration and evaluates their effectiveness from a management viewpoint. Such restoration has frequently been conducted without clear objectives. In the absence of clear objectives, assessment of success is impossible and there are many ways in which projects can fail. Setting objectives for long-term survival of restoration sites requires consideration of future environmental conditions; global climate change is likely to impact salt marshes as a result of the direct effects on species distributions by warmer temperatures and indirectly at many locations from the effects of rising relative sea level. Increased carbon dioxide concentration may also change the relative abundance of species. Disturbance factors are key considerations for effective restoration and include the effects of changed hydrology, introduced species, grazing, and pollution.
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Garbutt and Boorman review the science and technology behind managed realignment, which means recreating intertidal habitats on wetlands that had been previously disconnected from sea-level controls and tidal inundation. Rising sea levels and changes in global weather patterns, coupled with the high cost of maintaining sea defenses, have led coastal managers to look for more cost-effective and sustainable methods of flood protection. Managed realignment, consisting of the landward adjustment of coastal defenses and subsequent tidal inundation of agricultural land to create intertidal habitats, has been increasingly used since the early 1990s to fulfil flood protection requirements throughout northwest Europe. Managed realignment allows tidal ingress through a simple breach in a flood defence, reintroducing tidal inundation to formerly enclosed land often for the first time in centuries. By using the results of case studies, the authors have shown that this technique of coastal management can quickly produce intertidal mudflats that are colonized by invertebrates and, given the appropriate elevation, salt marsh plants. The method has the twin benefits of reducing sea defence maintenance costs and recreating intertidal habitats lost elsewhere to land claim and erosion, a statutory requirement under European environmental law. Lewis discusses the pros and cons and the practice of mangrove restoration. Successful mangrove forest restoration requires careful analyses of a number of factors in advance of attempting actual restoration. First, for a given area of mangroves or former mangroves, the existing watershed needs to be defined along with hydrologic alterations that may have taken place. Second, careful specific site selection must take into account the history of the site. Third, clearly stated goals and achievable and measurable success criteria need to be defined and incorporated into a proposed monitoring program. Restoration methodology must acknowledge the history of routine failure in attempts at mangrove restoration and propose the use of alternative, proven successful techniques. Finally, after the initial restoration activities are complete, the proposed monitoring program must be initiated and used to determine whether the project is achieving interim measurable success to indicate whether any mid-course corrections are needed. Baldwin et al. provide an evaluation of restored TFWs. While there have been a number of recent TFW restoration projects in North America, little information on these systems has been compiled. Furthermore, criteria for successful restoration of TFW are lacking. The most common restoration approaches used involve either excavation or placement of dredged sediment to restore tidal hydrology compatible with vegetation
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establishment. As for most types of wetland restoration projects, a majority of TFW restoration projects include postconstruction vegetation monitoring. However, a number of variables have also been studied including geomorphology, hydrology, salinity, soil seed banks, fish, birds, and invertebrates. Based on a review of assessment approaches and monitoring studies, the authors propose criteria for evaluating the success of TFW restoration projects. A novel approach taken in this review is to define success within the context of reasonable expectations at restored sites given limitations such as substrate characteristics or constraints imposed by intensive urban or agricultural land use. In a case study, these authors apply these criteria to evaluate restored TFW in the Anacostia River watershed (Washington, DC) in comparison with urban and nonurban reference sites. They conclude that while successful restoration of TFW is challenging, restoration success can be improved by using a coordinated approach of construction and vegetation establishment.
3.7. Coastal wetland sustainability and landscape dynamics Insights into future wetland sustainability through modeling techniques, and landscape dynamics on the expansion of mangroves into salt marsh environments, are provided by the reviews in this section. Rybczyk and Callaway review the state of knowledge of surface elevation models. Primary production, decomposition, and mineral matter accumulation are all feedback functions of relative elevation. The elevation rises with inputs of matter, which occurs at a time scale of 1 year for autochthonous organic matter production (plant growth) to 100 years for the input of mineral matter from rivers and occasional hurricanes. The elevation responds at timescales of 10 years for the decomposition of labile organic matter and hundreds of years for the compaction and consolidation of shallow sediments and the decomposition of refractory organic matter. The compaction of deep sediments and the secondary consolidation of sediments, as well as the geosyncline downwarping and eustatic motion, occur at scales of thousands of years. Models of the elevation of coastal wetlands yield encouraging results and the authors put forward a number of suggestions for improving these models from their extensive experience. Saintilan et al. review salt marsh– mangrove interactions where they coexist in the intertidal wetlands of many temperate latitude coastlines. In these settings, mangroves may be close to physiological limits of tolerance in relation to a range of environmental variables including
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temperature, salinity, and inundation frequency. Changes in the distribution of mangroves and salt marsh can thereby provide insights into the effects of climatic variability over a range of timescales. Two case studies are presented in detail. In southern United States, mangrove–salt marsh interactions are influenced by frost frequency. In southeast Australia and New Zealand, widespread encroachment of mangrove into salt marsh has been linked to relative sea level rise. The implications of these trends are discussed in the context of anticipated increases in temperature and sea level over the coming centuries. In many places where mangroves and salt marshes coexist and competitively interact, their interactions can be studied over a range of timescales. Stratigraphic and palynological evidence has been used to reconstruct distribution of mangroves and salt marsh communities over geological timescales with the greatest clarity emerging from the Holocene. At this scale, interactions between mangroves and salt marshes are governed by geomorphic processes, most notably patterns of sedimentation following the postglacial marine highstand. Reyes reviews the state of knowledge of spatially explicit landscape wetland models. Such models are needed to assess coastal wetland sustainability and landscape dynamics. The focus was on models that incorporate ecological dynamics and feedbacks into the landscape to represent ecosystem processes. In most cases, the watershed and local/regional models incorporated new modeling approaches on how to link ecosystems, communities, and populations to the environmental tableau. His review emphasizes new methodologies and state-of-the-art developments. He gives particular emphasis to landscape models that simulate long-term changes due to climate change, rising sea level, and changes in land use/land cover patterns. He shows the wide range of approaches to constructing physical and ecological models and the compromises needed to arrive at realistic models.
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C H A P T E R
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T HE M ORPHOLOGY AND D EVELOPMENT OF T ROPICAL C OASTAL W ETLANDS Colin D. Woodroffe and Gareth Davies
Contents 1. Introduction 2. Mangrove and Associated Wetlands 3. Sedimentation and the Development of Wetlands 4. Sea-Level Controls on Wetland Development 5. Sea-Level Change and the Diversification of West Indian Mangroves 6. Sea-Level Change and the Evolution of Mangrove Habitats in the IWP 7. Impact of Future Climate and Sea-Level Change 8. Summary and Concluding Remarks References
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1. INTRODUCTION Tropical coastal wetlands represent particularly productive ecosystems that contribute to both terrestrial and marine biodiversity. In this chapter, the principal wetlands that will be considered are associated with mangrove shorelines. Mangroves characterize the upper intertidal zone on many low-energy tropical coasts, often with salt marsh and associated wetlands that can form landward of such halophytic vegetation (Robertson and Alongi, 1992; Alongi, 1998). Seagrass, which can be extensive seaward of mangroves and in other near-shore settings, is beyond the scope of this chapter. There are several factors that have favored the development of these extensive coastal wetlands over the past 7,000 years (the mid–late Holocene). In particular, the formation of low-lying coastal plains has been in response to the pattern of relative sea-level change. For such wetlands to develop, it is generally necessary for extensive near-horizontal topography to be available at sea level. This occurs most frequently as a consequence of a particular set of geomorphological conditions that have accompanied a relatively stable period of sea level. It is sea level, or the level of the water table that is generally closely related to (though slightly above) sea level, that constrains wetland development. In some cases, modern sea level fortuitously Coastal Wetlands: An Integrated Ecosystem Approach
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coincides with a horizontal substrate. In other instances, the near-horizontal topography is the result of landforms built during former Pleistocene sea-level highstands, or a complex of habitats resulting from deltaic-estuarine sediment deposition and Holocene coastal plain development. Coincidence of sea level with suitable topography is critical.
2. MANGROVE AND ASSOCIATED W ETLANDS Mangroves can be defined as trees, shrubs, or palms, exceeding 0.5 m in height that occur in the upper intertidal zone. They are not a single taxonomic group but rather a diverse range of plants with adaptations enabling survival in this otherwise inhospitable saline and anaerobic environment. Adaptations include viviparous propagules, for example, in many genera seeds remain attached and germinate on the tree and then are buoyant during a short aquatic dispersal phase. Many mangroves have developed mechanisms to tolerate salt, and the majority have root systems that enable the plants to respire despite being anchored in saturated, non-porous soils depleted of oxygen. Above-ground root systems include pneumatophores, prop roots, and buttresses, some of which provide structural support and most of which are covered with lenticels that promote gas exchange. There is a significant climatic control on the distribution of tropical coastal wetlands. Mangroves are predominantly tropical although extending into subtropical regions along the eastern coasts of major continents where ocean currents ameliorate temperatures. Although salt marsh does occur in the tropics, it usually represents a minor component in comparison with mangroves but becomes increasingly important toward the poleward limits of mangroves (Alongi, 1998). Temperature controls distribution in a broad sense, mean air temperatures of the coldest month having to be at least 20C. Exposure to winter frosts constrains the latitudinal limit (Woodroffe and Grindrod, 1991). The genetically distinct population of the southeast Australian and New Zealand mangrove Avicennia marina var. australasica reaches its southernmost limit in Corner Inlet (38450 S) in Victoria (Duke, 1992). Coastal wetlands are more extensive in the wetter parts of the tropics, and this occurs for several reasons. First, deltas are more extensive at the mouths of rivers in high rainfall regions because greater sediment loads are eroded from the catchment and carried to the coast. Second, inundation is more common and widespread, both because of the river-fed flood waters and also as a result of direct precipitation. Mangroves do occur on arid coasts, however, and there are wetlands in arid and semiarid settings, but these appear to be limited by strongly hypersaline groundwaters (Semeniuk, 1983). Mangrove forests proliferate in wetter environments in terms of both areal extent and species diversity. For example, the wetter east coast of Australia has 20 species of mangrove compared to only four at comparable latitude on the drier west coast (Duke et al., 1998). The global distribution of mangroves comprises two provinces, one centered in the West Indies (termed Atlantic East Pacific, AEP, by Duke et al., 1998) and
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another in the Indo-West Pacific (IWP) extending from the east coast of Africa to the central Pacific Ocean. The IWP is more diverse in terms of number of mangrove taxa, as well as most other groups of organisms, such as corals. By contrast, only four mangrove species dominate in the West Indies. Rhizophora mangle is the most distinctive although there are other Rhizophora genera that occur in West Africa and South America. Avicennia germinans, termed the black mangrove, is the second most significant mangrove, generally occurring landward of Rhizophora. Laguncularia and Conocarpus are found at the landward margin and are typically less extensive. Mangrove associates in the AEP include the succulent Batis maritima and the fern Acrostichum aurem (also found in parts of the IWP). Those in South America include the genus Pelliciera and in wetter areas and upstream in estuaries, Mora oleifera. Grass and sedgelands occur landward of mangroves in the Everglades of Florida and along much of the northern coast of South America. The structure of mangrove forests is relatively simple in the West Indies (Figures 1 and 2a–c). Based on ecological studies in Florida, together with several small islands in the Caribbean, Lugo and Snedaker (1974) and Lugo et al. (1976) defined five principal types of mangrove: overwash, fringe, basin, scrub, and riverine. Their research indicated functional differences between these mangrove types, and the classification has been widely adopted for the AEP region (Bacon, 1994). Overwash mangrove occurs on islands that are overtopped by the tide, as in the Bahamas and Fringe
st
f cre
Ree
Overwash
r ve Ri
Lagoon
Scrub
Paleochannel
Riverine
Basin
Peat Mud Rhizophora Avicennia
Figure 1 Mangrove forest types distinguished in theWest Indies, illustrated schematically, and their typical occurrence in the landscape (following Lugo et al., 1976; Bacon, 1994).
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(a)
(b)
(c)
(d)
(e)
(f)
Figure 2 Examples of types of mangrove forests. (a) Fringe Rhizophora, Grouper Gardens, Twin Cays, Belize; (b) Overwash Rhizophora, Rodriguez Cay, Florida Bay, where mangrove peat has accumulated over carbonate sediments as shown in Figure 4; (c) aerial view of Twin Cays, Belize within which a range of mangrove types can be seen; (d) mangroves of the seaward margin of the McArthur River Delta, Northern Territory, Australia, following Cyclone Kathy in 1984; (e) mangrove-lined bank of King Creek, Shoal Bay, Northern Territory, Australia, showing a multispecies riverine mangrove; (f ) eroding bank of the Daly River, Northern Territory, Australia. The lumpy sediments in the lower bank region contain abundant mangrove stumps probably formed during the mid-Holocene‘‘big swamp’’phase.
on the Belize barrier reef. Fringe mangrove comprises the seaward Rhizophora zone, flooded during each tidal cycle. Basin mangroves consist of more inland stands that are less regularly flooded or drained. Scrub mangrove comprises dwarfed stands, often of Rhizophora, that appear to be nutrient-limited. Riverine mangrove occurs along the banks of rivers and has the largest trees and the highest productivity. Mangroves are much more diverse in the IWP, the center of diversity lying in Indonesia and northern Australia where there are more than 30 species. There are
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several species of Rhizophora in the IWP, but also several other genera in the Rhizophoraceae, such as Ceriops and Bruguiera. Although there is considerably greater diversity of habitats in the IWP than in the AEP and many of these show complex mosaics of mangrove species, a general zonation pattern can be recognized on open shorelines (Macnae, 1968). Usually, this comprises a seaward zone of Sonneratia, then a prominent zone of Rhizophora, often with Ceriops and Bruguiera, and a landward zone of other mangrove such as Lumnitzera and Excoecaria. There are also several species of Avicennia in the IWP, and this genus demonstrates the broadest tolerance to environmental factors, especially salinity, being able to grow throughout the intertidal zone (see Figure 5). There is greatest diversity of mangrove habitats in southeast Asia and northern Australia, reflecting geomorphological development during the mid and late Holocene under relatively stable sea level. The vegetation landward of the mangroves varies, controlled primarily by regional climatic trends. In arid and semiarid areas, as flanking the macrotidal estuarine systems of Western Australia, there are extensive saline flats, and the only vegetation that survives is low shrubs and herbs, such as samphire (Jennings, 1975; Thom et al., 1975). In the wet–dry tropics of northern Australia by contrast, there are broad convex alluvial plains, covered by seasonally inundated grass and sedgelands (Woodroffe et al., 1989, 1993). In the perhumid tropics of much of Malaysia and Indonesia, similar plains support domed accumulations of woody peat beneath peat swamp forest (Anderson, 1964; Morley, 1981; Staub and Esterle, 1994). Brackish communities of Nypa fruticans and Heritiera littoralis often dominate the transition to inland peat swamp forest, as in the Mahakam Delta. Prior to its clearing in the 1970s for intensive rice cultivation, the plains of the Mekong Delta were largely covered by forests of Melaleuca cajuputi, with Casuarinas on the sandy beach ridges; these species provide firewood and other services, and re-establishment is being encouraged, especially in Melaleucarice forest farming (Douglas, 2005). Proceeding up estuaries, there is a change in species occurrence. In northern Australia, Sonneratia caseolaris and Sonneratia lanceolata are restricted to upstream locations where they grow in the intertidal zone, but at higher absolute elevations than intertidal mangroves further downstream (Finlayson and Woodroffe, 1996). Variations in salinity along the estuaries are important physiological controls. The most salt-tolerant species tend to be the slowest growing under low salinity conditions and thus are often out-competed by less tolerant species; evidently salt tolerance occurs at the expense of competitive ability (Ball, 1998).
3. S EDIMENTATION AND THE D EVELOPMENT OF W ETLANDS Tropical coastal wetlands form near the water table under low-energy conditions and can grow on substrates ranging from mud (silt and clay) to fibrous peat (derived primarily from mangrove roots). The source and supply of sediment determine the nature of the substrate, and an organic-rich mud is typical where there is both a large supply of sediment and prolific mangrove growth. Those
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extensive coastal wetlands associated with large river deltas are built on terrestrial sediment that is delivered from the catchments. The largest area of mangroves in the world, the Sundarbans, occurs in the abandoned delta of the former mouths of the Ganges–Brahmaputra–Meghna system, down which over a billion tonnes of sediment a year is supplied each year. Mangrove and associated wetlands have also been extensive, until large-scale clearance, at the mouths of other megadeltas in Asia (Woodroffe et al., 2006). By contrast, the mouth of the Amazon has not developed a progradational delta, but riverine muds carried by longshore currents nourish a suite of rapidly migrating landforms along the northeast coast of South America (Froidefond et al., 1988). Considerable volumes of mud can also be advected into coastal wetlands from offshore by tidal processes. This is particularly true of macrotidal systems, such as those along the coast of northern and northwestern Australia, where tidal flows are large enough to entrain considerable volumes of mud and carry them in suspension into mangrove forests where they are deposited when velocities are reduced (Woodroffe et al., 1989; Wolanski, 2006a). The effectiveness of tidal pumping is demonstrated on the Ord River where reduction of river discharge as a result of dam construction to form Lake Argyle has been accompanied by rapid accumulation of tidal sediments in the upper part of the estuary (Wolanski et al., 2001). Mangroves also occur in carbonate settings. The substrate on which mangroves establish may be either biologically produced, as in the case of coral reefs, or precipitated, as in the case of calcareous muds on the Great Bahama Bank. Mangroves can colonize a near-horizontal reef platform, and successive stages of mangrove colonization and consolidation have been described in response to the evolutionary stage of reef-flat development on islands of the Great Barrier Reef (Stoddart, 1980). Sediment dynamics is a function of reef growth by corals and associated organisms, but sediment accumulation beneath mangroves also incorporates organic matter produced by the mangroves themselves. Many mangrove forests grow on sediments that are primarily organic, indicating the considerable productivity of these ecosystems. In those systems where there is only limited inorganic sediment supplied from outside the wetland, the substrate is generally peaty, comprising the fibrous root material of the mangroves. These highly organic peats characterize sediment-starved settings, such as limestone islands in the West Indies, the oceanward margin of the Everglades, and isolated islands in the Pacific Ocean (Woodroffe, 1992; Ramcharan, 2004). It is therefore essential to consider the nature of the substrate beneath mangrove forests in terms of sediment supply, distinguishing mud from outside, termed allochthonous, and from the in situ production of sediment, termed autochthonous (mangrove-derived peat). Whereas deep thixotropic muds are typical of the many turbid mangrove shorelines of the Indo-Pacific, these flexible halophytes can also grow on several meters of fibrous peat produced by ancestral mangroves. This dichotomy is another useful distinction to consider in relation to the associated wetlands that form in the hinterland of mangroves. For example, peat swamp forests of southeast Asia grow on autochthonous peats formed beneath forests of trees, in ombrogenous wetlands. Other rain-fed wetlands occur on alluvial floodplains behind subtle levees built from the sediment that seasonal tropical rivers bring from their catchments.
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4. SEA-L EVEL C ONTROLS ON W ETLAND D EVELOPMENT Tropical coastal wetlands develop where near-horizontal topography coincides with sea level, and it is clear that past sea-level changes have influenced their distribution and extent. Paleoenvironmental reconstruction provides insights into the nature of this response and may enable a clearer understanding of how such wetlands might respond to future environmental change, including sea-level rise. At a global scale, plate-tectonic setting exerts a first-order control on the distribution of coastal wetlands. Shorelines along a plate margin, such as the west coast of North and South America, are generally experiencing gradual uplift and are backed by steep mountain ranges from which short, steep rivers drain, carrying sediment into deeper waters. Such coasts provide few areas suitable for the development of wetlands, in contrast to the passive, or trailing-edge, coasts that occur on the east of the American continents. These receive large sediment loads from the long river systems that traverse the continents, such as those of the Mississippi, Orinoco, and Amazon rivers. There are more extensive low-lying areas associated with the sedimentary basins on these coasts, which are more conducive to the development of suites of landforms within which wetlands can form. Even larger sediment loads are delivered by the Indus, Ganges–Brahmaputra– Meghna, Irrawaddy, Mekong, Red, and Pearl rivers into the tropical waters of the IWP region from the tectonically active Himalayan massif. Sediment deposition has formed broad, flat delta plains in the lower reaches. Human impact on these catchments has been considerable, with land-use change often resulting in increased loss of soil from the landscape, but the construction of dams reducing delivery to the coast, in some cases to essentially zero (Syvitski et al., 2005). Steep catchments along the tectonically active island arcs, such as Indonesia and Papua New Guinea, contribute a disproportionately large sediment load to the oceans, accompanied by delta development along many of the coasts of these islands (Milliman and Syvitski, 1992). Coastal progradation has formed near-horizontal topography, providing suitable habitat for extensive wetland development. If the sea level changes, then there are ramifications throughout the wetland system. Wetlands can best form where there has been sufficient time for nearhorizontal substrates to have been deposited, such as has occurred in many parts of the tropics over the past few millennia as a result of a relatively stable period of sea level and its coincidence with near-horizontal underlying substrates. The Earth’s climate is presently in a warm period, an interglacial, and sea level has consequently regained a level occupied during previous interglacials, implying that some of the landforms are inherited from previous highstands of the sea. The peak of the last ice age occurred about 20,000 years ago, and sea level at that time was around 120 m below its present level (Lambeck and Chappell, 2001). Ice melted relatively quickly, and the postglacial sea appears to have risen to its present level by around 6,000 years ago, with only a minor amount of ice melt since that time. Those areas that were covered by ice have experienced rapid uplift since the ice load has been removed and are continuing to uplift. Areas that were marginal to the large ice sheets have also experienced isostatic land movements;
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in many cases, they are continuing to undergo subsidence as the mantle adjusts to replenish the ice-loaded regions. Most tropical coastal wetlands are sufficiently remote from former ice sheets (and tropical glaciers contributed such a minor component as to have only negligible isostatic responses), that the pattern of sea-level change experienced is similar to the global eustatic pattern, namely a postglacial sea-level rise up to 6,000 years ago but little change since. This is in contrast to temperate (or polar) latitudes, where there are ongoing isostatic movements of the land as a result of the redistribution of mass as ice has melted and ocean volume increased. Knowledge of Holocene sea-level trajectories, combined with paleoenvironmental information on mangrove forests, can be used to infer how such forests adjust to different rates of sea-level change. This approach has been most effectively developed in the case of reefs, because fossil reefs are better preserved and more accessible. Several different responses, including drowning, backstepping, catch up, keep up, prograding, and emergence, have been identified for reefs (Neumann and Macintyre, 1985). Similar types of response are shown schematically for mangrove forests in Figure 3. It can be inferred that mangroves lined shorelines when the sea was at its lowest; mangrove substrates clearly have not been able to keep pace with the most rapid rates of postglacial sea-level rise that have been experienced. If they had, then mangrove shorelines would not have shifted since the last glacial maximum (when the coastline was up to several 100 km seaward of the present one), and mangrove sediments 120 m or more thick would have been deposited across continental shelves such as the Timor Shelf (Yokoyama et al., 2001). Postglacial mangrove sediments have been identified on the broad Sahel and Sunda shelves, but these have been drowned by sea-level rise (Hanebuth et al., 2000). In shallower waters, and when rates of the sea-level rise were slightly less, it seems that mangrove forests may have backstepped. In this case, mangrove sediments evidently did not keep pace with sea-level rise but mangrove re-established at a more landward location at higher elevation. This response can be inferred for both the Belize reefs and the Great Barrier Reef where mangrove peat is encountered on the lagoon floor, and appears to have formed in an environment equivalent to, but now disconnected from, mangroves presently found on the modern mainland shoreline (Woolfe and Larcombe, 1998). It appears that mangrove sedimentation was unable to keep pace with sea-level rise in this setting. Basal transgressive mangrove sediments, generally organic-rich peaty muds at the base of cores, are found beneath coastal plains in northern Australia (Woodroffe et al., 1993). The more landward of these mangrove sediments were able to keep pace with sea-level rise as it slowed prior to reaching present level, and the mangrove forests appear to have persisted into the ‘‘big swamp’’ phase that characterized these estuarine plains (discussed below). Progradation is also recorded by the build-out of northern Australian mangrove shorelines over the past 6,000 years. In the case of sea-level fall, mangrove substrate is left emergent, as shown, for example, by the highly oxidized muds found in Malaysia (Geyh et al., 1979). Detailed evidence of relative sea-level change from many tropical locations reveals a slight fall of sea level during the past 6,000 years or even a series of oscillations (Larcombe et al., 1995). These subtle adjustments can be
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Figure 3 The response of mangrove shorelines to sea-level change. A schematic sea-level curve is shown with rapid sea-level rise, decelerating, then stabilizing before a gradual fall to present. The typical response, as indicated by the stratigraphy of mangrove sediments, is indicated.
explained either by local isostatic flexure (primarily hydro-isostatic flexure, the response of ocean basins to loading of water), or sediment-isostasy (flexural response to the loading of sediment, such as seems likely to have occurred where large loads of sediment have been deposited, as in the Asian megadeltas), or more generally by changes to the total volume of the ocean basins. There are, of course, exceptions to the general pattern of sea-level change, especially for areas that are undergoing vertical movement. For example, fossil reefs on uplifting shorelines, such as along the Indonesian arc, have been repeatedly raised by coseismic activity. In the West Indies and Florida, evidence indicates that sea level has continued to rise at a decelerating rate in those regions, reflecting the ongoing adjustment of the forebulge region in eastern North America. It seems that the West Indies have thus experienced a different sea-level history from that of other tropical areas, a fact that underpins the differences between the coastal wetlands of the New and Old World tropics.
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Mangroves are most extensive on mainland coasts where they occur in association with estuaries or deltas at river mouths. Such settings can be viewed within the classification of coastal depositional environments proposed by Boyd et al. (1992), which emphasizes the relative dominance of river, wave, and tidal processes. Deltas occur where river processes dominate, and it is possible to distinguish tidedominated and wave-dominated deltas. Estuaries can also be subdivided into wavedominated and tide-dominated. The suite of landforms on a coast differs between transgressive coasts, where the sea is rising, and prograding coasts more typical under stable sea-level conditions. Such a situation is seen on the Australian coast; Harris et al. (2002) have modified the approach based on the river–wave–tide ternary diagram and they indicate that, in the Australian context, estuaries mature into deltas. In their schema, the threshold between an estuary and a delta is identified as the point at which the initial estuarine embayment infills completely (or almost completely) and the river supplies sediment to the coast, with the consequence that the sandy coastline progrades on deltas but not on estuaries (Heap et al., 2004).
5. SEA-LEVEL C HANGE AND THE D IVERSIFICATION OF W EST INDIAN M ANGROVES The mangrove forests of Florida have been subject to a number of detailed studies (Bowman, 1917; Lugo and Snedaker, 1974). The mangroves of Florida Bay are adjacent to shallow marine carbonate banks and show the typical zonation with a seaward fringe of R. mangle and a more landward zone of Avicennia, flanking the extensive sedge- and grass-dominated freshwater wetlands of the Everglades. In an important paper, Davis (1940) considered that mangroves promoted sedimentation and had substantial land-building capability, and were building out into Florida Bay. By contrast, Egler, an ecologist, regarded the deep peats beneath the mangrove as having accreted during a period of sea-level rise and concluded that mangroves must have invaded previously freshwater environments (Egler, 1952). More detailed stratigraphic and palynological studies confirmed the latter interpretation, and mangrove peat was used to reconstruct the early sea-level curves for the region (Scholl and Stuiver, 1967). The stratigraphy typical of carbonate banks in the region is shown in Figure 4. Such stratigraphy indicates that in the late Holocene the sea was rising, as recorded by a basal mangrove peat that underlies most of the marine carbonate sediments. Such a sequence of sediments is found in Florida Bay, comprising an intertidal peat overlain by shallow-water carbonate and is transgressive, recording a relative rise of sea level (Enos and Perkins, 1979). On the other hand, mangroves are observed to now be spreading over marine carbonate banks, as observed by Davis. This is a regressive stratigraphy, recording the progradation of mangroves (Parkinson, 1989). A similar stratigraphy is found elsewhere in the region, for example, beneath the extensive mangrove forests of the Cayman Islands (Woodroffe, 1981) and Jamaica (Hendry and Digerfeldt, 1989). Such a pattern of decelerating sea-level rise is
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4 ka
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Figure 4 Typical sea-level curve for the Florida and West Indian region, and the schematic response of mangrove peat and carbonate environments as recorded in the stratigraphy of Florida Bay and several islands of theWest Indies.
observed widely throughout the Caribbean (Toscano and Macintrye, 2003) and represents the gradual response of this region to the rebound from ice melt of the North American ice sheet. However, a slightly different sea-level history in inferred
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for Belize (Gischler, 2006), and in South America it appears from the stratigraphy of mangrove forests that sea level reached its present level considerably earlier (Cohen et al., 2005), as it has in the IWP.
6. S EA -LEVEL C HANGE AND THE E VOLUTION OF MANGROVE H ABITATS IN THE IWP It is becoming clearer that during glacial periods, much of the Sunda Shelf experienced greater seasonality; wide, multichanneled, or braided sand and gravelbedded rivers carried large quantities of coarse sediment to shorelines on the exposed shelf. More subtle changes of climate are recorded by interbedded sands and freshwater peat horizons, such as those at Pantai Remis, south of Ipoh (Kamaludin et al., 1993). As sea level rose, marine sediments were deposited (Verstappen, 1975), this being recorded in numerous pollen sequences (Woodroffe, 1993). The stratigraphy and pattern of habitat change in response to the general sea-level history of the southeast Asian and northern Australian region is shown in Figure 5. Transgressive sediments record the rapid rise of sea level across much of the region. Mangrove deposits on the Sunda Shelf indicate that a fringe of mangrove forests must have traversed much of the shelf as it was flooded following ice melt (Hanebuth et al., 2000). The Holocene evolution of a number of systems in northern Australia has been interpreted, primarily on the basis of stratigraphic and radiocarbon-dating analyses, with particular emphasis on depositional environments but also including the evolution of channel form. Sea-level rise drowned the valleys, resulting in the landward transgression of mangroves over terrestrial environments and the formation of broad, open embayments (Figure 6). Initially, sediment supply could not keep pace with the rapid creation of accommodation space, and the catch-up transgressive sequence is recorded as a layer of organic-rich mangrove mud overlying pre-Holocene terrestrial sediments (Woodroffe et al., 1993). Sea level stabilized in northern Australia around 7,000–6,000 years BP, and many estuarine systems are filled by 10–15 m of ‘‘big swamp’’ mangrove muds that were able to keep up with the final decelerating rise in sea level. This sedimentary sequence has been described from the South Alligator system in particular (Woodroffe et al., 1985, 1989), but has also been shown to occur in neighboring systems (Chappell, 1993; Woodroffe et al., 1993), and can be inferred for the Fitzroy and Ord rivers in western Australia as well as more generally along the north Australian coast (Jennings, 1975; Thom et al., 1975; Chappell and Thom, 1986; Crowley, 1996). The nature of the sedimentary infill that has been deposited following sea-level stabilization varies, depending on the availability of material and the ability of the local environment to transport it, but typically it consists of either sands transported by wave or fluviotidal processes or muds associated with lower energy mangrove environments (Semeniuk, 1985). Within the estuaries themselves, vertical accretion has continued, but pollen analysis indicates a transition from mangrove to alluvial
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Figure 5 Typical sea-level curve for the Southeast Asia and northern Australian region, and schematic response of mangrove mud and adjacent environments. The slight emergence is a result of late Holocene relative sea-level fall; such a pattern is seen in Malaysia.
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Early Holocene
Mid Holocene
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Arid
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Figure 6 Three Holocene stages of estuarine evolution that occurred throughout much of northern Australia (based on Woodroffe et al., 1989, 1993) but that can be inferred to have typified much of southeast Asia (Woodroffe, 1993). The morphology of the plains that have developed during the late Holocene depends on climate as shown by the schematic cross sections. In arid areas, a saline flat remains over mangrove sediments; in the wet ^ dry tropics (as in the top end of Australia), seasonally inundated sedge-grassland dominates convex floodplains, whereas in the perhumid tropics, such as in Malaysia and Indonesia, domed peat occurs covered by peat swamp forest.
sediments as the floodplain has matured and mangroves have been replaced by grass and sedgeland vegetation, as observed there today. In many cases, mangrove had been replaced by 4,000 years BP although in other places it persisted until more recently (Clark and Guppy, 1988). Simultaneously with the transition to freshwater plains, progradation at the coast has built out broad coastal plains. Progradation appears to have been gradual but at a slowing rate, except when episodic high wave-energy events have created chenier ridges or, less frequently, beach ridges (Woodroffe and Grime, 1999). Despite coeval transgressive–regressive stratigraphy, modern environments along the coast of northern Australia show considerable variability. Mangroves are generally of limited extent in the more arid environments of western Australia
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(Semeniuk, 1996; Lessa and Masselink, 2006). High salinity restricts mangrove colonization and the tidal mudflats that form accrete into salt flats. In those areas that receive sufficient rainfall, floodplains may continue to accrete, with levees developing that restrict tidal waters to the major channels during the dry season but impound freshwater on the floodplains during the wet season (Woodroffe et al., 1993). Subtle variations in topography across the plains (associated with lowelevation cheniers, scroll bars, and levees) can be very important in determining this balance; for example, the Mary River comprises a series of plains from which the wet season flood waters would have evaporated, without a fluvial connection to the sea until recent expansion of tidal systems reconnected the river to the sea (Mulrennan and Woodroffe, 1998a). Fluvial and tidal channel deposits associated with channel migration can also become abundant at these elevations if rates of channel migration are high enough and the supply of sediment is adequate, as shown by a series of deltas flanking the Gulf of Carpentaria, formed over this late Holocene period (Jones et al., 2003). Tropical forests have replaced mangroves in the wettest regions of Queensland (Crowley and Gagan, 1995). Geomorphological features constrain mangrove assemblages in tidal creek settings in northern Australia. Six major geomorphologically defined habitat types have been distinguished in tidal creek embayments by Semeniuk (1985), who identified relatively distinct mangrove assemblages within each although the diversity is ultimately restricted by the regional species pool. The influence of habitat type results from the different soil and groundwater regimes associated with each habitat, as these factors exert a strong effect on mangrove competition, and environmental gradients are reflected in corresponding species gradients (Semeniuk, 1985). The evolution of these landforms entails a change in the distribution of each habitat that will depend on the relative influence of different geomorphological processes. For example, alluvial fans are more extensive in those tidal embayments with greater fluvial inputs. Similarly, spits and cheniers are built more frequently in systems exposed to higher wave energies. Such patterns are likely to be associated with a corresponding shift in the distribution of mangrove assemblages, and the mangrove types that occur in these complex mangrove forests are more diverse than those summarized in the West Indian situation in Figure 1. The tidal river systems of northern Australia began as open embayments drowned by the rapidly rising postglacial sea level. Following sea-level stabilization, the nature of sedimentation appears to have varied with distance from the coast. Seaward regions of the embayments filled with a mixture of shelly fine sand and mud presumably delivered from offshore as a result of pumping by tidal currents (Chappell and Woodroffe, 1994; Wolanski, 2006a, 2006b). Around 6,000 years ago, many of the embayments contained substantial intertidal mangrove forests, interspersed with tidal channels (Figure 6). These environments infilled rapidly with fine-grained sediments derived from offshore (cf. sedimentation rate far exceeded the estimated rate of fluvial sediment delivery; Chappell and Woodroffe, 1994). As sediments accreted, and as indicated by the pollen record, lower intertidal species (Sonneratia) were replaced by mid-intertidal species (Rhizophora/Ceriops), in turn replaced by higher intertidal species (Avicennia) that were subsequently replaced by freshwater floodplain species (Woodroffe et al., 1986). In the case of
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the South Alligator and Daly rivers, the infilling of the mangrove swamps was also associated with the onset of channel migration throughout the estuarine and coastal plains, resulting in reworking of big-swamp sediments and lateral accretion of laminated channel margin and shoal sediments (Chappell, 1993; Woodroffe et al., 1993; Figure 2f). In smaller systems, the banks near the mouth may be reworked by waves, but fluvial discharges appear less efficient at eroding the banks. In southeast Asia, a similar pattern of stratigraphy and chronology has been recorded (Woodroffe, 1993; Staub and Gastaldo, 2003; Hope, 2005). Early studies of extensive peat swamp forests in the region had shown that these were formed on domed peat deposits in the wettest settings (Figure 6) and that they had formed over mangrove muds (Anderson, 1964). Pollen analysis has also revealed the transition from mangrove to peat swamp forest, and transition through successional stages of that forest as the peat has built up beneath the trees (Anderson and Muller, 1975). The Anderson model of peat swamp development is widely applicable in southeast Asia (Hope et al., 2004) although Taylor et al. (2001) indicate that a similar peatbased lowland tropical swamp can also develop in a topogenous setting, as it has done at Nee Soon in Singapore.
7. IMPACT OF FUTURE C LIMATE AND SEA-LEVEL C HANGE It is clear that mangrove forests have undergone, and survived, major geomorphological transformations during the Holocene, primarily as a consequence of global sea-level rise. It seems highly likely that the subtlety of past response to sealevel change may hold important lessons as to how these systems will respond to future climate and sea-level change. Whereas temperature and carbon dioxide concentrations may have some effects on mangrove growth, this is already highly spatially variable and these climate drivers seem unlikely to have major impacts on ecosystem functioning (Saenger, 2002). Changes in storm intensity may affect mangrove distribution, as may a suite of other factors (Field, 1995), but it seems clear that one of the most major impacts is likely to be from any acceleration of sealevel rise (Semeniuk, 1994). The stratigraphic record indicates that mangrove systems have survived rapid rates of sea-level rise in the past, and it provides a rich archive of information on the nature of the response (Woodroffe, 1990). At the ecosystem level, it is clear that mangroves do not face total extinction as a result of sea-level rise. Stratigraphy may offer insights into the rates at which mangroves can keep up, as opposed to catch up, with sea-level rise. Ellison and Stoddart (1991) considered that collapse of mangrove ecosystems might be imminent on small islands where there is not a large supply of inorganic sediment, but their view is not shared by other researchers familiar with these environments (Bacon, 1994; Snedaker et al., 1994). The paleoenvironmental record could potentially yield insight into the critical rate of sea-level rise beyond which mangrove systems change from ‘‘keep up’’ to ‘‘catch up’’, indicating a threshold beyond which they might lag future sea-level rise. It is important, however, to note that the distinction is primarily about
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whether the seaward fringe has kept pace. Stratigraphy confirms that the most fundamental response of mangroves to sea-level rise is incursion into more landward habitats. It will not always be the case that the space within which this can happen will be available in future. This issue of coastal squeeze, where landward areas have been embanked or reclaimed for alternative uses, has been identified previously (Woodroffe, 1990; Delafontaine et al., 2000; Gilman et al., 2006). Perhaps less well investigated is the question of whether there is hinterland to invade. The presence of broad prograded plains, such as those developed on many coasts in the IWP region, imply that rapid sea-level rise in these areas would lead to inundation of these plains. There has not been the detail of stratigraphic reconstruction to demonstrate whether such episodes have recurred in the past. It is clear that increased tidal processes will widen tidal channels (Wolanski and Chappell, 1996), but the intertidal systems can be anticipated to have only a limited ability to continue to build up and to accentuate the levees before sea-level rise exceeds this capacity and widespread inundation of the plains occurs (Woodroffe, 2007). This has been likened to re-establishment of the big swamp. In the case of salt marshes, the stratigraphy of coastal plains around much of Europe and eastern North America contains interbedded marine muds and freshwater peats that are interpreted to record alternation of freshwater wetlands and more marine environments in the past (Allen, 2000). Managed realignment is already practiced in the United Kingdom, such reversion being promoted as a management strategy in the face of sea-level rise (Wolters et al., 2005). It is important to recognize that mangrove shorelines are highly sensitive to changes in sea level, and modifications in the distribution of mangroves may be one sign of sea-level rise (Blasco et al., 1996). On the other hand, it is evident that mangroves are naturally dynamic and that substantial changes in mangrove distribution are already occurring as a part of the geomorphological behavior of these systems (Lucas et al., 2001). Mangrove incursion into more landward salt-marsh habitats, and loss of salt marsh as a result, has been observed along a series of estuarine systems in southern New South Wales (Saintilan and Williams, 1999), but this does not appear to be related specifically to sea-level change and may be a more complex response to other anthropogenic pressures such as landuse change (Harty, 2004). Widespread inundation of plains might occur under future rapid sea-level rise, but more insidious might be saline incursion into freshwater wetlands through an expanding network of tidal creeks. Tidal creek systems are a component of the natural dynamics of these complex coastal systems, but increased salinization through creek extension is also occurring on several systems already. Freshwater wetlands are separated from tidal waters by grasslands, low chenier ridges, and channel levees that reach elevations only a few centimeters above high tide level, acting as barriers to saltwater intrusion, experiencing occasional inundation by storm surges and by seasonal freshwater floods (Woodroffe et al., 1993; Winn et al., 2006). In recent decades, erosion of mangrove habitats and the growth of tidal channels have been recorded in several parts of northern Australia, including the Pilbura (Semeniuk, 1980, 1983) and Alligator River regions (Winn et al., 2006). However,
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saline intrusion has occurred most dramatically on the Mary River plains (Knighton et al., 1991, 1992; Mulrennan and Woodroffe, 1998b). As the tidal channels extend and deepen, they create local increases in the tidal range and transfer saltwater to regions previously protected from tidal influence. This has resulted in the salinization of large areas of freshwater wetland, death of Melaleuca, and upstream extension of mangrove species. The growth of channel networks has undergone exponential growth in terms of number and length although the causes continue to be debated (Mulrennan and Woodroffe, 1998b). Whereas it has not been demonstrated unambiguously that tidal creek extension is related to sea-level change, these examples do demonstrate the speed with which saline intrusion can proceed by this mechanism, which is likely to become much more widespread if sea-level rise accelerates. Low-lying plains, such as those developed in much of the IWP region during the past few millennia, are likely to be subject to subsidence and compaction. This is accelerated in some regions through extraction of groundwater, which, for example, has been shown to have led to rapid subsidence beneath Bangkok (Phienwej and Nutalaya, 2005). Groundwater can influence other aspects of the behavior of wetlands (Perillo et al., 2005; Gallardo and Marui, 2006). Whether similar processes are involved in vegetation change on the relatively unpopulated plains of northern Australia is unclear although direct evidence for this is weak. Nevertheless, detailed measurements of ground surface elevation in mangroves not only demonstrate considerable spatial variation in rates of sedimentation, with the most rapid potential rates accentuated by the presence of prop roots (Krauss et al., 2003), but it also appears that surface elevation changes are only partially a record of sediment deposition. Using controlled sedimentation tables, substrate elevation has been shown to also incorporate much more subtle changes of ground surface in response to ground water changes within the substrate (Rogers et al., 2006). These complexities imply that the response of mangrove shorelines to sea-level change will differ between systems (Semeniuk, 1994; Woodroffe, 1995). There may be other subtle changes as well; for example, gap regeneration appears to be an important process in the cycle of mangrove tree replacement (Duke, 2001) although to differing extents in terms of which trees remain standing (Pinzo´n et al., 2003). These and other processes may adjust in response to other aspects of climate change, such as changes in storm frequency and intensity.
8. S UMMARY AND C ONCLUDING R EMARKS Mangroves have often been interpreted as ‘‘trees that reclaim land from the sea’’, in view of their vivipary and unique aerial root systems (Carlton, 1974). This builds on ideas of Davis (1940) that mangrove peats in Florida overlie marine carbonate sediments. In fact, stratigraphic studies have indicated that in many cases, these calcareous marls were laid down in more freshwater environments, as now found in the Everglades, and within the stratigraphy there is a complex pattern of initial transgression followed by subsequent regression (Parkinson, 1989).
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Mangrove forests are some of the most conspicuously zoned of plant communities (Watson, 1928; Chapman, 1976), and this has often been interpreted in terms of a succession of species that might subsequently become replaced by a more landward, climax plant community. Whereas pollen studies do provide evidence of such gradual replacement; for example, throughout much of the southern Asia and northern Australia region, re-evaluation of areas in which these ideas were initially applied shows little apparent change over the past half a century (Alleng, 1998). In geomorphologically dynamic areas such as the Niger delta or the delta of the Grijalva in Mexico, mangroves appear to opportunistically colonize habitats (Allen, 1965; Thom, 1967). However, even where sediment supply might appear ample, mangrove shorelines can demonstrate stability and a morphological resilience, as in Sherbro Bay in West Africa, implying that they are in equilibrium with the hydrodynamics and sediment transport through the system (Anthony, 2004). Geomorphologically, mangrove shorelines have adapted to past patterns of sealevel change, some of which have occurred more rapidly than at rates occurring at present or anticipated in the immediate future. Mangrove systems are adjusted to substantial sediment loads, and sedimentation rates can be high and often spatially variable. Climate change threatens many ecosystems with conditions to which they are not adapted. Mangrove forests, by comparison, are relatively well adjusted to extreme environmental conditions, whether in terms of salinity tolerance or inundation by salt water. Sea-level rise, at the rates presently experienced or at accelerated rates being predicted, should not represent undue stress for mangrove forests. However, geomorphological adjustments are inevitable. Mangroves are generally well able to adapt to changing conditions, and the interactions with adjacent systems and the landward migration of mangroves may seem unexpected or undesirable. Under the most extreme conditions presently foreseeable, such as melting of the West Antarctic ice shelf, more rapid sea-level rise may challenge mangrove systems and trigger more rapid changes than currently seen, but these seem unlikely to pose greater threats than are already caused by man. These anthropogenic pressures, combined with additional climate change, seem certain to lead to continued decline of mangrove ecosystems, unless the economic and ecological values of these systems are more broadly appreciated, and opportunities to preserve and extend their distribution by conservation and planting are undertaken with greater urgency.
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C H A P T E R
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T EMPERATE C OASTAL W ETLANDS : M ORPHOLOGY , S EDIMENT P ROCESSES , AND P LANT C OMMUNITIES Paula D. Pratolongo, Jason R. Kirby, Andrew Plater, and Mark M. Brinson
Contents 1. Introduction 2. Factors Controlling Sediment Dynamics 2.1. The “ramp” model of salt marsh accretion 2.2. The “creek” model of salt marsh accretion 2.3. Storms and salt marsh erosion 3. Factors Controlling Patterns of Vegetation 3.1. Zonation of vegetation 3.2. Ecological development 4. Geographic Variation 4.1. Northern Europe 4.2. Eastern North America 4.3. Western North America 4.4. Mediterranean 4.5. Eastern Asia 4.6. Australasia 4.7. South America 5. Human Impact and Climate Change 5.1. Human impact 5.2. Climate and sea-level change 6. Summary References
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1. INTRODUCTION In addition to their ecological, cultural, and aesthetic significance, temperate coastal wetlands play an important role in determining the sustainability of coastal biological populations through their dynamic interaction with physical and chemical properties of the environment. The interaction of biota, hydrology, and Coastal Wetlands: An Integrated Ecosystem Approach
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sediments is clearly evident in the ecological and geomorphologic characteristics of temperate coastal wetlands. They offer perhaps the clearest illustration of the extent to which coastal habitats respond to abiotic factors such as tidal inundation, sea level, climate, groundwater, accommodation space (the space between sediment surface and tidal level), sediment supply, water quality, and water and sediment dynamics, as well as human and other biotic activities. In addition, their paleoenvironmental records provide both ecological and chronological information on their evolution in response to many of the same forcing factors and reveal a variety of data concerning past climate change, vegetation history and paleohydrology, sealevel trends, and alteration by human activities (Tooley, 1986). In their broadest sense, temperate coastal wetlands include a wide spectrum of environments from tidal flats to fen woodland communities or barren salt flats, in a continuum that can span several kilometers (Figure 1). While tidal flats at the seaward end are similar in terms of the physical constrains precluding vegetation establishment, plant communities at the landward margin are subjected to climatic and groundwater influences. For fen wetlands, sea level is an important control on the groundwater position that provides the waterlogged conditions necessary for their development. Hageman (1969) termed the area where freshwater wetlands persist under the control of relative sea-level movements as the “perimarine zone” and the seawards wetlands where marine and brackish sediments are directly deposited as the “tidal flat and lagoonal zone.” In arid climates where freshwater inputs are scarce, the perimarine zone would extend through salt flats where the low soil water potential eliminates all but the most tolerant halophytes or excludes macrophytes altogether. This chapter encompasses general aspects of wetlands along this continuum, including wetland environments below the highest astronomical tide which experience direct tidal inundation, with decreasing frequency and duration as a function of increasing elevation within the tidal frame, and those wetlands landward where
Figure 1 Temperate coastal wetland continuum (after Waller, 1994).
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the water table is linked to the sea-level influence. For simplicity, these are referred to in this chapter as “intertidal” and “perimarine” wetlands, respectively. However, more detailed descriptions on the geographical variations will focus on intertidal wetlands, particularly on salt marshes, given that tidal freshwater wetlands will be covered thoroughly in a different chapter of this book. In the following sections, we aim to illustrate the main geomorphic and sedimentary development processes, factors that control zonation of vegetation (including disturbance and successional change), their geographic variation globally, and major modification by human activities. The section on geographic variation will compare these wetlands from different localities within the temperate zone. The impacts of climate change are briefly discussed and future sustainable management techniques are outlined, which demonstrate how coastal wetlands can offer humanity a means of environmental protection and improved quality of life.
2. FACTORS C ONTROLLING SEDIMENT D YNAMICS The development of coastal wetlands of differing character is related to a variety of biotic and abiotic factors. In the case of the intertidal zone, the primary abiotic control on wetland structure and function is a combination of tidal inundation frequency, depth, and duration, known as “hydroperiod” (French, 1993) or “hydropattern” (Weinstein and Kreeger, 2000; Jackson, 2006). With reference to reviews of intertidal wetlands (Dijkema, 1987; Pye and French, 1993), a further distinction can be made between macrotidal and intertidal wetlands formed primarily through the accumulation of externally derived mineral matter, and autochthonous wetlands found in more microtidal settings where plant biomass accumulation maintains elevation within the tidal frame (French, 2006). In the perimarine zone, water level and water quality combine to determine wetland properties. In humid regions, wetlands in close proximity to the high tide position typically show a high degree of waterlogging and organic matter accumulation. In dry climates, on the contrary, the combination of scarce freshwater availability and high evaporation rates often determines extremely elevated levels of surface and groundwater salinity in the supratidal zone, limiting plant growth and precluding organic accumulation in soils. Although supratidal salt flats occur to a greater extent along tropical coastlines, their formation is also common in temperate regions under extremely dry conditions. While salt marsh vegetation plays an important role in determining the rate of sediment accretion and stabilization (Brown, 1998; Reed et al., 1999; Cahoon et al., 2000), the distribution of species reflects the overarching control of elevation within the tidal frame (Figure 2). If salt marshes and the contiguous tidal flats are conceptualized as a ramp that increases in elevation landward, there is a resulting decrease in both the frequency and duration of tidal inundation as elevation of the surface approaches that of the highest astronomical tide. The elevation control on hydroperiod creates a gradient to which plants and animals respond in terms of their distribution (Gray, 1992; Gardner et al., 2002).
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Figure 2 Species distribution in salt marshes of northwestern Europe and North America as a function of hydroperiod (after Carter, 1988).
2.1. The “ramp” model of salt marsh accretion The “ramp” model of salt marsh sedimentation considers the rate of salt marsh accretion as a direct function of surface elevation and, hence, hydroperiod. Here, hydroperiod and the availability sediments determine the amount of particles settling on the salt marsh surface; sites that are lower in the tidal frame will, therefore, experience more sediment deposition on each tide. Suspended sediment concentration itself is determined by tidal velocity, rainfall, and biological activity (Murphy and Voulgaris, 2006). As salt marsh elevation increases through sediment accretion, hydroperiod and net sediment accretion are reduced. As environmental stress induced by hydroperiod decreases with elevation and the influence of gravitational drainage of the freshwater table increases, the uppermost part of the intertidal zone switches from incremental minerogenic sediment accretion to the accumulation of organic matter in situ. Hence, the net balance between biological productivity and decomposition governs sedimentation on the upper salt marsh, particularly where the freshwater influence leads to the establishment of species less tolerant of salinity and where aerobic decomposition may become important. This balance between sedimentation through the accumulation of both mineral and organic matter enables continued vertical accretion at the intertidal/perimarine interface. There is, however, a great deal of spatial variation in the extent to which salt marshes are characterized by organogenic sediment accumulation, mainly as a function of tidal range, climate, and salt marsh ecology. Autocompaction plays a role in mediating vertical accretion rates, especially for relatively thick organic-rich deposits (Allen, 1999). This is less likely as a control for intertidal salt marshes where sediment supply is largely allochthonous minerogenic material (Cahoon et al., 2000) (Figure 3).
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ESLR and deep subsidence Mineral matter
Relative wetland elevation
Temp. Above and bellowground primary production
Decomposition
Sediment Cohort 1 Labile organic matter Refractory organic matter Live roots Mineral matter Pore space
Compaction
Sediment cohort 2 Sediment cohort n
Factors raising marsh elevation Allogenic factors Autogenic factors Other factors
Mineral matter Detrital organic matter Labile organic matter Refractory organic matter Live roots Pore space Water content
Factors lowering marsh elevation Autocompaction Decomposition Sea-level rise Subsidence
Figure 3 Factors contributing to marsh surface elevation relative to sea level (Allen, 1990; Rybczyk et al., 1998).
Feedbacks by vegetation exert control over the simplistic, conceptual ramp model in a number of ways as following: 1. Increased bed roughness results from the presence of leaves and stems, thus increasing frictional drag and near-bed turbulence and enhancing sediment deposition potential (Leonard and Luther, 1995; Leonard and Croft, 2006). 2. Increasing biomass also increases the surface area available for sediment capture, for example, leaf surfaces (Stumpf, 1983). 3. The presence of roots binds and further protects sediment once deposited (Garofalo, 1980). A protective binding function is also provided by filamentous algae that cover the mudflat surface, thus providing a stable surface on which halophytes may become established (Coles, 1979; Underwood and Paterson, 1993). Important exceptions to the ramp terminating at an upland margin or in a perimarine wetland are found in the flood tide deltaic marshes located behind barrier islands and in the outer estuary fringing marshes.
2.2. The “creek” model of salt marsh accretion Micromorphology is further reflected in the “creek” model of salt marsh sedimentation whereby the network of drainage channels (creeks or gullies) that crosses the
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salt marsh surface acts to capture and funnel the incoming tide. Here, the rising tide spills onto much of the salt marsh surface only when it exceeds the capacity of the creeks. This results in an observed increase in sediment grain size with proximity to the creek margin as a function of the competence of the overbank flow decreasing with distance from the creek margin (Christiansen et al., 2000; Leonard et al., 2002; Temmerman et al., 2004). With perhaps the exception of fringing shoreline salt marshes found in the outer estuary, levees generally develop on creek banks where more of the coarser sediment load is deposited, much as they do in the development of levees on river floodplains. There is clear evidence that supports the viability of both the ramp and the creek models for salt marsh accretion. However, they operate at very different time and space scales; while the ramp model accounts for widespread and gradual trends in vertical accretion, the creek model is focused more on the local scale and potentially rapid development of three-dimensional sedimentary features. The role of estuarine hydrodynamics and mixing on salt marsh geomorphology also has to be considered in the supply and deposition of fine sediment. In a simple model of salt marsh sedimentation (Orme, 1990), deposition of fine sediments with low settling velocities is restricted to the slack water periods at or near high tide. The velocity profile of the incoming and outgoing tidal waves shows a roughly Gaussian distribution with time, and the highest flow velocities occur at mid-tide, decreasing to zero during slack water (Allen, 2000). The incoming flood tide, therefore, has decreasing competence for sediment transport as it approaches high tide over the intertidal zone, with progressively finer grains settling from suspension. Conversely, sufficient velocities for sediment entrainment on the outgoing tide are only attained after the ebb has propagated for some time. Other studies have demonstrated a continuous settling of suspended sediments of different sizes delivered by flood tide velocity pulses in adjacent channels (Bayliss-Smith et al., 1979). Reed et al. (1999), in a macrotidal marsh in the coast of North Norfolk, England, found higher concentrations of suspended sediments over the marsh surface related to “flood velocity pulses” in the creek (French and Stoddart, 1992). These findings indicate that flows in tidal channels play a critical role in providing sediments for deposition on the marsh surface. The mixing of marine and freshwaters in estuaries further enhances the deposition of fine grain sizes due to their cohesive nature through a combination of ionic bonding and organic grain coatings (Alldredge and Silver, 1988). Hence, estuarine mixing is associated with flocculation of fine grains into aggregates, which then settle as coarser grain sizes (Krone, 1978). Furthermore, the dynamics of estuarine circulation determine where this “turbidity maximum” occurs and where mud deposition is more likely (Manning et al., 2006).
2.3. Storms and salt marsh erosion In addition to the overarching influence of the antecedent conditions and sea-level fluctuations, vegetation also responds to shorter term events such as storms in several ways. van Proosdij et al. (2006) proposed that in the Bay of Fundy, wave activity increases suspended sediment concentrations and the transport of suspended
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sediment further up onto the mid-marsh to upper marsh. A similar process was also found in the Villa del Mar salt marsh (Bahı´a Blanca Estuary) during storm activity and fine sediments were transported as fluid mud. In contrast, Pethick (1992) found that in storms in Essex, United Kingdom, where waves are effectively propagated into the estuary or perhaps where locally produced waves reach sufficient height, the result may be erosion of the outer and lower parts of the salt marsh.
3. FACTORS C ONTROLLING P ATTERNS OF V EGETATION 3.1. Zonation of vegetation The lower limit of salt marsh is unambiguously defined as the seaward margin of emergent vascular plants (Adam, 1990). These areas are regularly inundated by salt water and usually consist of hardy pioneer genera such as Salicornia, Suaeda, Aster, and Spartina (Doody, 1992). Above this in the mid-level salt marsh where hydroperiod is less, a greater diversity of plants can colonize. This floristic variation is illustrated by the identification of 28 different communities of salt marsh vegetation in the United Kingdom alone, as described by Rodwell (2000) and expanded on in Section 3.4. The salt marshes are, therefore, dominated by halophytic plants although salinity is not the dominant control on plant establishment at the seaward margin. In low marsh areas, the salinity is comparatively low due to regular tidal flushing that prevents the accumulation of salt. Instead, substrate stability, oxygenation, and sulfide toxicity are key factors controlling plant establishment at the seaward margin (Adam, 1990; Mendelssohn and Morris, 2000). Salinity is higher in mid-marsh areas where there are prolonged periods between tidal inundations when evaporation takes place, which can lead to the buildup of high concentrations of salt, most notably in arid climates. Similarly, disturbance (wrack deposition) and biotic interactions (Bertness, 1991) can greatly influence plant establishment in the high marsh. From early descriptions by Yapp et al. (1917), salt pans are patches of bare soil that develop in shallow depressions particularly in the mid-salt marsh to upper salt marsh. Although very different mechanisms of formation may operate in different marshes (Pethick, 1974; Perillo et al., 1996; Adam, 1997), salt pans at some locations may be related to the storage of saltwater in shallow depressions. Here, the enhanced salinity induced by evaporation and incomplete soil-water flushing leads to dieback and lack of reestablishment (Hayden et al., 1995). The result of the interaction between hydrodynamics, elevation, and vegetation is to produce a shore-parallel zonation of plants which, in reality, is made more complex and spatially variable by the micromorphology of the marsh surface. For example, enhanced sediment accretion has been observed on hummocks covered by Puccinellia maritima (Langlois et al., 2003). These local increases in sedimentation may be a function of vegetation height and stem density (Boorman et al., 1998). Micromorphology also influences drainage and potential erosion of the marsh. In humid climates with sufficient freshwater inputs, there would be a gradational transition from the upper salt marsh limit to a freshwater reed swamp, which
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creates a clear zonation of plant species at different elevation levels. This transition is often restricted, however, due to anthropogenic activities such as drainage, reclamation (filling a marsh to raise elevation), and emplacement of coastal defenses. Beyond these freshwater communities, tall herb fen wetlands would develop, where waterlogging is sustained by freshwater seepage, freshwater flooding, and high groundwater levels due to proximity to the coast. Here, depending on the amount and nutrient status of the receiving water, an array of floristically different wetland vegetation communities may develop (Rodwell, 1995), including freshwater marshes and swamps described in Whigham et al. (2009). In arid climates, however, salinities can exceed the limits of even the most tolerant halophytes, and salt flats devoid of vascular vegetation develop near the upland boundary. Other biota also occupy distinct zones on coastal wetlands (diatoms and foraminifera) that reflect the relative ecological tolerance of different species to changing environmental conditions (Scott and Medioli, 1978; Zong and Horton, 1998). This phenomenon has been used in semiquantitative reconstructions of past coastal environments (Vos and de Wolf, 1993) and, more recently, quantitative reconstructions of relative sea level (Horton, 1999; Gehrels, 2000; Edwards and Horton, 2000, 2006; Horton and Edwards, 2006). Statistical investigations demonstrate that the dominant control on the distribution of microorganisms on salt marshes is flooding duration as a function of elevation (Gehrels, 2000).
3.2. Ecological development The ecological development of temperate coastal wetlands is a product of both autogenic and allogenic factors (Walker, 1970; Jackson et al., 1988; Singer et al., 1996; Waller et al., 1999). These factors determine spatial patterns in the wetland vegetation communities and also their development over time. Autogenic factors are usually biotic processes that operate naturally within an ecosystem to dictate the direction and nature of changes in plant succession until the vegetation reaches “equilibrium” with the environment (Schofield and Bunting, 2005). However, the patterns and pathways of successional change within coastal marshes are largely externally controlled by the operation of allogenic factors such as sea-level change and storms. This is most clearly expressed by the reconstruction of coastal plant communities from biostratigraphic profiles using pollen analysis (Godwin, 1940; Shennan, 1986; Waller, 1994). Some of the patterns that emerge include sedimentation of allochthonous material (both organic and inorganic) in response to relative sea-level change, trends in groundwater level and quality, and disturbances from storms and flooding, climate change, subsidence, and vegetation disturbance due to human impact, which reflect the interaction of coastal, terrestrial, and anthropogenic processes (Clark and Patterson, 1985; Clark, 1986). As such, ecological succession, principally controlled by autogenic processes by definition, gives way to state change, principally controlled by changes that result firstly in altered hydrology and salinity and secondly by the vegetation (Walker, 1970; Brinson et al., 1995). In the perimarine zone of humid regions, the most important allogenic factor controlling groundwater level, and thus sedimentation and vegetation
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development, is base level or sea-level change. In these temperate humid climates, it is clear that the principal factor initiating peat formation over formerly dry basement surfaces was raising water levels during the early stages of the Holocene (Jelgersma, 1961; Hageman, 1969; van de Plassche, 1982; Pons, 1992; Vos and van Heeringen, 1997). In arid climates, in contrast, a combination of high rates of evapotranspiration and low freshwater inputs precludes peat formation in the perimarine zone. Although sea level may be the long-term regional control on water table height, species composition and abundance of vegetation can respond to interannual and seasonal variations in rainfall (Callaway and Sabraw, 1994; Noe and Zedler, 2001). These large but short-term variations in the controlling factors make the effects of sea-level rise more difficult to evaluate.
4. G EOGRAPHIC VARIATION Given that all tidal marshes share a similar range of geomorphologic settings and hydrological driving forces, they would be comparable in terms of ecosystem structure and function (Brinson, 1993). However, even within this variation substantial differences arise among major geographic regions due firstly to precipitation and secondly to the biogeographic distribution of species. In addition, human activities have also exerted profound changes in coastal wetlands. Worldwide, half of wetlands are estimated to have been converted to other uses during the past century (Dugan, 1993), and given the concentration of human settlements near shorelines, a large number of coastal wetlands through the world have disappeared or radically changed in area and functions (Holmer, 2009). However, the processes of wetland loss and degradation have been quite variable in space and time and have led to major regional differences in the extent and ecological integrity of remaining wetland areas. Introduction of nonnative species that rapidly expand into native communities has contributed to this trend and, in some cases, may have far-reaching consequences on both biotic and physical processes. The geographical grouping used here is based on the work of Chapman (1960). In his early attempt to classify coastal systems, salt marshes of the world fall into distinct groups characterized by different types of vegetation. In this updated review, we also incorporate major regional geomorphic settings that affect the distribution of vegetation.
4.1. Northern Europe Temperate tidal marshes in Northern Europe zone occur along the Atlantic and Channel coasts from the Iberian Peninsula northward to Denmark, through the barrier complex of the Friesian Islands and the Wadden Sea, and including the southern coasts of Great Britain. Some of the better known salt marshes of Europe are found at the Mont Saint-Michel Bay in the south of the Gulf of Normandy. Since the 11th century, 133 km2 of salt marshes have been converted to agriculture. Nowadays, the salt marshes extend over 40 km2 with most area under sheep grazing (Lefeuvre et al., 2000).
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Typical vegetation pattern includes a pioneer zone with a sparse cover of Spartina anglica and Salicornia dolichostachya/fragilis. Landward, P. maritima appears as small clones, and a middle marsh establishes with higher vegetation cover and diversity. Aster tripolium, Salicornia ramosissima, Suaeda maritima, Halimione portulacoides, and Atriplex portulacoides may also appear in this zone (Bouchard et al., 1995; Langlois et al., 2003), while Festuca rubra, Juncus gerardii, and Elymus athericus are described as common dominants in the upper marsh (Figure 2). Recently, E. athericus, a native clonal grass, has been aggressively invading the middle and low marshes, where it often forms dense monospecific stands replacing the natural A. portulacoides marshes (Bouchard and Lefeuvre, 2003). This pattern of colonization has been reported elsewhere in northern Europe (Bockelmann and Neuhaus, 1999; Vale´ry et al., 2003), and an increase of edaphic nitrogen is thought to be facilitating E. athericus invasions. Well-studied tidal marshes also occur in the Severn estuary, in the head of the Bristol Channel, the back-barrier environments along the Frisian Islands, and the open coast of the Wadden Sea. In the few remaining salt marshes on these complexes, typical vegetation zones and plant associations are similar to those described for northern France, with the recurring presence of S. anglica in the pioneer zone. This fertile allopolyploid arose by the end of the 1880s in Southampton Water, UK, by chromosomal doubling from Spartina xtownsendii, a hybrid between the introduced Spartina alterniflora and the native Spartina maritima (Ayres and Strong, 2001). Although S. anglica invaded large areas during the first 30 years after being reported (Raybould, 1997), during the late 1920s and early 1930s the species began to show signs of loss of vigor at some locations, leading to widespread diebacks. This process of recession, first described by Goodman et al. (1959), has continued up to the present. However, the concept of comprehensive spread followed by sudden decline cannot be generalized, and there are some locations where S. anglica has maintained its vigor and capacity for colonization of mudflats, displacing eel grass (Zostera) and algae (Raybould, 1997). In addition to a rapid natural spread following its appearance, S. anglica was also extensively planted to stabilize soft sediments, resulting in a considerable expansion throughout the British Isles and nearby Europe over a relatively short period. The rapid colonization of S. anglica over extensive flats can be viewed as an autogenic control whereby plants may cause a rapid buildup of sediment, and thereby significantly affect coastal dynamics (Allen, 2000).
4.2. Eastern North America In the temperate zone of eastern North America, marshes dominate the intertidal zone from southern Canada to the northern Gulf Coast, excluding subtropical coastlines of Florida, south of approximately 30 N (Savage, 1972), where the salt marshes are replaced by a mixture of mangrove forests and tidal marshes. The presence of barrier islands is a conspicuous feature all along the Atlantic coast of North America, occurring from Maine to Texas. Barrier islands are formed and maintained by changing sea level that makes them extremely dynamic and ephemeral landforms on a geological timescale. In eastern North America, different barrier
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complexes share a common origin, closely related to the sustained sea-level rise during the last 4,500–5,000 years. The Holocene landward migration of shorelines led to the transgression of older barrier islands and dune ridges, and the inundation of lowlands behind them created extensive protected back-barrier environments, suitable for wetland development (Oertel et al., 1989). Despite similarities in the Holocene sea-level history, there are several structural differences between North American and European coastal marshes. In eastern North America, S. alterniflora grows in the lower intertidal zone, with the lower elevational limit of occurrence near the tidal datum of mean low tide (McKee and Patrick, 1988). In contrast, vegetation of European salt marshes is typically confined to the upper intertidal zone, at elevations between mean high tide and spring tide (Lefeuvre, 1996). Moreover, pioneer species in European salt marshes form a short prairie, instead of those dense tall stands of S. alterniflora described on the Atlantic coast of North America. Following Chapman’s geographical groups, tidal marshes of eastern North America can be divided into three subgroups: the Bay of Fundy, New England, and Coastal Plain. 4.2.1. Bay of Fundy The Bay of Fundy is located on the east coast of southern Canada, where it forms an extension of the Gulf of Maine. Semidiurnal tides of about 4 m high near the mouth amplify to an average height of about 11 m in the upper reaches of the bay. The most remarkable feature in this area is that marshes here grow under the strong tidal currents generated by some of the largest tides in the world. Characteristic marsh topography results in a narrow low marsh fringing a wider high marsh area. At some locations, undercutting of creek banks generates a sharp topographic division separating the two zones. These escarpments also appear in southern New England marshes (Miller and Egler, 1950) but, being a transition between temperate and subarctic regions, marsh geomorphology in the upper Bay of Fundy is more strongly affected by ice formation. The movement of blocks of frozen sediment causes both erosion and sedimentation in the intertidal zone (Dionne, 1989). Typical salt marsh vegetation is flooded frequently by water as much as 4 m deep, and consists almost exclusively of S. alterniflora. The high marsh zone is flooded infrequently for short periods by only extreme high tides. High marsh vegetation is more diverse, usually dominated by Spartina patens, although different species such as Salicornia europea, Triglochin maritima, and J. gerardi may also be dominant (Chmura et al., 1997). Common species also found in these high marshes include Atriplex patula, Glaux maritima, S. maritima, Spergularia canadensis, Limonium nashii, Hierochloe odorata, Elymus arenarius, Distichlis spicata, and P. maritima (van Proosdij et al., 1999). At some locations, Chmura et al. (1997) have described a middle zone between the high and low marshes where S. alterniflora and Plantago maritima are codominants. The presence of P. maritima and extensive stands of Juncus balticus, common species at higher latitudes, indicates the transition from the subarctic to the more temperate marshes of New England. Phragmites australis and Iva frutescens often mark the upper limit of the high marsh (Jacobson and Jacobson, 1989).
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4.2.2. New England Tidal marshes in New England occur along the Atlantic coast from Maine to New Jersey. In the words of Chapman (1960), the distinctive feature of these marshes is that they are built of a marine peat. The particular processes driving marsh peat development in New England raised an early scientific interest, as reflected by the works of Mudge (1862), Shaler (1886), Davis (1910), and Johnson (1925). After Redfield (1972), a new comprehensive model was accepted, highlighting the relevance of Holocene sea-level history on the evolution of New England marshes (Figure 4). In Redfield’s model with a sufficient sediment supply, the marsh would have responded to slowly rising sea level (approximately 1 mm/year) by extending landward and covering uplands with marsh peat. Marshes prograded seaward through fine sediment accumulation and transgressed landward through high marsh development over the older intertidal peat. Similar to marshes in the Bay of Fundy, vertical zonation includes a low marsh flooded daily by tides, and covered by dense pure stands of S. alterniflora. At higher elevations, marshes are dominated by S. patens, with the upland border dominated by monocultures of J. gerardi. The occurrence of a narrow strip of I. frutescens usually delineates the terrestrial border (Bertness et al., 2002), and where a brackish transition exists, given sufficient freshwater input, Zizania aquatica, Scirpus americanus, Typha spp., and P. australis often appear as dominants (Chapman, 1960; Nixon, 1982). Although New England marshes are typically described as a regional unit, the more severe winter temperatures in northern marshes and the high frequency of icing events may cause severe damage to low marsh vegetation, preventing the establishment of dense and pronounced monocultures of S. alterniflora common in southern marshes. Besides climate, northern and southern marshes differ in the recent history of human alteration. During 1960–1990, the human population along the Long Island Sound increased at twice the rate experienced in the Gulf of Maine (Shriver et al., 2004). As a result, the remaining marshes in southern New England are smaller and well drained, while the larger marshes in the Gulf of Maine are more likely to have waterlogged pannes. Vegetation of these pannes consists of forbs such as Agalinis maritima, A. patula, G. maritima, L. nashii, P. maritima, Salicornia europaea, Suaeda linearis, and T. maritima, many of which occur at low abundances further south (Shumway and Bertness, 1992) but form a persistent community in the north.
High water level Time = 3 High marsh peat
Upland
Time = 2 Intertidal peat
Sand
Time = 1 Time = 0
Figure 4 Evolution of Barnstable Marsh, Massachusetts, USA (after Redfield, 1972). Each time sequence would be affixed to the upland surface on the left.
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A recent expansion of P. australis has been observed on the landward edges of New England salt marshes. This large reed is a common species historically appearing at lower abundances in nonsaline marshes. Over the past 150 years, the species has increased its cover in freshwater and brackish tidal wetlands along the middle Atlantic, displacing other well-established wetland plant (Saltonstall, 2002). Although P. australis expansion was thought to be limited by high soil salinities (Chambers et al., 1999), observations in southern New England indicate that this reed is actually invading salt marshes, and the spreading could be related to humaninduced changes in nutrient regimes (Bertness et al., 2002). Disturbance and periodic long-term changes in tidal regime (Chambers et al., 2003), in addition to nutrient enrichment, appear to play an important role in P. australis expansion. 4.2.3. Coastal Plain Marshes of the Coastal Plain occur in the broad coastal zone from New Jersey to the Gulf Coast. Along the southeastern Atlantic Coast, salt marshes usually develop in extensive shallow areas lying behind Pleistocene barrier islands (Figure 5). In the Gulf Coast, extensive salt and freshwater tidal marshes are associated with the Mississippi River Delta. South of the Chesapeake Bay to northern Florida, S. alterniflora dominates these extensive intertidal low marshes that cover most of the coastal marsh area. On levees and creek banks, the tallest and most productive form of the species appears as a narrow band. In the back levee low marsh, where tidal inundation is still several hours a day, a wide zone of vigorous plants extends up to higher elevations, where tidal inundation is just a couple of hours a day, and the short form of S. alterniflora becomes dominant. At even higher elevations, where flooding by tides is irregular, a high marsh dominated by Juncus roemerianus and S. patens is common along with inclusions of Salicornia spp., D. spicata, and Limonium spp. and salt bare pans may also occur. In the brackish transitions to freshwater communities, Spartina cynosuroides commonly forms extensive stands that may mix with smaller stands of S. americanus and Pontederia cordata (Wiegert and Freeman, 1990). In the Albemarle-Pamlico Sound, North Carolina, coastal rivers flow into large sounds with negligible tides. Consequently, marshes are dominated by the higher elevation species (J. roemerianus and S. patens) from the eroding shore line to the landward upland forest (Moorhead and Brinson, 1995). In the Mississippi River Delta marshes, there is a general pattern of decreasing salinity from the coast inland. Visser et al. (1998) recognize nine different vegetation types: (1) two in the polyhaline zone, one occupied by black mangrove (Avicennia germinans with S. alterniflora and Batis maritima) and the other dominated by oyster grass (S. alterniflora sometimes in association with J. roemerianus); (2) in the mesohaline zone a codominant mixture of S. alterniflora, S. patens, and D. spicata; (3) mesohaline wiregrass dominated by S. patens; (4) an oligohaline wiregrass dominated by S. patens, but with a higher species richness; (5) an oligohaline mix dominated by Sagittaria lancifolia; and (6) a freshwater zone, composed of three different types – the fresh bull tongue dominated by S. lancifolia with the presence of ferns, the fresh maidencane dominated by Panicum hemitomon, and the fresh cutgrass dominated by Zizaniopsis miliacea.
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Barrier Is Delmarv lands of the a Penin sula Albem arle So und Palm lico S ound Sea I sland C of So uth C oastal Re arolin g a and ion Geor gia
icola Bay Apalach nd ay ile B Sou Mob ssippi si Mis
Ga lve sto Ma Sa ta n Ba Co n A gor y nt da rp us oni Ba y o Ch Ba ris y ti Ba y
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Back levee low marsh
Salt pan
Salicornia spp. Distichlis spicata Juncus roemerianus
Levee marsh
Low marsh Transition marsh
High marsh
Spartina alterniflora Short form Spartina alterniflora Tall form
Figure 5 Location of main wetland complexes associated with barrier island systems along the temperate Coastal Plain of the United States, and cross section of a typical Southeast Atlantic coast salt marsh (after Wiegert and Freeman, 1990).
S. alterniflora marshes along the Mississippi River Deltaic Plain have recently undergone extensive diebacks, characterized by the premature browning of this species. This phenomenon has affected approximately 150 km of coastline and has damaged over 100,000 ha of coastal salt marshes along the Gulf Coast (www.brownmarsh.net). Several factors may be interacting to cause marsh browning, but it is now believed that combined elevated salinity and soil drying may interfere with nutrient uptake.
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4.3. Western North America Coastal wetlands on the Pacific coast of North America are less extensive due to the morphology of the shoreline, and the tidal regime is different from the Atlantic coast, with a more pronounced mixed tide, which likely influences low and high marsh distribution. Main wetland areas occur in San Francisco Bay estuary, a drowned river valley, formed by crustal movements during the late Pliocene and later inundated by rising Holocene sea level (Atwater et al., 1977). Natural salt marshes in the lower portion of the estuary are dominated by Spartina foliosa in the low marsh and Salicornia virginica in the high marsh, above mean high water level. Marshes dominated by S. virginica are also found in diked wetlands where tides have been excluded. Within the high marsh, Jaumea carnosa and T. maritima appear at lower abundances (Figure 6), and D. spicata grows on elevated sites (Josselyn, 1983). However, these natural marshes occupy a relatively small area, since wetlands in San Francisco Bay have been converted to salt ponds, urban and industrial land uses, and agriculture on a large scale. S. foliosa has a narrow range of distribution, limited to the Pacific coast of North America, from Humboldt Bay to Baja California. Similar to S. alterniflora, S. foliosa
(a)
Spartina foliosa
Grindelia stricta
Salicornia virginica Jaumea carnosa Triglochin maritima
Pan Tidal creek
(b) exotic Spartina spp. Short form
exotic Spartina spp. Tall form
Accreted sediment
Figure 6 (a) Cross section of natural salt marshes in the lower portion of the San Francisco Bay estuary (after Josselyn, 1983) and (b) modifications introduced by the invasive exotic species of Spartina (after California State Coastal Conservancy/US Fish and Wildlife Service, 2003).
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takes on a tall or “robust” form, which grows at lower elevations, and a “dwarf” form at elevations closer to mean high water. Further north, exotic species are invading including the hybrid S. alterniflora S. foliosa in the San Francisco Estuary, where it has been reported to increase in area a 100-fold since the 1970s, and Spartina densiflora, a South American native, ranking second in area covered, also appearing in Humboldt Bay. Finally, S. patens and S. anglica, the latter a hybrid species of S. alterniflora and the European S. maritima, are also found, but have not yet dispersed far beyond their introduction sites (Ayres et al., 2004). In Willapa Bay, nearly one-third of the area originally covered by mudflats is now infested with S. alterniflora. This species was introduced into Willapa Bay during the late 1800s but was not identified until the 1940s. During the first 50 years, the population expanded slowly, but from 1945 to 1988 the plant spread rapidly throughout the bay, resulting in severe habitat alteration as unvegetated mudflats were converted to salt marshes (Simenstad and Thom, 1995)
4.4. Mediterranean Wetlands in the Mediterranean region are mainly associated with deltas, such as those of the Rhoˆne, Ebro, Nile, and Po, where the deceleration in the rise of sea level occurred from 7,000 to 6,000 years BP favored a rapid deltaic progradation, dependent on alluvial sediment supplies (Vella et al., 2005). Similar to the Mississippi Delta, Mediterranean systems are subsiding, and many coastal wetlands in this region might face elevated rates of relative sea-level rise (Day et al., 1995). The Rhoˆne Delta (La Camargue) is one of the most important natural areas remaining in this region. Natural halophytic vegetation is composed of low shrubby species, instead of the herbaceous one of the genus Spartina. Typical salt marsh communities would include species such as Sarcocornia fruticosa and Arthrocnemum macrostachyum in the low marsh; an intermediate zone with Limonium virgatum, Limonium girardianum, Frankenia pulverulenta, and Artemisia galla; and Juncus spp. in a high marsh. (Chapman, 1960). However, these natural marshes have virtually disappeared as a result of human agricultural practices, and ricefields, although completely different in structure and function, have become the main remaining wetland habitat. Similar conditions apply for other deltas, especially in the northern Mediterranean, where ricefields may offer a valuable habitat for waterbirds (Tourenq et al., 2001), but their hydrology is no longer controlled by sea level.
4.5. Eastern Asia Major areas of marshes in eastern Asia occur along the temperate coasts of the East China Sea and the Bohai Sea where large rivers like the Changjian (Yangtze River) and the Huanghe (Yellow River) develop extensive deltas. The high precipitation due to the monsoonal climate and the presence of large rivers and the high sediment supply, in combination with the Holocene relative sea-level history, are major reasons for the presence of extensive deltas in Asia. These monsooncontrolled coasts have been formed under a stable or slightly falling sea level over
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the past 6,000 years. In most coastal regions of eastern Asia, the middle Holocene highstand was 2–4 m above the present relative sea level (Saito, 2001), and deltas began to build by sediment deposition and shoreline progradation as sea level fell to the present position. The Yangtze River, the third largest in the world, carries a huge amount of fine sediments into the East China Sea. Most of these sediments are deposited in the estuary, developing an extensive delta that is prograding at a rate of 15–20 km2/year (Yang et al., 1997). The intertidal zone in the Yangtze Delta extended over about 1,550 km2, but an area of 750 km2 has been reclaimed in the past half century, and approximately onethird remains as salt marsh. The lower intertidal zone is typically characterized by mudflats lacking any vegetation cover, with Scirpus mariqueter and Scirpus triqueter becoming the pioneer species at elevations higher than neap tide high water level. In the high marsh, P. australis is the common dominant species and a variety of species such as Imperata cylindrical, Suaeda glauca, Juncus setchuensis, and Carex scabriflora may also appear in small patches (Gao and Zhang, 2006). In 1979, S. alterniflora was transplanted into tidal marshes of coastal China to stabilize tidal flats. Since its introduction, this species has gradually invaded the former P. australis communities, and the upper S. mariqueter zone, aggressively replacing native species (Cheng et al., 2006). At present, the plant zonation of coastal wetlands in the Jiangsu Province typically includes S. alterniflora as the pioneer species dominating the upper intertidal zone, a Suaeda spp. community landward, and a supratidal zone, where P. australis, I. cylindrical, Aeluropus littoralis, Scirpus karuizawensis, and Zoysia macrostachya may be dominant (Liu et al., 2007) (Figure 7).
4.6. Australasia Similar to eastern Asia, the coasts of Australia and New Zealand have also experienced a stable or slightly falling relative sea level since the middle Holocene highstand. However, the internally arid continent in Australia constrains the sediment supply to coastal areas, precluding the formation of large deltas and allowing the persistence of unfilled estuaries behind barriers. Although salt marshes are associated with temperate coasts worldwide, in Australia they also occur in tropical coastlines, landward of mangroves (Figure 8). The only exception to this pattern occurs in Victoria, the southern distribution of Avicennia marina, where the introduced S. anglica grows at lower elevations than mangroves. Temperate salt marshes in Australia only occur in the southern coasts of the continent. These marshes are generally smaller in area but support a higher number of species compared to the northern ones. A general pattern of zonation includes, from low to high marsh, communities dominated by Sarcocornia quinqueflora, Sporobolus virginicus, and Juncus kraussii. The arid climate on the southwestern coast determines a more open vegetation cover in the high marshes, and a more shrubby physiognomy, while on the wetter eastern coasts denser vegetation occurs, with a higher presence of grasses and sedges (Adam, 1990). Most estuarine areas in New Zealand are shallow lagoons, protected by sand bars. Typical marsh zonation includes a pioneer zone dominated by S. quinqueflora and Juncus maritimus, and a
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Jiangsu coast
Yangtze river
Supratidal zone
Aeluropus littoralis Phragmites australis Imperata cylindrical Scirpus karuizawensis Zoysiam jacrostachys
Hightidal zone
Intertidal zone
Lowtidal zone
Subtidal zone
Suaeda salsa Suaeda glauca Spartina alterniflora
Figure 7 Typical sequence of plant species in coastal salt marshes of the Jiangsu Province, China (after Liu et al., 2007).
Highest spring Mean high spring Juncus
Mean sea level
Sarcocornia Avicennia Samolus
Suaeda
Figure 8 Salt marshes occurring landward of mangroves in South NewWales, Australia (after Adam, 1990).
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middle marsh with Juncus articulatus. At higher elevations, beyond mean high tides a salt meadow develops, where species such as Selliera radicans, Cotula coronopifolia, and Plagianthus divaricatus start to appear (Bishop, 1992).
4.7. South America Similar to North America, wetland development along the rugged western coasts of South America is restricted mostly to small barrier lagoons and fjord-head deltas in the south of Chile. In contrast, the Atlantic coast is occupied by a large number of coastal salt and freshwater marshes, from the temperate southern coasts of Brazil to Tierra del Fuego Island in Argentina. Although similar in the geographical configuration of the coast, lying along wide and shallow continental shelves, a falling or fluctuating relative sea level characterized South American late Holocene (Isla, 1989), instead of the rising sea level described in North American Atlantic coast. In the northern coasts of Argentina, several authors locate a highstand about 6,000 years BP, when the relative sea level reached around 6 m above present (Cavallotto et al., 2004; Isla, 1989; Violante and Parker, 2000). Southwards, along the Patagonian coasts, several Holocene marine shorelines can be found at different elevations, with a southward increase in terrace elevation for terraces of the same age, suggesting a tectonic uplift effect (Rostami et al., 2000). The late Holocene falling trend in relative sea level has resulted in wide low-lying areas of former estuarine environments and typical regressive forms, like extensive plains composed of beach-ridge and lagoonal deposits. This regressive landscape, which is at present unconnected to tides, would become increasingly inundated under the recent rising trends in relative sea level. In the coastal plain extending southward from the Rı´o de la Plata Estuary, in the northern coasts of Argentina, S. alterniflora is the dominant species in the low marsh, forming monospecific stands in the zone inundated by all high tides. At higher elevations, S. densiflora marshes are the most widespread, where typical associated species such as Apium sellowianum, Limonium brasiliense, and Cortaderia selloana may also appear. At some places, monospecific stands of Sarcocornia perennis occur between S. alterniflora and S. densiflora marshes, and a mixed marsh type with S. densiflora, S. perennis, Juncus acutus, and D. spicata is commonly found in elevated sites (Cagnoni, 1999). In a gradient of increasing aridity southward, vegetation becomes sparse in the intertidal zone. At Bahı´a Blanca, at the northern limit of the Patagonian desert, the intertidal S. alterniflora marshes only occur as discontinuous patches near the mouth (Figure 9). Under the seasonally hypersaline conditions in the head of the estuary, vegetation is virtually absent in the intertidal zone except for the circular mounds of S. perennis colonizing the upper portion. At higher elevations, typical southern species such as Heterostachys ritteriana, Suaeda patagonica, Cressa truxiliensis, and Allenrolfea patagonica form extensive halophytic communities. In the southernmost coasts of Argentina, at San Sebastia´n Bay on Tierra del Fuego Island, S. perennis is the pioneer species in the upper intertidal zone, while Puccinellia sp. is much less abundant and scattered. At higher levels, pastures dominated by Puccinellia magellanica, Puccinellia biflora, and S. perennis usually cover the abandoned tidal flats (Collantes and Faggi, 1999).
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Figure 9 Isolated patches of Spartina alterniflora colonizing extensive mudflats in the Bahı´a Blanca Estuary (courtesy of Gerardo Perillo).
5. HUMAN IMPACT AND C LIMATE C HANGE 5.1. Human impact The evolution of salt marshes over time shows a strong abiotic influence from relative sea-level and tidal drainage pattern as well as both direct and indirect effects from climate and human populations. Human alterations are perhaps more readily associated with land claims and engineering works for navigation and erosion/flood protection, which directly alter salt marsh geomorphology and estuarine dynamics, as well as environmental pollution in more recent times (Doody, 1992; Viles and Spencer, 1995). Indirect human alterations include the delivery of pollution from far-field sources originating from rivers, and allied changes to terrestrial sediment delivery. Land clearance since the Mesolithic and, especially, the period from the Neolithic to the Roman era has significantly impacted on the erosion potential in the United Kingdom (Dearing and Zolitschka, 1999; Kalis et al., 2003; Macklin, 1999) and during the 18th through the 20th centuries on the east coast of North America (Pasternack et al., 2001). When coupled with increased runoff, sediment delivery to the coast has been greatly enhanced. Irrespective of large-scale changes in coastal sediment budget, the preservation in the sedimentary record of human impacts on the landscape is evident in the pollen records of clearance and introduced species (Brush et al., 1982; Mudie and Byrne, 1980), heavy metal records of mining activity (Plater and Appleby, 2004), and more recently, near-field pollution from urban and industrial expansion (Croudace and Cundy, 1995). The combination of land claim and resulting changes in estuary dynamics, that is, reduction in prism, has had an important bearing on the long-term persistence or “preservation” potential of temperate coastal wetlands. The fact that many former and extensive perimarine wetlands are largely absent from European coastal
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lowlands is primarily due to land claim (e.g., the Somerset Levels, Romney Marsh, and Fenland systems in the United Kingdom). Similarly, several examples exist where large-scale “reclamations” have led to significant changes in the nature and pattern of intertidal drainage and, hence, the erosion of coastal wetland sediments. Indeed, cyclic changes in the proximity of shore-parallel drainage creeks may result in the formation of “stepped” salt marsh profiles where a more recent phase of accretion is found seaward of a former marginal cliff and levee. Examples of this are found in the Severn Estuary (Allen and Rae, 1987), the Morecambe Bay (Pringle, 1995), and the Wadden Sea (Jakobsen, 1954). Former marshes in northern Europe have been extensively diked over the past 2,000 years, and the few surviving areas are reduced in size and strongly modified. On the German coasts of the Wadden Sea, approximately 1,000 km2 of former coastal marshes have been diked over the last millennium. Continuous embankments of newly accreted foreland have shaped the mainland coasts of the northern Netherlands, where anthropogenic salt marshes extend over 190 km2, and virtually no natural tidal marshes remain (Lotze, 2004). In the Severn estuary, about 840 km2 of marshes have been impounded since the end of the Roman occupation, whereas only 14 km2 of active marsh remains (Allen and Duffy, 1998) and similar relations between active and embanked marshes would also apply for France and Denmark (Allen, 2000). Prior to European settlement in the Bay of Fundy, marshes covered wide areas of the Minas Basin and the upper reaches of Chignecto Bay, where large amounts of fine sediments accumulate. However, during the past 400 years lowlands have been intensely diked and reclaimed, with an estimated reduction of about 70% of the former marsh area (Gordon and Cranford, 1994). Agricultural expansion is the major cause of wetland losses. Since the early 1800s, wetland conversion to agriculture is estimated at over 20 million hectares, including 65% of the coastal marshes of Atlantic Canada. In Eastern Asia, several human activities, such as dredging and deepening of navigation channels, water diversions to northern China and large-scale land reclamation are thought to remarkably change the Yangtze Delta environment in the next few decades (Xiqing, 1998). The Three Gorges Dam will be the largest hydroelectric dam in the world. Its construction began in 1994, the reservoir began to fill in 2003, and it is expected to become fully operational in 2009. Since cultivation in this region dates back to more than 7,000 years, human activities have played a mayor role in shaping present landscapes. With more than 50% of the world’s population living in Asia, and most of the world’s ricefields occurring in Asian deltas (Galloway and Melillo, 1989), human pressures on the contributing watersheds have greatly influenced the past and present coastal processes. The future development of big Asian deltas may be driven, to a great extent, by largescale engineering projects. The temperate coasts of Australia have also been greatly affected by human activities. Although losses have been small, compared to the original wetland area (Adam, 1990), reclamation for industrial and human development has been concentrated on the southeastern coasts. Besides reclaimed areas, the main threats to salt marshes are ecosystem degradation due to alterations in hydrologic regimes, pollution, and weed invasion.
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The presence of large wetland complexes, comparatively little affected by human alterations, is an attribute that distinguish South America. Major transformations in the Bahı´a Blanca Estuary relate to the presence of the largest deepwater harbor in Argentina, but human activities are concentrated in a relatively small section on the shore, leaving wide tidal flats in the southern estuary, as well as most of the islands in a nearly unmodified condition. Southwards, demographic pressures decrease, and large sections of the Argentinean coast remain little affected by anthropogenic transformation.
5.2. Climate and sea-level change There has been considerable discussion as to how temperate coastal wetlands will fare in the future under climate-enhanced sea-level rise (Reed, 1990; Simas et al., 2001). Indeed, early scoping studies (Boorman et al., 1998; Titus, 1987) predicted the large-scale loss of coastal wetlands as a consequence of sea-level rise exceeding sediment supply (Temmerman et al., 2004), or the influence of sea-level rise on marsh productivity (Morris et al., 2002). However, there is some evidence to suggest that, at some locations, the geomorphic response of salt marshes is not sediment limited. Many salt marshes built from allochthonous sediment show a significant excess of vertical sediment accretion relative to sea-level rise (French, 2006), where the net sediment flux across the intertidal zone during the tidal cycle is far in excess of that deposited. Recent work on back-barrier intertidal wetland sediments of Romney Marsh and Dungeness, United Kingdom, suggests that rapid tidal accretion (in the order of 0.5 m/year) follows changes in coastal morphology and accommodation space driven by storms and coastal progradation (Stupples and Plater, 2007). Deltaic regions are particularly vulnerable to a relative sea-level rise because of a rapid subsidence (Day et al., 1995). Under this scenario, river sediment supply is one of the most important agents in shaping deltaic evolution, and human-induced changes in sedimentary fluxes have been occurring in this region since the beginnings of the historical period. Dam construction and the increase in water demand for agriculture, industry, and tourist development has dramatically reduced the sediment load of rivers, and is thought to be a major cause of deltaic degradation when coupled with subsidence (Sanchez-Arcilla et al., 1998; Stanley and Warne, 1993). Erosion of salt marsh sediments, especially microtidal systems that are more strongly influenced by storms (Stumpf, 1983), also raises concerns regarding geomorphic response to relative sea-level rise. While temperate coastal wetland subenvironments are expected to advance landward with rising tidal levels, the outer edge of the salt marsh and mudflat may experience higher tidal flow velocities and wave energy, thus leading to erosion (Pethick, 1993). Although this then supplies mud to the tidal sediment budget, these resuspended muds contain the accumulated pollutants of years past, many of which are now banned or strictly regulated due to their detrimental environmental and health impacts (Valette-Silver, 1993; Williams et al., 1994). For example, the erosion of sediment from the Mersey Estuary will release DDT, mercury, lead, and radionuclides into suspension for their potential
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redeposition (Fox et al., 1999). In this respect, pollutants that are currently regarded as being locked away in sedimentary reserves may become the main sources of environmental pollution in years to come.
6. S UMMARY Temperate coastal wetlands include a large variety of environments, ranging from tidal flats to fen woodland communities or barren tidal flats. While tidal flats typically occur at the seaward end, there is a large variation in plant communities at the landward margin, depending on climatic and groundwater influences. In humid climates, sea level controls the groundwater position providing the waterlogged conditions necessary for fen wetlands development. Hageman (1969) termed “perimarine zone” the area where freshwater wetlands persist under sea-level control. In arid climates where freshwater inputs are scarce, salt flats occur in the perimarine zone, where the elevated levels of soil salinity eliminate vegetation. In the intertidal zone, the primary abiotic control on wetland structure and function is the hydroperiod. In the “ramp” model of salt marsh accretion the hydroperiod and the availability of sediments determine the amount of particles settling on the salt marsh surface. Under the ramp model, sites located lower in the tidal frame experience more sediment deposition on each tide and, as elevation increases, hydroperiod and net sedimentation are reduced. In the “creek” model of salt marsh accretion, tidal channels capture and funnel the incoming tide, explaining the observed increase in sediment grain size with proximity to the creek margin. Levees generally develop on creek banks where more of the coarser sediment load is deposited. While the ramp model accounts for widespread and gradual trends in vertical accretion, the creek model is focused more on the local scale and potentially rapid development of three-dimensional sedimentary features. The result of the interaction between hydrodynamics, elevation, and vegetation is a shore-parallel zonation of plants which is made more complex and spatially variable by the micromorphology of the marsh surface. The lower marsh, regularly inundated by salt water usually consists of hardy pioneer genera, and at higher levels a greater diversity of plants can colonize. In humid climates there would be a gradational transition from the upper salt marsh limit to a freshwater reed swamp, which creates a clear zonation of plant species at different elevations. In arid climates, however, salinities can exceed the limits of even the most tolerant halophytes, and salt flats devoid of vascular vegetation develop near the upland boundary. Tidal marshes through the temperate zone would be comparable in terms of ecosystem structure and function. However, substantial differences arise among major geographic regions firstly due to precipitation and secondly due to the biogeographic distribution of species. In addition, human activities have also exerted profound changes in coastal wetlands, and the processes of wetland loss and degradation have been quite variable in space and time, leading to major regional differences in the extent and ecological integrity of remaining wetland areas.
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The evolution of salt marshes over time is strongly influenced by climate change and human populations. Human alterations are associated with land claims and engineering works for navigation and erosion/flood protection, as well as environmental pollution. Indirect human alterations include the delivery of pollution from far-field sources and allied changes to terrestrial sediment delivery. Considering a future climate-enhanced sea-level rise, early studies predicted the large-scale loss of coastal wetlands as a consequence of sea-level rise exceeding sediment supply, or the influence of sea-level rise on marsh productivity. However, there is some evidence to suggest that at some locations, the geomorphic response of salt marshes would not be sediment limited.
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P OLAR C OASTAL W ETLANDS : D EVELOPMENT , S TRUCTURE , AND L AND USE I. Peter Martini, Robert L. Jefferies, R. I. Guy Morrison, and Kenneth F. Abraham
Contents 1. 2. 3. 4. 5. 6. 7.
Introduction Geology/Geomorphology Oceanography Climate Structure of Coastal Wetlands Vegetation of Polar Coastal Wetlands Fauna of Polar Coastal Wetlands 7.1. Invertebrate fauna 7.2. Vertebrate fauna using coastal wetlands 8. Environmental Hazards 9. Conclusions and Research Priorities References
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1. INTRODUCTION Polar coastal wetlands mostly consist of salt- and brackish/fresh-water marshes, laida (coastal tundra inundated by seawater during storm surges or by freshwater at the time of snow and ground-ice melt), and coastal tundra plains with numerous ponds and shallow lakes in Arctic and sub-Arctic zones affected by permafrost (Figure 1). Vegetated coastal wetlands are found along every northern coastline, although locally they are often poorly developed especially on coasts of polar deserts. The deserts have a very cold climate (less than 10C average during the warmest month of the year), very low precipitation (less than 250 mm/year to as low as 45 mm/year), and extreme poverty of life (Callaghan et al., 2005). In the north, coastal marshes are well represented in low Arctic and sub-Arctic lowlands. They may bound wide tidal flats in protected embayments or develop behind coastal barriers partially inundated by tides where fine sediments (mud to fine-grained sand) can accumulate and vegetation grows under waterlogged Coastal Wetlands: An Integrated Ecosystem Approach
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Figure 1 Wetlands. (a) Distribution of wetlands on Earth; polar wetlands are characterized by the presence of permafrost. (b) Distribution of Arctic and sub-Arctic zones in which polar wetlands are located. (c) Distribution of the major types of Wetlands Regions of Canada (the Boreal Wetland Zone is further subdivided into various sectors as indicated in Zoltai, 1980) (FB, Foxe Basin; fh, Fury and Hecla Strait; GL, Great lakes; HB, Hudson Bay; hs, Hudson Strait; JB, James Bay; np, North Point; md, Mackenzie River Delta; pk, The Great Plain of the Koukdjuak). [Compilation of data from U.S. Dept. of Agriculture (1996), UNEP/GRID (2006), Zoltai (1980)].
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conditions. Laida and other coastal tundra also develop on low-lying polar coastal plains. Equivalent types of marine coastal wetlands do not develop to any extent in the Southern Hemisphere because few ice-free lands occur south of the Antarctic Circle. In addition, surface marine currents flow latitudinally and, unlike the Northern Hemisphere, do not refrigerate the continents at lower latitudes. Some of the largest polar salt marshes have developed in the Hudson Bay Lowland, on the coasts of Hudson and James bays. Some of the largest laida and tundra coastal plains occur in the southeastern corner of the Foxe Basin, along the Arctic Coastal Plain of Alaska and Yukon, and along the Russian Arctic coast. Brackish marshes and other coastal wetlands develop extensively on deltas/ estuaries of rivers that flow into Hudson and James bays and the Arctic Ocean, such as in the deltas of the Mackenzie River (Canada) and Lena River (Russia) (Figure 1). Multiyear, multidisciplinary studies of wetlands and their land uses have been carried out along the western coasts of James Bay (JB) and southwestern Hudson Bay (HB) and to a lesser extent in the Foxe Basin (FB) and some of the Canadian Arctic Islands. Variations in sediment, soil, vegetation, and distribution of infauna and migratory birds were recorded along selected transects from the sea to the upper marshes. Regional physiographic variations of the coastlines along a south–north transect were also examined (Figure 1c). In addition, detailed multiseason analyses of local estuarine areas and selected coasts have been made, such as at North Point (np) in James Bay (Figure 1c), which allow for benchmark comparisons to be made with coastal wetland sites elsewhere.
2. GEOLOGY /G EOMORPHOLOGY The Arctic regions have been affected by several orogenies (mountainbuilding episodes) and are mountainous over large tracts. A series of terranes (fragments of the Earth’s crust) accreted during the Mesozoic in present Arctic areas where mountain ranges developed in northeastern Russia, Alaska (such as the east–west trending Brooks Range (BR)), Canadian Arctic (such as the Richardson Mountains (RM) and Innuitian Mountains), and northern Greenland. Older Caledonian mountain belts developed during the Paleozoic in eastern Greenland, northern Europe, and the central parts of Russia (Ural Mountains) (Figure 2). The Kamchatka Peninsula in northeast Russia and the Aleutian Islands to the west of Alaska are two active tectonic areas where continuing subduction of the Pacific tectonic plate has in the past generated, and is still generating, active volcanoes and rugged terrain. Nevertheless, extensive Arctic lowlands occur in between or in front of these mountain chains along the Arctic Coastal Plain of northern Alaska, in the Mackenzie River delta in NW Canada, and in parts of the Arctic Russian coastal areas west and east of the Ural
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Mountains. In northeast Canada, other extensive coastal plains occur in Arctic and sub-Arctic zones that extend southward along the coasts of Hudson and James bays bounded by the North American Precambrian Shield and underlain by low-lying Paleozoic rocks of old inland basins. Most polar areas have been variously glaciated during Late Pleistocene except in a few cold but dry areas in northernmost North America and Russia (Figure 3a,b). The large ice sheets were thicker at their epicenters (domes) and thinned toward saddle and peripheral areas. This had several consequences that still affect polar coastal areas. The weight of the glaciers depressed the Earth’s crust, and as the glacier melted, differential isostatic rebound has led to uplifts of more than 200 m. The rebound is still continuing at rates that vary from around 1 m per century where the ice was thicker near the centers of glaciation at mid-high latitudes to minimal rates of uplift in other areas. The isostatic uplift also led to land emersion from large lakes and seas that had formed in front of the glaciers, and subsequently extensive coastal plains developed. Land emersion (regression) continues in areas near former ice domes where the land uplift is more rapid than the present sea-level rise. This occurs along the western coasts of Hudson and James bays, along the Gulf of Bothnia between Finland and Sweden, and in the White Sea in northwest Russia. Where the land uplift was small and is no longer occurring and where there is no active neotectonism, a marine transgression is taking place, such as along the Arctic Coastal Plain of Alaska, northwestern Canada, and parts of the Siberian coastal plains. The extent of glaciation in central-north Russia is still unresolved: two major hypotheses have been put forward (Grosswald, 1998; Velichko et al., 1997). One hypothesis proposes a wide, thick glaciation during the Late Pleistocene, which could have generated postglacial isostatic uplift similar to that of North America and Fennoscandinavia (mainly Finland, Sweden, and Norway). The other
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Figure 3 Pleistocene glaciations of northern lands (after Flint, 1971; Fulton, 1989). (a) North America. (b) Arctic Russia; the wide glaciation hypothesis considers theYamal and Yenissei area to have also been glaciated in Late Pleistocene (lateWeichselian) ( adapted from Raab et al., 2003).
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suggests a restricted glaciation that generated smaller areas of postglacial isostatic uplift. For example, uplift is still occurring in the White Sea area in the Kola Peninsula, northwest Russia (Jeglum et al., 2003).
3. O CEANOGRAPHY The inshore Arctic seas are generally shallow, particularly along the Russian coasts, Foxe Basin, and James Bay. They and their respective coasts are subjected to a harsh climate and a seasonally variable ice cover that is more severe at the northernmost latitudes where ice is present throughout the year. Sea ice influences marine currents, tides, and waves that, in turn, affect the stability of coastlines although their potential action is limited to a few months of the year. The marine currents of the Arctic Ocean and the adjoining inland seas are complex (Figure 4). In North America Arctic waters enter the Canadian inland seas from the Fury–Hecla Strait in the northwest corner of Foxe Basin and are carried down into Hudson and James bays to latitudes of about 51 N, cooling off the surrounding lands. Conversely, the northern lands of Fennoscandinavia are warmed by a branch of the North Atlantic Drift current up to approximate latitude of 70 N, well above the Arctic Circle. The tides of the northern seas are generally microtidal (less than 2 m in amplitude), but several shores experience mesotidal excursions, such as those of Hudson Bay, James Bay, and the Barents Sea. Macrotidal excursions with tides exceeding 5 m and locally 10 m occur in a few embayments, such as Bristol Bay in west Alaska, Bowman Bay in the southeast Foxe Basin, Frobisher (Iqualuit) Bay in southeast Baffin Island, parts of Ungava Bay along the Hudson Strait, Mezen Gulf in east White Sea, and the Gulf of Shelikov in the northern Okhotsk Sea. The salinity of the Arctic seas may in some places vary drastically from season to season due to the formation and melting of the ice cover and to the variable input of fluvial freshwater during spring–summer freshets (floods). “Unlike tropical oceans, which are temperature-stratified (there is a thermocline), the Arctic Ocean [and adjacent Arctic seas] is [mostly] salinity-stratified (there is a halocline), although at high latitudes the ocean is much less stable. The temperature profile of Arctic waters is nearly uniform at 0 to 1C”(Linacre and Geerts, 1998). Brackish-water conditions develop in and near the estuaries of major rivers during the spring–summer freshets, and more marine saline conditions are reestablished later when the river discharge decreases drastically. One dramatic case occurs in the shallow, southern James Bay, where freshwater is injected into the shallow sea by the northward flowing river floods. The anticlockwise marine current of the bay moves these waters northeastward, freshening the eastern coast of the bay significantly more than the western coast.
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Figure 4 Marine currents (arrows) in the Arctic Ocean and adjacent seas (north-directed arrows just west of Iceland and Norway represent movement of warm water into the Arctic) (after AMAP, 1998).
4. C LIMATE Average annual low temperatures and amounts of precipitation decrease substantially from south to north, particularly where polar deserts are present, such as in the northernmost parts of the Russian and Canadian Arctic Islands. The cold climate and the strong variation in day length throughout the year greatly
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affect assemblages of plant and animal species. There is a particularly low species richness in polar deserts. A cryosphere has developed in northern areas, which includes ice formation over water bodies and within the ground (permafrost). The distribution of permafrost in northern lands does not regularly follow latitudinal alignments; rather, it is influenced by the heat redistribution brought about by atmospheric movements and marine currents. Accordingly, there is a southern dip of continuous and discontinuous permafrost in continental and mountainous areas of central Asia and, more relevant for this chapter, a latitudinal dip in central-east Canada where anticlockwise marine currents bring cold Arctic waters to low latitudes in Hudson and James bays (Figure 5). The effect of marine currents is dramatically evidenced in James Bay, where warmer fluvial waters are injected into the bay from the south and are moved northeastward along the shore by marine currents, leading to a significant northward shift of the permafrost zones on the east side of the bay.
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Furthermore, whereas new (modern) permafrost is actively forming at midlatitude where the land is actively emerging from the sea, Pleistocene relict permafrost is present in the high Arctic. Such relict permafrost has undergone adjustment to the variable Holocene climates, particularly during the postglacial temperature maximum (approximately 8,000–7,000 years BP) in northwest Canada and northern Alaska, as demonstrated by peat accumulations of that age, but which are now frozen and inactive. This has led to the development of various geomorphologic structures in these landscapes including numerous thermokarst features (Washburn, 1973; Ruz et al., 1992). Frozen soil constitutes a barrier to free groundwater movement. In the high Arctic, a thin (<1 m) surficial active layer thaws during the summer. Water moves by convection only within this thin layer, as the underlying permafrost for the most part impedes percolation, and, thus, waterlogged conditions are favored (Woo, 2002). Farther south in the discontinuous and sporadic permafrost area of the sub-Arctic, the groundwater movement is less influenced by the ground ice: near-surface water flow is only partially obstructed by locally persistent ice lenses. The groundwater flow is still very slow, though, because of the very low hydraulic conductivity of peat.
5. STRUCTURE OF C OASTAL W ETLANDS The characteristics of polar coastal wetlands depend on abiotic and biotic conditions. They range from extensive seashore meadows showing a transition from saltwater to freshwater on large coastal plains such as the sub-Arctic and Boreal marshes of James and Hudson bays, to narrow strips of depauperate vegetation in the mid- to high-Arctic where narrow wetlands mostly developed in swales between beach ridges and brackish-water systems in deltaic areas. Salt marshes are a characteristic landscape feature of low-lying Arctic coastlines (Jefferies, 1977; Macdonald, 1977; Bliss, 1993). The best-developed marshes of the low Arctic to sub-Arctic are those on the southern and western coasts of Hudson and James bays (Jefferies et al., 1979). They have developed in the last 300–400 years as the Hudson Bay Lowland has continued to emerge as a result of the isostatic uplift (Andrews, 1973; Mo¨rner, 1980). Two major types of marsh have formed: one on open coasts and the other in swales on coasts with beach ridges. 1. The widest salt marshes occur inland from open, extensive sand and mud flats (Figures 6a,b), and brackish marshes are formed in estuarine areas and on mainland coasts freshened by fluvial plumes (Figure 6c,d). Less welldeveloped marshes form on steeper shores with higher waves and low sedimentation rates, where limited fine-grained deposits occur in bouldery areas. The salt marshes open to the sea are much impacted by ice. During the melting season, ice floes are grounded and later lifted and removed by tides, or ice pressure ridges form (Figure 7a). This leads to the removal of marsh
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Figure 6 Structure of well-developed sub-Arctic coastal marshes, western James Bay. (a,b) Marsh of open coasts with a well-developed gradation from algal high tidal flats at the shoreline to inland freshwater marshes and peat-bearing fens. The diagrams show a cross section of the substrate stratigraphy, the surficial depositional features (barbed short line indicate sandy ripple marks), and erosional features by ice floes, progressive colonization of the raised coastal plain by grassy vegetation, shrubs, and trees, and, in the top short columns, the stratigraphy of the top 20 cm of the laminated and cross-laminated (inclined lines) recent sediments and peat (symbols of diagrams are explained in vertical profile in “b”) (after Martini et al., 2001). (c,d) Extensive brackish- to freshwater marshes on river-influenced coasts. (e,f.) Coast with beach ridges and bilaterally structured coastal wetlands in the interridge swales. (UM, upper marsh; LM, lower marsh; HTF, high tidal flat; UTF, upper tidal flat; LTF, lower tidal flat).
material frozen to the underside of the floes, or to scouring due to ice push. As the land emerges further and becomes more vegetated, typical jigsaw patterns of pools develop reflecting earlier ice action (Figure 7b). 2. Marshes also form in interridge swales that are inundated by high tides, where they acquire a bilateral vegetation distribution pattern with salt species closer to the mid-swale tidal creek and brackish and freshwater species farther
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away from it. Locally, the marsh deposits of the swales have well-sorted sand grains (dispersed or in thin lenses) received from the bounding beach ridges, either blown in by strong winds or as a result of washover events during heavy storms (Figure 6e,f ). In low Arctic to sub-Arctic zones, the generally muddy, wet sediments of marshes are modified (ripened) as incipient soils develop. Incipient Bkg to Bg horizons with grayish brown colors (2.5Y5/2) (Protz, 1982) occur in the upper marshes, and ferrans (iron precipitates) may form around plant roots in betterdrained parts of the system. Along transects from the shoreline inland, salt-marsh soils show a gradual increase in thickness of the surficial organic layer (never reaching the 30–40 cm thickness to qualify as peatlands), a decrease in sodium and chloride concentrations with an associated drop in electrical conductivity (e.g., from a seasonal average of 6.1 mS/cm in the lower marsh to 2.2 mS/cm in the upper marsh at North Point (np) in southwest James Bay; Figure 1c), a marked decrease in pH in the upper brackish/freshwater marshes, and a decrease in calcium carbonate equivalents. In some cases, a landward salinity inversion occurs with brackish marshes formed near the coast and saltwater marshes developing farther inland (Martini, 2006). Salinity of inshore seawater is often low (about
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12 g of dissolved solutes per litre of soil solution), because of the inflow of freshwater from large rivers that drain into Arctic seas. The more seaward sections of marshes are less saline than the landward sections where disturbance, lack of tidal cover in summer, and drying out of the terrain can produce extreme hypersaline soils (about 120 g of solutes per liter of soil solution), which are devoid of vascular plants (Iacobelli and Jefferies, 1991). In southern James Bay, the inland more saline marshes located beyond the reach of storm surges derive salt from groundwater desalinating marine argillaceous silts of the substrate (Price and Woo, 1988). In the mid- to high-Arctic zones, salt marshes open to the sea are generally poorly developed and infrequent. The coasts are affected by storm waves during the period of open seawater and commonly develop beach ridges, spits, and barrier beaches. On the isostatically rising lands, the beach ridges show various height and spacing. They alternate with interridge lows occupied by shallow lakes and ponds (Figure 8). Sparse vegetation grows in these interridge coastal wetlands, which may be described as true oases where they occur in the northernmost polar deserts. Wetlands with numerous interlaced channels and lakes separated by patterned grounds develop on Arctic deltas, particularly along the Beaufort Sea (such as those
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Figure 7 Effect of ice push and erosion on salt marshes. (a) Freshly developed ice push structures in high tidal flats, precursors to a more mature, pond-riddle marsh. (b) Mature salt marsh with jigsaw pattern that has developed because of formation of pools initiated by ice action.
Figure 8 Coastal ridges of Igloolik Island in northwest Foxe Basin, Nunuvut, in the mid-Arctic Zone of Canada.
of the Colville and Mackenzie rivers; Figure 9a) and along the Arctic Russian coast (such as those of the Lena River (Figure 9c,d) and Kolyma River). Pingos (small conical hills with a core of solid ice) develop in shallow lakes where the coastal areas are low lying, such as in the Mackenzie River delta (Figure 9b). Thermokarst greatly affects Arctic coastal zones, but the change in the landscape differs depending on rates of uplift relative to sea-level rise. For example, the coastal plain of southeast Foxe Basin, the Arctic Coastal Plain of Alaska, and the Mackenzie River delta are characterized by numerous thermokarst lakes. There sea-level rise outpaces any residual postglacial isostatic rebound, and during the ensuing marine transgressions, considerable coastal erosion occurs and the coastal lakes become breached and invaded by the sea (Figure 10c,d; Ruz et al., 1992; Wolfe et al.,
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Figure 9 Coastal wetlands associated with polar river deltas. (a,b) Interlaced channels, lakes (several are thaw lakes) in the Mackenzie River Delta, Northwest Territories, Canada [(a) from NASA and (b) aerial view of part of the delta with pingos developed in a coastal lake, from Canadian Geological Survey]. (c,d) Interlaced channels, thaw lakes, and patterned ground of the Lena River delta, Russia [(c) from NASA; (d) aerial view of well-developed patterned ground in interchannel areas, adapted fromWilliams (1994)].
1998). Local salt marshes can develop inside the newly formed, protected embayments. In the Foxe Basin, instead, the land is still undergoing isostatic uplift and the thaw lakes of the coastal tundra remain isolated and are only locally breached and joined by creeks (Figure 10a,b).
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Figure 10 Thermokarst structures in coastal zones. (a) Thaw lakes in the Arctic Coastal Plain, Cape Harlett, Alaska (adapted from Bowen, 2005). (b) Breached thaw lake Mackenzie River Delta (after Ruz et al, 1992). (c,d) Isolated lakes in Great Plain of the Koukdjuak, southeastern Foxe Basin, Nunavut, Canada; (c) satellite image of the entire plain; (d) air view of part of the plain.
6. V EGETATION OF P OLAR C OASTAL W ETLANDS Environmental conditions such as cold climate and icy conditions within the immediate coastal zones are often severe enough to restrict species richness. Furthermore, regional development portends a likely future scenario where some low-lying sedimentary coastlines and their associated biota become increasingly vulnerable to rise in sea level and the occurrence of tidal surges associated with global climate change. Other damaging oceanographic changes are likely to occur.
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For example, the present mean annual discharge of freshwater to the Arctic Ocean and the Hudson Bay is estimated at 5,250 km3/year (Shiklomanov et al., 2000), and significant increases (up to 30%) in discharge rates are projected to occur in response to climatic warming (Walsh et al., 2005). These changes in runoff will affect water levels, salinity, and nutrient fluxes in estuarine wetlands, all of which may be expected to alter biological production and biodiversity (Walsh et al., 2005). The extent to which the Arctic biota present at this interface between terrestrial and marine systems can adapt (genetic response) or modify behavior patterns (phenotypic plasticity) to this ongoing change is largely unknown. Sea ice persists until late spring (mid-June) on many shores, which restricts plant growth in the early growing season. At that time, most sea ice is located at the seaward end of marshes and its presence may protect the lower marsh vegetation from grubbing by geese (Section 7.2.1). Ice rafting is common when the ice breaks free of the shore at melt. The underlying sediment and vegetation are still frozen to the base of the ice and are carried to different locations in melt or tidal water. The remaining exposed sediment may be colonized by inward clonal growth from adjacent intact graminoid swards. The only studies of the nutritional status of Arctic salt-marsh soils are those conducted in the southern Hudson Bay region, where the results indicate that the soils are severely nitrogen-limited for plant growth (Cargill and Jefferies, 1984; Ngai and Jefferies, 2004). Addition of ammonium or nitrate salts leads to a rapid increase in the aboveground biomass in summer, but quickly growth becomes phosphorus-limited because of a colimitation of this element when nitrogen shortage is alleviated (Cargill and Jefferies, 1984). For the immediate coastal freshwater mires, mostly poor fens, plant growth is limited by both nitrogen and phosphorus (Ngai and Jefferies, 2004). The plant species richness of Arctic and sub-Arctic coastal salt marshes is low compared to temperate marshes, and prostrate graminoids dominate the vegetation. The common vascular species that colonizes suitable sites along low-lying, muddy seashores throughout circumpolar regions is Puccinellia phryganodes (Figure 11a; Hulte´n, 1968). In Arctic North America, plants are triploid and although they flower, seeds are not produced. In northern Fennoscandinavia and in the White Sea region of the Russian Federation, there are reports of tetraploid races of this grass, but it is not known if seed set occurs (R.M.M. Crawford, personal communication). Hence, at least in North America and northern Russia, plants are dispersed by clonal propagation. Individual leaves, shoots, and tillers are able to establish in soft sediment and develop into plants (Chou et al., 1992). Individuals are extremely resilient to environmental stressors; they can survive encased in pack ice for months and have been grown successfully in an anaerobic jar in an atmosphere of nitrogen for a number of weeks (Crawford et al., 1994; Crawford and Smith, 1997). Another widespread circumpolar species is Carex subspathacea (Figure 11b) and this includes the closely related species, Carex salina and Carex ramenskii, which may be variants of C. subspathacea. The species appears to be less salt tolerant than P. phryganodes and tends to occur in areas that receive fresh or brackish drainage water from adjacent lowlands. Unlike the grass, C. subspathacea can grow in anoxic soils where drainage is impeded. Seed set is episodic and most growth occurs via clonal reproduction. Both graminoids have well-developed
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Figure 11 Typical plants of Arctic and sub-Arctic coastal marshes of James and Hudson bays. (a) P. phryganodes colonizing plants of salt marsh (inset: inflorescence from Aiken et al., 2003; Dalwitz et al, 2003). (b) C. subspathacea (adapted from Aiken et al., 2003; Dalwitz et al., 2003). (c) H. tetraphilla brackish marsh. (d) Thick silt deposits trapped by H. teraphilla.
rhizomatous and/or stoloniferous systems and the fine root systems are confined to the upper soil layers (<7.5 cm). Other species that are common include Triglochin palustris and Triglochin maritima (the former grows more seaward than the latter), Cochlearia officinalis, Plantago eriopoda, Potentilla egedii, Ranunculus cymbalaria, Stellaria humifusa, Carex ursina, Carex maritima, and Festuca rubra (Kershaw, 1976; Jefferies, 1977; Jefferies et al., 1979). All flower infrequently but seed set is uncommon and depends on prevailing local weather conditions. Weather conditions are also important in the previous year when flower buds are laid down in most species. There have been few studies of the seed bank in Arctic and sub-Arctic salt marshes (Staniforth et al., 1998;
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Chang et al., 2001). The results show that the composition of the vegetation and the soil seed banks are only loosely correlated (approximately 50%), reflecting the poor contribution of the dominant graminoid species mentioned above to the seed bank. Some species are overrepresented in the soil seed bank compared with their abundance in the vegetation. They are weedy species typical of degraded or disturbed soils. Seed banks in impacted and fragmented sites do not recover quickly (Chang et al., 2001). The annual growth habit is confined to the low Arctic, generally close to the Arctic/sub-Arctic boundary. Within salt marshes, three species are represented: Salicornia borealis, Koenigia islandica, and Atriplex patula, which mostly grow on an organic substrate in the upper levels of salt marshes or in supratidal marshes (flooded by seawater only during storm surges) but only set seed in favorable years. S. borealis does not have a well-developed long-term seed bank, and annuals are not the primary colonizers of exposed mudflats, as in temperate salt marshes. This functional group of annuals is a prime candidate to study the response of plants in the low Arctic to climate change. There are indications that S. borealis is spreading north on the Cape Churchill Peninsula, Manitoba in the last decade, but much of this spread may be related to the loss of vegetation and exposure of sediments as a result of goose grubbing (see below). Where salt marshes grade into beach ridges or dunes, Leymus mollis var. arenarius is common. This ecological equivalent of marram grass (Ammophila arenaria) in temperate latitudes is widespread in well-drained, disturbed, sandy habitats in the Arctic. Other species that occur in this type of habitat where there is some soil organic matter include F. rubra and the related species, such as Sedum rosea, Parnassus palustris, Primula stricta, Bartsia alpina, Polygonum vivipara, and Chrysanthemum articum. Because of the large outflow of freshwater from rivers, brackish conditions often prevail in the tidal reaches of the estuaries and on open marine coasts close to the river mouths. The salinity ranges from about 3 g of solutes per liter up to 12 g/L. The range is the result of the movement of the tidal salt wedge along the lower reaches of the rivers. Because of the high rates of sedimentation in some estuaries associated with the deposition of the sediment load from rivers and the brackish conditions that prevail, soft sediments are often available for colonization by plants intolerant of full salinity. In these muddy estuaries, species such as Hippuris tetraphylla, Hippuris vulgaris (less salt tolerant), T. palustris, Potamogeton filiformis, P. pectinatus, C. officinalis, and R. cymbalaria readily establish in soft sediment or on shallow river bottoms if the outflow is not rapid. This flora is not confined to the vicinity of river mouths, but can occur in coastal areas beyond the intertidal zone where relict salt or brackish ponds are present as a result of isostatic uplift. Hippuris species develop an extensive rhizome system, and the stands of shoots produced by clonal propagation are very effective at trapping soft sediment, leading to a rapid change in coastal topography (Figure 11c,d).
7. FAUNA OF POLAR C OASTAL W ETLANDS Although species diversity in the Arctic is considered relatively low compared to other regions, the Arctic terrestrial fauna nevertheless contains some 6,000 species, which is about 2% of the global total (Chernov, 1995; Callaghan et al.,
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2005). From an ecological point of view, the fauna of polar coastal wetlands can be divided into two basic categories: an infauna consisting of organisms that live their life cycles within or in close association with the wetland; and a second group, an exfauna that makes use of the resources within the wetland on a seasonal basis, and whose members are migratory or nomadic. Polar coastal wetlands, in fact, play a critical role in the life cycles of many migratory animals, particularly birds; many species of waterfowl and shorebirds use Arctic and sub-Arctic wetlands both for breeding and/or as stopovers on migration to reach their breeding areas. The flora and infauna of the wetlands provide the food resources upon which the birds depend to complete their annual cycles.
7.1. Invertebrate fauna Among invertebrates, primitive groups are better represented (such as springtails: 400 species, 6% of the global total of species) than advanced groups (such as spiders: 300 species, 0.1% of the global total). Typically, there is a reduction in invertebrate species and families with increasing latitude, and the distribution of some groups, such as spiders, is patchy (Chernov, 1995; Pickavance, 2006). The common invertebrate species in the far north tend to be widely distributed and only a few species may become dominant at high latitudes (e.g., 12 species of springtail in the northern Taimyr, Russia; Chernov and Matveyeva, 1997). Although there has been a steady increase in the description of invertebrate assemblages at site-specific locations within the Arctic, the roles of individual species and functional groups in community dynamics are poorly understood. Coastal wetlands and their fauna fall into two broad categories: those at or near the shore involving marine or saltwater-influenced habitats, and those occurring slightly inland involving mostly brackish and freshwater habitats. 7.1.1. Invertebrate fauna of coastal saline areas Marine intertidal invertebrates and other organisms have received limited study in many Arctic and sub-Arctic areas (although they are highly important food sources for a variety of birds, fish, and even mammals). The extent of intertidal areas that develops in different regions will depend on the geomorphology and sedimentary characteristics of the area, and intertidal organisms not only have to survive severe climatic conditions, but are likely to be subjected to annual removal or disruption caused by ice scour, which can affect the flats themselves and the near-shore waters to depths of up to 5 m. In James and Hudson bays, for instance, the bivalve Macoma balthica is the most common burrowing mollusk and much of the intertidal stock is removed by ice and wave action during the winter and spring but is replenished by spatfall (larval production) originating from subtidal populations each year. M. balthica has an extensive geographic range, inhabiting temperate to Arctic coastal waters in the North Atlantic and North Pacific oceans, and forms a prominent food resource for birds and fish in intertidal areas in Iceland, Hudson and James bays, and Alaska. Densities in James Bay average 2,000–3,700 individuals/m2, with highest recorded densities in zones of eelgrass Zostera marina of up to 12,800 individuals/m2
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(Martini and Morrison, 1987). These densities are consistent with those recorded in other areas: Alaska, maximum 4,000 individuals/m2 (Powers et al., 2002); St Lawrence Estuary (southeast Canada), maximum 2,700 individuals/m2 (Azouzi et al., 2002); and Dutch Wadden Sea, 3,250 individuals/m2 (Piersma and Koolhaas, 1997). Sub-Arctic tidal flats may be characterized as having fairly high densities of infauna invertebrates, but low species diversity. In James Bay, the intertidal fauna consist principally of M. balthica and the gastropod Hydrobia minuta (Martini et al., 1980), while on the Copper River Delta, Alaska, M. balthica, the amphipod Corophium salmonis, and the polychaete Eteone longa account for over 95% of individuals identified on the mudflats (Figure 12; Powers et al., 2002). Other organisms occurring in James Bay and other sites include gastropods, mussels, limpets, nematodes, oligochaetes, and polychaetes, as well as foraminifera, copepods, ostracods, amphipods, cladocerans, ectoprocta, and barnacles (Martini et al., 1980). Oligochaete worms and Dipteran larvae are numerous along the edge of the short-grass salt marsh (consisting primarily of P. phryganodes) and are important food items for shorebirds. Oligochaetes (family Naididae, genus Paranais) numerically account for about 63% of the macrobenthos in the salt marshes, and their distribution is strongly correlated with electrical conductivity and the organic carbon content of the sediments. In the coastal ponds, Dipteran larvae of the families Chironomidae, Heleidae, and Tipulidae occur in densities of up to 5,500 m–2 (Clarke, 1980). Intertidal mudflats in high Arctic locations have received less study. It would appear that numbers and variety of organisms are low, as environmental conditions are correspondingly more severe than in sub-Arctic localities. At Zackenberg in central northeast Greenland, coastal mudflats contain low to moderate densities of nematodes, tardigrades, and crustaceans, and these are preyed on by shorebirds during July and August (Caning and Rausch, 2001; Meltofte and Lahrman, 2006). Red knots (Calidris canutus) have been observed feeding on crustaceans on intertidal mudflats on the central east coast of Ellesmere Island, Canada, during the postbreeding period, and both knots and ruddy turnstones (Arenaria interpres) feed on crustaceans along shorelines in the Alert area on northeast Ellesmere Island, Canada. At the other end of the globe, there has been little study of intertidal areas in subAntarctic wetlands and mudflats. The Atlantic mainland coast of Tierra del Fuego, although surrounded by the sub-Antarctic oceanographic zone, is cool temperate in terms of climate and vegetation, and it does contain important coastal wetlands. At Bahı´a Lomas, in the Chilean sector of Tierra del Fuego near the entrance to the Strait of Magellan, preliminary investigations of the infauna show that polychaetes, bivalve mollusks, isopods, and amphipods are predominant in sandy habitats, their abundance dependent on sediment size and type (Ponce et al., 2003). 7.1.2. Invertebrate fauna of near-coast, freshwater areas In areas of inundated less regularly by the tide and farther inland, a rich insect fauna develops. In southern James Bay, Kakonge et al. (1979) identified 318 species of invertebrates (105 families, 14 orders) among which mosquitoes and biting flies were a prominent component. The insects play an important role in ecological processes in the marsh, including contributing to soil fertility through
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Figure 12 Locations of wildlife places mentioned in the text in the Arctic and sub-Arctic. (AL, Alert; AN, Arctic National Wildlife Reserve, Alaska; BI, Bylot Island; CH, Churchill, Manitoba; CP, Chukotka Peninsula; CR, Copper River Delta, Alaska; DS, Dewey Soper Migratory Bird Sanctuary, Baffin Island; EI, Ellesmere Island; FB, Foxe Basin; HB, Hudson Bay; JB, James Bay; LR, Lena River Delta, Russia; MR, McConnell River Migratory Bird Sanctuary; NL, Nelson Lagoon, Alaska; QM, Queen Maud Gulf Migratory Bird Sanctuary; RL, Rasmussen Lowlands; SI, Southampton Island;WI,Wrangel Island;WS,Wadden Sea;YF, Yakutat Forelands, Alaska; YK, Yukon and Kuskokwim river deltas Alaska; ZA, Zackenberg, Greenland) (after UNEP/GRID-Arendal, 2005).
aeration and transfer of organic particles into the soil, litter breakdown (by springtails, mites, nematodes, rotifers, and some exotic earthworms), the role as major secondary producers, and the function of providing a food resource for migrating birds. Mosquitoes were reported to occur in densities of 5 million per acre (13.35 m/ha) on the coast of Hudson Bay (West, 1951). Chironomids often reach densities of many thousand per square meter in freshwater and brackish
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water (Pinder, 1983), and they are a major component of the macrobenthos in ponds at northern latitudes (Andersen, 1946; Butler et al., 1981) as well as being an important component of the diet of shorebirds (Summerhayes and Elton, 1923; Holmes and Pitelka, 1968) and waterfowl (Bergman and Derksen, 1977; Danell and Sjo¨berg, 1977). Arthropod species characteristic of the sub-Arctic and the Boreal forest are frequently transported on southerly winds to the low Arctic where they survive the summer (Danks, 1981).
7.2. Vertebrate fauna using coastal wetlands 7.2.1. Avifauna Birds form one of the most prominent components of the fauna using coastal wetlands. Most species may be categorized as waterbirds, principally waterfowl such as ducks, geese (12 breeding species), and swans, but also loons, shorebirds, gulls, and terns. Other types of birds including birds of prey (such as owls and raptors) and passerines also use coastal wetlands. Within the Arctic, 450 species of birds, which make up the majority of vertebrate species, have been recorded breeding (Callaghan et al., 2005); the majority are migratory and migrate in winter to southern latitudes and many inhabit coastal wetlands (Schmiegelow and Mo¨nkko¨nen, 2002). The total number of wetland birds that breed in the Arctic is estimated at between 85 and 100 million individuals (Callaghan et al., 2005). Polar wetlands not only serve as breeding grounds for many millions of waterbirds, but also play a key role as important migration stopover sites enabling the birds to travel between their breeding grounds and their more southerly wintering areas. The migration systems connecting southern wintering and Arctic breeding areas are generally known as flyways – worldwide there are some 8 recognizable flyway systems for waterfowl (Figure 13) and 10 (nine linking Arctic and Boreal breeding areas with wintering zones and one in South America linking austral
Figure 13
Major global flyway systems used by waterbirds (adapted from ACIA, 2004).
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sub-Antarctic, cool-temperate, breeding, and wintering areas) for shorebirds (Figure 14). Some birds travel between Arctic and sub-Arctic wetlands to subAntarctic wetlands during the course of their annual travels: the North American red knot (C. canutus rufa) (Figure 15a), for instance, migrates from breeding areas in the central Canadian Arctic, through areas along the coasts of Hudson and James bays, through temperate and tropical areas, to wintering areas on the cool-temperate, intertidal areas of Tierra del Fuego at the southern tip of South America (Morrison and Harrington, 1992; Morrison, 1984; Morrison and Ross, 1989). Areas supporting important concentrations of shorebirds worldwide are located near coastal regions of high productivity (Butler et al., 2001).
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Figure 14 Major global shorebird flyway systems linking arctic breeding wetlands with “wintering areas” some of which are sub-Antarctic wetlands (adapted from Piersma and Lindstro«m, 2004).
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Figure 15 Birds utilizing the polar coastal wetlands. (a) Red knots in flight. (b) Lesser snow geese, light-colored, flightless adults in the background during breeding period, and darker colored goslings in the foreground.
Bird breeding areas. Geese form a prominent component of the avifauna using polar coastal wetlands, with populations of six species totaling an estimated 5.7 million birds in the North American Arctic and nine species totaling an estimated 2.5 million birds in the Eurasian Arctic (Zo¨ckler, 1998) in the early 1990s. Some geese occupy a relatively restricted part of the Arctic, while others have a wide geographical distribution. For instance, about 95% of Ross’s geese (Chen rossii) historically nest in the Queen Maud Gulf Migratory Bird Sanctuary in the central Canadian Arctic (Kerbes, 1994; Ryder and Alisauskas, 1995; Kerbes et al., 2006), whereas lesser snow geese (Chen caerulescens caerulescens) breed from northwest Greenland, through the Canadian and Alaskan Arctic, to Wrangel Island and the Chukotka Peninsula in Russia (Mowbray et al., 2000) (Figures 12, 15b). In the North American Arctic, hundreds of thousands of geese and other waterfowl nest in areas such as the Yukon and Kuskokwim river deltas [which support most of the world’s emperor geese (Chen canagica)], Arctic National Wildlife Refuge in Alaska, the Rasmussen Lowlands, Queen Maud Gulf Migratory Bird Sanctuary, McConnell River Migratory Bird Sanctuary, and Dewey Soper Migratory Bird Sanctuary in the Canadian Arctic. Similar numbers are supported in the Eurasian Arctic in areas such as the Lena River delta in Russia (Gilg et al., 2000) and in sub-Arctic wetlands in Iceland (Rowell and Hearn, 2005) (Figure 12). Shorebirds also form a prominent component of the breeding avifauna of sub-Arctic and Arctic wetlands. Zo¨ckler (1998) reported 13 species of calidrid sandpipers with an estimated population of 8.1 million individuals that breed in the North American Arctic (including Greenland), while 17 species involving 6.3 million individuals breed in the Eurasian Arctic. When more southerly areas around Foxe Basin, Hudson Bay, and James Bay are included, some 27 species of shorebirds use central areas of Canada (Morrison and Gaston, 1986). Some species nest directly in wetland habitats, such as red phalarope (Phalaropus fulicarius) and dunlin (Calidris alpina), whereas others, such as ruddy turnstone and red knot, nest in nearby upland habitats. In the latter cases, the upland (tundra) habitats used
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for nesting are often found in close association with wetter habitats, where the birds feed and raise their young. Shorebirds breed right to the northern limit of land in North America and Eurasia, although it is in the more southerly large wetlands that the highest numbers and diversity are found, such as Yukon and Kuskokwim river deltas in Alaska, Rasmussen Lowlands in Canada, and coastal wetlands in eastern Siberia. Bird staging areas. During migration, sub-Arctic and Arctic coastal habitats support equally large populations of waterfowl and shorebirds en route to and from the breeding grounds. In many cases, geese and other waterfowl acquire nutrients that they bring to the breeding grounds farther north in the form of body stores, which are used to form eggs or enhance survival (Alisauskas and Ankney, 1992). A similar phenomenon occurs with shorebirds. In Iceland, for instance, red knots en route to the eastern Canadian high Arctic from European wintering quarters, not only accumulate large amounts of fat but also alter their physiological makeup, increasing the size of organs and muscles used for flying (pectoral muscles) and decreasing the size of organs and muscles that are less important during the flight (stomach, intestines, and leg muscles) so that they transform themselves into virtual flying machines (Piersma et al., 1999). Not all the stores are used during the flight, and an important function of the significant amounts of fat and protein that remain on arrival in the Arctic appears to be to enable the birds to retransform their physiological makeup into one suitable for breeding [such as liver, heart (decreased during flight), stomach, and intestines increase in size], or perhaps for survival if early season conditions are difficult (Morrison and Hobson, 2004; Morrison et al., 2005). These aspects of bird migration emphasize the interconnected nature of the Arctic, sub-Arctic, and other wetlands farther south. The ability of birds to acquire the needed stores during migration has important survival implications. Shorebirds departing Iceland in better than average condition were shown to have a higher survival when faced with difficult weather conditions in the Arctic (Morrison, 2006; Morrison et al., 2007), and conversely, shorebirds prevented from reaching adequate departure weights at the final spring stopover area in North America suffered significantly decreased survival (Baker et al., 2004). Waterfowl also depend on food resources in sub-Arctic wetlands during migration. Examples in North America include the Mid-Continent populations of lesser snow geese and Atlantic brant (Branta bernicla) passing northward through James Bay during the spring (Jefferies et al., 2003; Ward et al., 2005) and European populations of brant using Icelandic stopovers en route to the Canadian Arctic (Ward et al., 2005) as well as pink-footed geese (Anser brachyrhynchus) using wetlands in northern Norway en route to breeding grounds on Svalbard (Drent et al., 2006; Glahder et al., 2006). In recent decades, habitats in migration and breeding areas used by lesser snow geese and Ross’s geese in the Canadian Arctic have been heavily damaged as a result of overgrazing by the geese, which have undergone spectacular increases in population size. On the wintering grounds in southern latitudes, the loss of coastal habitat and freshwater wetlands (associated with changing land use) has led to the birds
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becoming heavily dependent on agricultural crops, particularly high-yielding crops rich in nitrogen. Some species that have shown geometric population growth, primarily in response to changes in modern agriculture, are the lesser snow goose and Ross’s goose in North America (Abraham et al., 1996, 2005) and the barnacle goose (Branta leucopsis) in Europe (Van Eerden et al., 2005; Jefferies et al., 2006a). Winter counts of the Mid-Continent population of snow geese increased in a geometric manner between 1970 and 2000, from 0.8 million in 1969 to 2.7 million in 1994, indicating that the entire population was most likely between 4.5 and 6 million in the mid-1990s (Abraham and Jefferies, 1997; Jefferies et al., 2003). At La Pe´ rouse Bay near Churchill on the Hudson Bay coast, the breeding colony increased from less than 2,000 pairs in 1968 to 44,500 pairs in 1997 (Jefferies et al., 2003). The main factors involved appear to be the increased availability of food from agriculture, as well as the availability of refugia from hunting, lower harvest rates, and possible climate change on the breeding grounds (Jefferies et al., 2003; Kerbes et al., 2006). The birds have, in effect, escaped from density dependence in the coastal marshes, and hunting losses (where hunting is permitted) have not been able to keep pace with the increases in the population sizes of the different species (Abraham et al., 1996). With numbers of lesser snow geese exceeding 5 million, the large population may be expected to have adverse effects on Arctic coastal vegetation, depending on the densities of the birds and their foraging behavior (grazing, grubbing, shoot pulling of sedges), which is related to bill size and shape. Grubbing, in which the geese pull up the roots and rhizomes of the plants, and shoot-pulling lead to destruction of the vegetation and often total loss of the selected graminoid plants. The resulting physical and chemical changes in the exposed sediments and the continued exposure to geese foraging alter habitat succession and recovery (Abraham et al., 2005). Coastal wetland sites in the sub-Arctic in North America are particularly vulnerable to such disturbance because they serve as sites for both staging and breeding. During both of these phases of the annual cycle, the birds need to feed heavily, especially in early spring, to regain resources expended during migration and maintain or increase reserves for breeding. In North America, the Mid-Continent population of the lesser snow goose that breeds in the eastern Canadian Arctic has had a dramatic impact on coastal wetland plant assemblages and soils at a large spatial scale (Jefferies et al., 2003): vegetation loss has been so extensive that it can be readily detected by remote sensing (Jano et al., 1998, Jefferies et al., 2003; Didiuk and Ferguson, 2005, Jefferies et al., 2006b; Alisauskas et al., 2006). The loss of vegetation is triggered by geese grubbing for roots and rhizomes in thawed ground in early spring. However, it is the subsequent abiotic changes, including the development of hypersalinity, loss of organic matter, and compaction of sediments that limit the potential for recolonization. This is compounded by biotic factors such as loss of the seed bank, the absence of sexual reproduction in P. phryganodes (at least in North America), and irregular seed set in C. subspathacea. Because of the cumulative impact of grubbing, patches of exposed sediment coalesce into larger and larger units of exposed sediment largely devoid of vegetation. Reestablishment of vegetation is long-term and requires erosion of
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hypersaline, consolidated sediment and the buildup of unconsolidated soft sediment, in which plants can establish themselves. A similar loss of vegetation is occurring in coastal freshwater marshes, although the abiotic and biotic processes are different (Jefferies et al., 2003; Alisauskas et al., 2006). Erosion of peat, following loss of vegetation, can lead to exposure of underlying glacial gravels and marine clays in coastal locations and can alter the trajectory of succession (Handa et al., 2002). For the greater snow goose (C. caerulescens atlantica) breeding in the northeastern Canadian Arctic, grazing pressure on graminoids during the summer is high and reduces the plant production in wetlands, although the vegetation has not been damaged past the point of recovery, as observed in areas used by lesser snow goose populations (Gauthier et al., 2006). Goose abundance on Bylot Island, one of the largest breeding colonies of the greater snow goose, was still at only half the estimated carrying capacity of the island’s wetlands in 1997. Similarly, the European population of the barnacle goose is considerably lower than that of the North American snow goose (Madsen et al., 1999), and they have had relatively little impact on coastal habitats. During autumn migration, many shorebirds and waterfowl use the coasts of Hudson and James bays as well as other locations in the eastern Canadian Arctic to build up body reserves for the flight south (Morrison and Harrington, 1979; Morrison and Gaston, 1986). Shorebird distribution in James and Hudson bays was directly related to food abundance for several species, including semipalmated sandpiper (Calidris pusilla), red knot, and hudsonian godwit (Limosa haemastica), a relationship that was evident at several geographical scales (locally across the marsh, over 15 km stretches of coast, and over several hundreds of kilometers) (Morrison, 1983, 1984; Morrison and Gaston, 1986). In northern Alaska, a number of species of shorebirds move from tundra to littoral habitats after breeding, and coastal flats are important during the autumn migration (Connors et al., 1979; Andres, 1994). In western Alaska, 17 species of shorebirds regularly use the intertidal flats of the Yukon and Kuskokwim river delta during spring and fall; peak counts reached some 300,000 birds, consisting mostly of dunlins, western sandpipers (Calidris mauri), and rock sandpipers (Calidris ptilocnemis), and an estimated total of 1–2 million shorebirds were thought to use the area each year (Gill and Handel, 1990). Many millions of shorebirds used the Copper River delta flats during spring migration (Islieb, 1979; Bishop et al., 2000) and hundreds of thousands used the Yakutat Forelands area (Andres and Browne, 1998) in Alaska. Some 20 species of shorebirds used intertidal habitats in Nelson Lagoon on the Alaska Peninsula during fall migration (Gill and Jorgensen, 1979). Such areas are of critical importance during southward migration of species such as the bar-tailed godwit (Limosa lapponica), which accumulates up to 55% of its body weight in fat and undergoes physiological changes involving reduction in gut sizes before a spectacular migration across the Pacific Ocean, which can involve nonstop flights of 11,000 km lasting 6 or more days to wintering areas in New Zealand and eastern Australia (Piersma and Gill, 1998; Gill et al., 2005; Gill et al., 2006). The critical importance of these coastal wetlands in the life cycles of a variety of birds is clear.
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7.2.2. Mammal fauna A variety of mammals occurs in polar coastal wetlands, including bears (polar bears, grizzly bears, and brown bears), Arctic and red foxes, wolves, wolverines, caribou, arctic hare, mink, weasels, lemmings, and voles. While less numerous than waterbirds, mammals play an important role in ecological dynamics of Arctic systems. Lemmings and other rodents undergo pronounced cyclical patterns in abundance, and the associated responses of predators such as arctic foxes (Alopex lagopus) in turn influences the success of other birds and animals. In years of high lemming abundance, for instance, Arctic foxes spend much of their time hunting the lemmings so that there is relatively little predation pressure on nesting birds; in contrast, in low lemming years, shorebirds and their nests may be heavily depredated by foxes. Breeding success of the birds on the breeding grounds can be affected to the extent that the lemming cycles can be detected by observing the number of shorebird young reaching migration and wintering areas (Underhill, 1987; Underhill et al., 1989; Blomqvist et al., 2002); pomarine jaegers (Stercorarius pomarinus), which depend on lemmings for food, may not breed at all in some areas in low lemming years (Maher, 1970). In the Canadian Arctic, lemming cycles may be less synchronized over large areas than in the Russian Arctic, although they are clearly evident in the high Arctic. Polar bears (Ursus maritimus) are a top predator in marine ecosystems and are highly dependent on the presence of sea ice which provides habitat for their principal prey, ringed seals (Phoca hispida), in many regions of the Arctic (Stirling et al., 1999, Derocher et al., 2004). During the summer when the sea ice melts, polar bears come ashore in coastal areas. During this period when the normal diet of seals is not available, they exist on accumulated fat reserves and become fairly omnivorous predators and scavengers, feeding on berries, seaweed, and adults, young, and eggs of colonial nesting seabirds and waterfowl [including thick-billed murres (Uria lomvia), little auks (Alle alle), gulls, geese, ducks, and sometimes more terrestrial species]. Polar bears are occasionally cannibalistic and have been also known to hunt large terrestrial mammals such as caribou and muskox (Ovibos moschatus) (Stempniewicz, 2006). Polar bears can cause extensive damage to colonies of seabirds nesting in coastal locations; at East Bay, Southampton Island, in 1997, two bears destroyed an entire common eider (Somateria mollisima) colony, eating an estimated 12,000 eggs, rendering themselves temporarily immobile in the process (H.G. Gilchrist, personal communication). The predicted thinning and reduced coverage of Arctic sea ice resulting from climate warming are likely to substantially alter sea ice ecosystems (Loeng et al., 2005) and could result in deleterious effects on availability of food sources for polar bears. These predictions reflect the strong coupling between marine and terrestrial systems in response to climate change. Such effects are likely to be most evident at the southern distribution limit of polar bears, where early melt and late freezing of sea ice extend the period when the bears are on land, during which time little feeding occurs. Recently, the condition of adult bears has declined in the southwest region of the Hudson Bay and, the number of first-year cubs as a proportion of the population has fallen associated with the early breakup of sea ice and the cubs coming ashore in poor condition (Stirling et al., 1999; Derocher et al., 2004).
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Polar bears are increasingly likely to seek alternative food sources as the extent of sea ice declines and inshore time increases under climate warming (Stirling and Parkinson, 2006), although it is not clear to what extent individuals may be able to modify their feeding habits to utilize new food sources. For example, unpublished data (Ian Stirling, per.com.) indicate harbor seals (Phoca vitulina), which occur on rocky coasts, may be increasing in western Hudson Bay, possibly in response to climate warming and that this species is becoming more important in the diet of polar bears. In North America, caribou (Rangifer tarandus) herds use coastal wetlands and plains for both calving and wintering; in the summer, coastal areas are important for avoiding predators and biting and parasitic insects. Furthermore, caribou diets shift in summer from the lichen dominated winter diets to vascular plants, including wetland sedges, grasses, and other species.
8. ENVIRONMENTAL H AZARDS Most polar coastal wetlands have been subject to few direct and indirect anthropogenic influences compared to their temperate counterparts. They are, however, environments that are easily impacted and modified by either natural disturbances or human activities. Climate changes will affect these wetlands directly, as a result of rises in temperature and in amounts of precipitation that will affect growth of vegetation and the reproductive success of plant populations. There are also a plethora of indirect effects associated with the melting of sea ice and permafrost and changes in salinity and hydrology. For example, low-lying coasts, where wetlands occur, are vulnerable to the deleterious effects of the increased incidence of storm surges and the destructive effects of wave action in the absence of sea ice (Callaghan et al., 2005; Loeng et al., 2005; Cahoon et al., 2006). The direct effect of climate warming on melting and the changing regime of sea ice will strongly affect animal behavior, possibly leading to extinction of specialized species like polar bears in some areas of the Arctic (Derocher et al., 2004). Human activities may impact drastically on polar coastal wetlands because of the rapidity of the imposed changes. Rapid adverse effects are associated with megahydrological projects, resource extraction (hydrocarbon exploration and exploitation, mineral mining for lead zinc, gold, and diamonds), ecotourism, fishing, and increased hunting and gathering by indigenous people (Anisimov et al., 2001). Large-scale hydrocarbon exploration and production is ongoing in areas onand offshore in the Arctic Coastal Plain of Alaska, the Mackenzie River delta, the Pechora Basin, the Lower Ob Basin, and the Western Siberian Plain. These activities adversely impact wetland ecosystems and their wildlife because of the necessary infrastructure and the disturbances created at the different locations. A more ominous impact is related to the pollution of the whole coastal and marine environment that can derive from oil spills during exploration and the marine transport of petroleum through a possible ice-free Northwest or Northeast Passage
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(Anisimov et al., 2001). Both heavy metals and POPs (Persistent Organic Pollutants) are transported by water and air, and both bioaccumulate in trophic food webs and in wetland soils, thereby posing environmental risks to wildlife and human populations (AMAP, 2002).
9. C ONCLUSIONS AND RESEARCH PRIORITIES Earth is continuously changing. During the Quaternary Era, continent-wide glaciers developed and waned several times responding to alternating cold and warm periods. The biota, including human populations, adapted to these changes by migrating and recolonizing land affected by ice. However, humans are now capable of affecting rates of climatic and geochemical (pollution) changes that already are impacting polar regions, and will continue to do so. With amplification of temperature, the global warming will be more evident in the Arctic, and it can be expected to reduce the cover of both sea and land ice. As a result of changes in albedo and heat storage with the melting of ice, the redistribution of heat at the global scale can modify atmospheric properties and oceanic currents. This will probably lead to adverse consequences such as changes in atmospheric precipitation, increased frequency of storms, and a global sea-level rise. Coastal Arctic and sub-Arctic environments and the associated biota are particularly vulnerable to these climatic changes. The majority of human settlements in the Arctic are located on low-lying coasts and they will also be adversely affected by storms and tidal surges. The melting of the sea ice and the opening of Arctic sea routes present further possible hazards for low-lying coastal regions and their biota. Pollution, physical disturbance, and increased access to these remote localities are likely to result in indirect changes, many of which are unforeseen at this stage. Research priorities include the following. 1. Monitor rates of permafrost loss, and, in particular, map the ongoing northward shift of the boundary between the continuous and discontinuous permafrost zones (Tarnocai, 2006). This change is likely to lead to the partial drying of vast peatlands and the release of greenhouse gasses (carbon dioxide and methane) to the atmosphere and, thus, an increase in the rate of global warming. In addition, shrubs and grasses are likely to replace the wetland flora of aquatic plants and sedges. Such changes will affect the existing insect fauna and the availability of suitable nesting and feeding sites for migratory waterfowl and passerines that nest in the Arctic. 2. Improve and expand the determination of rates of isostatic land uplift and sealevel rise using a network of stations, in order to predict relative sea-level changes along coasts. This is especially important where severe coastal erosion is taking place. 3. Monitor the effects of global change on coastal environments and the ability of resident wildlife populations (such as vegetation, polar bears, foxes, and lemmings) and migratory populations (most bird species, caribou) to adjust to these changes. Recording of invasive species in northern latitudes associated
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with climate change is necessary. Increased attention needs to be paid to traditional knowledge and its contribution to our understanding of past and present changes in wildlife populations. More direct involvement of First Nations’ people in research activities is needed. 4. Expand monitoring to understand the impact of local (oil exploration and production, mining) and distant (long-range) transport of contaminants on human activities (fishing, hunting on land and on sea ice, societal change). Migratory birds, and waterfowl in particular, constitute a means of transport of geochemical materials from the industrial, agricultural south to the northern coasts.
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INTERTIDAL ECO-GEOMORPHOLOGICAL DYNAMICS AND HYDRODYNAMIC CIRCULATION Andrea D’Alpaos, Stefano Lanzoni, Andrea Rinaldo, and Marco Marani
Contents 1. Introduction 2. Intertidal Eco-Geomorphological Evolution 2.1. Poisson hydrodynamic model 2.2. Model of channel network early development 2.3. Model of marsh platform evolution 3. Results 4. Discussion 5. Conclusions Acknowledgments References
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1. INTRODUCTION The strong dynamic coupling of intertidal platforms, marsh vegetation, and tidal channel networks gives rise to a complex system, whose nonlinear dynamics is arguably one of the most fascinating examples of eco-morphodynamics: The collective temporal evolution emerging from the mutual interactions among hydrodynamic, morphological, and biological processes. Improving our understanding of the chief land-forming processes, which drive intertidal system morphogenesis and long-term evolution, is an intriguing problem and a critical step to preserve such delicate systems, exposed to the effects of climate changes and human interference. Wetland vegetation ecosystems host an extremely high biodiversity, exhibit one of the highest rates of primary production in the world, and play a fundamental role in determining the evolution of coastal lagoons and estuaries (Mitsch and Gosselink, 2000; Marani et al., 2006b). The decline of wetland areas worldwide and their potential sensitivity to abrupt sea-level fluctuations highlight their global importance and call for a deeper understanding of their dynamics. This has motivated several researchers who have produced a large literature, especially in the last two decades (e.g., see Allen, 2000; Friedrichs and Coastal Wetlands: An Integrated Ecosystem Approach
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Perry, 2001; Marani et al., 2006a). Most existing works, however, concentrate on specific aspects of intertidal dynamics, such as tidal propagation in estuaries and tidal channel hydrodynamics (e.g., Le Blond, 1978; Jay, 1991; Friedrichs and Aubrey, 1994; Lanzoni and Seminara, 1998; Savenije, 2001; Savenije and Veling, 2005); tidal asymmetries and sediment dynamics in tidal channels (e.g., Pethick, 1980; Boon and Byrne, 1981; French and Stoddart, 1992; Friedrichs, 1995; Schuttelaars and de Swart, 2000; Lanzoni and Seminara, 2002); morphometric analyses of tidal networks (e.g., Myrick and Leopold, 1963; Pestrong, 1965; Steel and Pye, 1997; Fagherazzi et al., 1999; Rinaldo et al., 1999a,b; Marani et al., 2002, 2003; Rinaldo et al., 2004; Feola et al., 2005; Marani et al., 2006b); sedimentation and accretion patterns over vegetated marsh platforms (e.g., Stoddart et al., 1989; French and Spencer, 1993; Leonard and Luther, 1995; Ward et al., 1998; Christiansen et al., 2000; Leonard and Reed, 2002); salt marsh ecological dynamics and patterns (e.g., Adam, 1990; Yallop et al., 1994; Marani et al., 2004; Silvestri and Marani, 2004; Belluco et al., 2006); saturated and unsaturated subsurface flows in salt marshes and their relationships with vegetation patterns (Ursino et al., 2004; Marani et al., 2005, 2006a); and the influence of wind waves on the hydrodynamics of shallow tidal areas (Carniello et al., 2005). Even though significant advances have been achieved in all these fields, the understanding of the collective eco-morphological behavior of intertidal systems still lacks a comprehensive and predictive theory, due to the strongly intertwined interactions of their physical and ecological components. A deeper understanding may thus be achieved only by elucidating the detailed feedbacks between ecological and geomorphological processes, which, in turn, requires a holistic approach encompassing the governing bio-morphological processes over the wide range of spatial scales involved (Rinaldo et al., 1999a,b; Marani et al., 2003, 2006b). In order to arrive at a mathematical description explicitly including intertidal biotic and abiotic processes, it is useful to provide a brief review of some modeling results concerning the different components of the system. A number of zero-dimensional models have been proposed to investigate the long-term vertical growth of salt marshes by assuming their accretion rate as a function of sediment supply and either marsh elevation or biomass (e.g., Randerson, 1979; Krone, 1987; French, 1993; Allen, 1995, 1997; Morris et al., 2002; Temmerman et al., 2003). These models consider the evolution of a salt marsh point as a representative of the whole platform and, although providing helpful insights into the response of the marsh surface to tidal forcing and sea-level variations, they are unable to represent important space-dependent features. The modeling of the differential accretion of the marsh surface induced by the spatial variability of sediment deposition rates has been relatively attempted only recently in a one-dimensional setting (e.g., Woolnough et al., 1995) and with the incorporation of some form of vegetation dynamics (Mudd et al., 2004). Three-dimensional analyses of sedimentation patterns in tidal marsh landscapes have been carried out both in the very short period (single inundation event) through complete hydrodynamic models (Temmerman et al., 2005) and in view of a long-term evolution through simplified process-oriented models (D’Alpaos et al., 2007a), thus emphasizing the strong control exerted by ecological processes
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on marsh morphodynamics. The purely geomorphological equilibria of unchanneled subtidal areas have been recently studied through conceptual (Fagherazzi et al., 2006) and numerical modeling (Defina et al., 2007). Marani et al. (2007, 2008) analyzed the fully coupled dynamics of landforms and biota in the intertidal zone, through a model of the coupled tidal physical and biological processes. They proved the existence of multiple equilibria, and transitions among them, governed by vegetation type, disturbances of the benthic biofilm, sediment availability and marine transgressions, or regressions, thus emphasizing the importance of the coupling between biological and sediment transport processes in determining the evolution of a tidal system as a whole. In spite of the fundamental control exerted by tidal channels on the hydrodynamics and sediment dynamics within intertidal systems, and their importance for nutrient circulations within intertidal habitats, the literature on the morphogenesis and long-term morphological evolution of tidal channel networks is not as well developed. Field and laboratory observations (see, e.g., Pestrong, 1965; Redfield, 1965, 1972; Tambroni et al., 2005) and conceptual models (e.g., Yapp et al., 1916; Beeftink, 1966; French and Stoddart, 1992; Allen, 2000) have, however, been developed. In the last few years, mathematical and numerical models of the morphogenesis and long-term morphological evolution of tidal channels have also been proposed. Schuttelaars and de Swart (2000) and Lanzoni and Seminara (2002) developed, within different theoretical frameworks, one-dimensional models that allow the investigation of the equilibrium configurations of estuaries and tidal channels. In particular, Lanzoni and Seminara (2002) observed that equilibrium configurations, allowing a vanishing net along-channel sediment flux, tend to be reached asymptotically. Fagherazzi and Furbish (2001) analyzed the long-term morphodynamic evolution of a reference cross-section composed by an incipient channel zone and a marsh surface zone, through a model simulating aspects of initial channel formation over an existing tidal flat. D’Alpaos et al. (2006) extended the analysis of Fagherazzi and Furbish (2001) tracking the channel crosssectional morphodynamic evolution coupled with the vertical growth of the adjacent emerging marsh platform, with particular emphasis on the role played by the hydroperiod and halophytic vegetation. They found that channel cross-sections tend to adapt quite rapidly to changes in the flow. The morphogenesis and longterm evolution of channel networks have been recently studied through the use of simplified, process-oriented models (Fagherazzi and Sun, 2004; D’Alpaos et al., 2005; Kirwan and Murray, 2007) based on the Poisson hydrodynamic model proposed by Rinaldo et al. (1999a,b). Fagherazzi and Sun (2004) developed a stochastic model for channel network formation in which water surface gradients drive the process of network incision. D’Alpaos et al. (2005) set up a mathematical model of tidal network ontogeny describing channel initiation and progressive headward extension through the carving of incised cross-sections where the local shear stress – controlled by water surface gradients – exceeds a predefined, possibly site-dependent, threshold value. In agreement with observational evidence and conceptual models of marsh evolution, these approaches decouple the initial channel formation from the evolution of the adjacent marsh platform (Steers, 1960; Pestrong, 1965; French and Stoddart, 1992). However, contrary to the
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model proposed by Fagherazzi and Sun (2004), D’Alpaos et al. (2005) account for feedbacks existing between channel geometry and local hydrodynamic conditions, instantaneously adapting network configuration to the local discharge (or to the local tidal prism), in accordance with observational evidence and modeling (Friedrichs, 1995; Rinaldo et al., 1999b; Lanzoni and Seminara, 2002; D’Alpaos et al., 2006). Moreover, D’Alpaos et al. (2007b) have recently tested the channel network model by simulating the rapid development of small creek networks within a newly constructed artificial salt marsh in the Venice Lagoon. They showed that the synthetic creeks tend to originate at locations that match those of the actual ones, thus supporting the assumption of the strong control exerted by the water surface elevation gradients in the process of channel incision. On the other hand, Kirwan and Murray (2007) proposed a model of the long-term evolution of channel networks through a simplified treatment of flow, sediment dynamics, and vegetation productivity. Water routing across the marsh platform is again based on the local gradients of a Poisson-parameterized surface (Rinaldo et al., 1999a), but part of the procedure used to represent channel erosion seems somewhat artificial. Marciano et al. (2005) used the Delft3D hydrodynamic and sediment transport model to produce channel patterns in a short tidal basin. The results seem to be strongly influenced by the initial conditions specified, and only when an initial bottom configuration close to the expected equilibrium basin hypsometry is assigned, the model produces well-developed branching structures. Moreover, model validation is not conclusive in comparing generated and observed structures as it is carried out on the basis of Horton’s hierarchical analysis, a formalism shown to be unable to discriminate different network statistics (Kirchner, 1993; Rinaldo et al., 1998). D’Alpaos et al. (2007a) have recently discussed the interplay of erosion, sedimentation, and vegetation dynamics and their effects on the intertwined eco-morphodynamic processes governing the evolution of the marsh platform and of the tidal channels cutting through it. Temmerman et al. (2007) developed a coupled morphodynamic and plant growth model, simulating plant colonization and tidal channel formation on an initially bare flat marsh surface. The interaction of different biotic and abiotic processes in particular environments (Bahı´a Blanca Estuary) has also been recently addressed (Perillo et al., 2005; Minkoff et al., 2006). A simplified cellular automaton model for the development of tidal creeks, accounting for observed bioturbation effects linked to crab–halophytic plant interactions, shows that, in such particular setting, this interaction exerts a relevant role in driving the development of tidal creeks, overcoming the role of water surface gradients. Hood (2006) suggested that in the particular environment represented by a rapidly prograding delta dominated by river discharge, tidal channels might be the result of depositional rather then erosional processes. It is worth at this point to remark that most of the contributions to the modeling of the morphogenesis and evolution of channel marsh systems discussed above do not rigorously address the problem of model validation. Very seldom a quantitative validation of models against observed morphologies is attempted, and the evaluation of model results is rather performed by qualitative visual appraisal or on the basis of lenient geomorphic measures (e.g., the traditional Hortonian measures
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considered by Marciano et al., 2005). Different from this common approach, D’Alpaos et al. (2005, 2007a,b) defined a distinctive network statistics, which they use for a quantitative validation of model results, showing that the synthetic network structures generated by the model indeed reproduced several observed characteristics of geomorphic relevance such as, among others, unchanneled length distributions (Marani et al., 2003). In the following, we describe a comprehensive theoretical framework aimed at extending our current understanding of the coupled eco-geomorphic evolution of intertidal environments, and our abilities to model it quantitatively, as defined by the literature discussed.
2. INTERTIDAL ECO-GEOMORPHOLOGICAL EVOLUTION The chief morphological processes involved in the evolution of an intertidal area include the incision and subsequent elaboration of a channel network within a platform that may be evolving from a tidal-flat to a salt marsh state. As mentioned before, the interaction between these processes is also coupled through the influence of biotic processes, such as vegetation or microphytobenthos colonization, affecting the sediment transport and stability. Tidal channel initiation can be ascribed to the concentration of tidal fluxes over a surface, for example a mudflat, possibly induced by the presence of small perturbations in the topography. Such perturbations produce local scour as a consequence of the excess shear stress exerted at the bottom. Channel incision favors a further flux concentration, generating a positivefeedback mechanism that leads to the development of the observed tidal patterns (Yapp et al., 1916, 1917; Beeftink, 1966; French and Stoddart, 1992; Allen, 1997, 2000; Steel and Pye, 1997; Fagherazzi and Furbish, 2001; D’Alpaos et al., 2006). It is generally agreed that the process of network incision is a rather rapid one (Steers, 1960; Pestrong, 1965; Pethick, 1969; Collins et al., 1987; French and Stoddart, 1992; Allen, 2000): a permanent imprinting is likely to be given to the tidal environment, possibly later followed by a slower elaboration of the network structure, for example, by meandering and by the adjustment of channel geometry to variations in the local tidal prism due to the vertical accretion of the flanking intertidal surface (Gabet, 1998; Marani et al., 2002). The transformation of a tidal flat into a salt marsh requires the presence of critical conditions like, for example, a sufficient sediment supply, and relatively low tidal and wind–wave energy conditions. As soon as the local elevation of the platform reaches a height suitable for halophytic plant development, the surface is colonized by vegetation, which promotes the sediment trapping during submersion periods (Leonard and Luther, 1995; Christiansen et al., 2000) and contributes organic material (e.g., Randerson, 1979; Leonard and Reed, 2002; Morris et al., 2002). When vegetation extensively encroaches the marsh surface, the increased drag caused by plants influences tidal velocity profiles and the rate at which water floods into and drains from the platform adjacent to a channel (an increasing
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function of plant density, e.g., see Leonard and Luther, 1995; Nepf, 1999; Lopez and Garcia, 2001). The presence of vegetation also influences the planimetric evolution of tidal channels due to its stabilizing effects on surface sediments and channel banks (Garofalo, 1980; Marani et al., 2002). The influence of benthic fauna on erosion/deposition processes and on sediment characteristics through bioturbation and biodeposition has been observed as well (Yallop et al., 1994; Wood and Widdows, 2002). It is worthwhile emphasizing that observational evidence and modeling support the concept of inheritance of the major features of channelized patterns from sand flat or mudflat to a salt marsh (e.g., Allen, 2000; Friedrichs and Perry, 2001; Marani et al., 2003). As discussed above, when the tidal landscape reaches an elevation that allows the colonization by halophytic vegetation, this freezes the configuration of the network that can, from then on, only undergo minor changes immaterial to its basic structure. A numerically feasible description of such complex interactions, particularly in the context of a long-term model, requires the formulation of simplified model components retaining the most relevant features of the governing processes. Such a description of the key hydrodynamic properties of the flow over an intertidal platform is provided by the model proposed by Rinaldo et al. (1999a,b).
2.1. Poisson hydrodynamic model Under the assumption that a balance holds in the momentum equations between water surface slope and the linearized friction term, Rinaldo et al. (1999a) suitably simplified the two-dimensional shallow water equations to a Poisson equation H2 1 =
@0 2 ð0 z0 Þ @t
ð1Þ
where 1(x; t) is the local deviation of the water surface from its instantaneous average value, 0(t), referenced to the mean sea level (hereinafter MSL); z0 is the average marsh bottom elevation, referenced to the MSL; and is a bottom friction coefficient (Rinaldo et al., 1999a; Marani et al., 2003 for a detailed description). Further assuming tidal propagation to be much faster within the channel network than over the shallower flanking marsh areas, that is, considering a flat water level, 1 = 0, within the network, allows one to determine the field of free surface elevations over the unchanneled marsh platform, at any instant t of the tidal cycle, by solving the Poisson boundary value problem (1). On the basis of the resulting water surface, flow directions can be obtained at any location on the intertidal areas by determining the steepest descent direction, and watersheds related to any channel cross-section may be thus identified. The above simplified Poisson model applies, in principle, to relatively short tidal basins, that is, when the length of the basin is much smaller than the frictionless tidal wavelength (Lanzoni and Seminara, 1998). Nevertheless, as thoroughly discussed by Marani et al. (2003) by comparison with observations and complete hydrodynamic simulations, the Poisson model leads to quite robust estimates of drainage directions and watersheds, and, through the use of the continuity equation
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(Rinaldo et al., 1999b), of the landscape-forming discharges, even when the hypothesis of a short tidal basin is not strictly met. On the basis of the water surface elevation field, the distribution of bottom shear stresses due to tidal currents at every point x on unchanneled areas can be determined as follows: = D H1
ð2Þ
where (x; t) is the local value of the bottom shear stress, is the specific weight of water, and D is the local water depth. The analysis of the spatial distribution of (x) for our case study sites in the Venice lagoon (Figure 1a) is valuable in suggesting possible general features. It emerges that the higher values of the shear stress usually occur at the tips of the channel network and near pronounced channel bends. This observation is confirmed by Figure 1b that shows the probability density function of the shear stresses at the channel network heads, tips, and at all other adjacent sites with the exception of the tips, others. Such observations corroborate the speculation that headward erosion and tributary addition (possibly originating at sites where the stress increases along bends) are the main processes responsible for channel elaboration during its early development (Pethick, 1969; Steel and Pye, 1997; Allen, 2000). We thus suggest that channel headward growth, driven by the spatial distribution of local shear stress, is the chief land-forming agent for network formation on real marsh platforms. Under the assumption of approximately stable network configurations, the observed probability distributions provide useful information on critical shear stress values, which can be used in numerical simulations. The notion that erosional activities can be primarily expected in those parts of the basin where the local value (x) exceeds a threshold value for erosion, c, (Rinaldo et al., 1993, 1995; Rigon et al., 1994) is found to produce reasonable structures of tidal drainage densities and associated features within tidal landscapes (D’Alpaos et al., 2005).
2.2. Model of channel network early development We briefly review here the channel network development model, presenting its more relevant features, and refer the reader to the paper by D’Alpaos et al. (2005) for a more detailed description. Field evidence supports the main assumptions, in particular the hypothesis – adopted also in a number of conceptual models of salt marsh growth (Pethick, 1969; French and Stoddart, 1992; French, 1993; Allen, 1997, 2000; Steel and Pye, 1997) – that during its initial development stage a tidal network quickly cuts down through the intertidal areas, acquiring a permanent basic structure (in analogy with the case of fluvial settings, Rodrı`guez-Iturbe and Rinaldo, 1997). Such a quick initial network incision is later followed by elaboration (Pestrong, 1972; Garofalo, 1980; Marani et al., 2002), which does not alter its major features and, possibly, by vertical accretion of intertidal areas, which usually become vegetated when a threshold bottom elevation is exceeded. These considerations indicate the existence of different timescales characteristic of the various processes and justify the choice of decoupling the initial rapid network incision
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from its subsequent slower elaboration and from the eco-morphological evolution of intertidal platforms. Furthermore, based on the computed spatial distribution of bottom shear stresses (Figure 1), which shows how higher values are located at channel tips, we assume that the mechanism dominating channel network development is headward growth driven by the exceedances of a critical shear stress, c, which we take to coincide with a stability shear stress required to maintain an incised cross-section through repeated tidal cycles (Friedrichs, 1995). Depending on the spatial heterogeneity of sediment, vegetation, and microphytobenthos, which influences channel
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network dynamics, c may be assumed as constant or space dependent. Whenever the local bottom shear stress, (x), exceeds c anywhere on the border of the channels, erosional activity and network development may be expected: The model of channel network incision is thus based on the evaluation of the bottom shear stress distribution. According to the model proposed, the evolution of the network proceeds as follows. (1) For a given configuration of the channel network (initially consisting of a single-channeled site), Equation (1) is solved using representative values of 0 and @0/@t and the (x) distribution is computed from Equation (2). (2) One of the sites where (x) exceeds the fixed threshold for erosion, c, is selected on the basis of a suitable procedure governed by a parameter, T (which may be considered as “temperature,” in analogy with the simulated annealing procedure proposed by Kirkpatrick et al. 1983), expressing the possibly spatially heterogeneous distribution of the critical stress, and becomes part of the network. (3) The new channel pixel is considered to be part of the channel axis and channel crosssections are instantaneously adapted to the tidal prism, P, flowing through them, defined as the total volume of water exchanged through any cross-section between low water slack and the following high water slack, that is, during flood or ebb phases. In fact, it has long been recognized that a power-law relation between the tidal prism, P, and the minimum cross-sectional area, , holds for a large number of tidal systems believed to have achieved dynamic equilibrium (O’Brien, 1969; Jarrett, 1976). More recently, Friedrichs (1995), Rinaldo et al. (1999b), and Lanzoni and Seminara (2002) explored, in several tidal systems, the relationship between and spring (i.e., maximum astronomical) peak discharge, Q, which is directly related to the tidal prism, finding that a near proportionality between and Q also exists for sheltered sections. Friedrichs (1995) explains the existence of such relationship by relating the equilibrium cross-sectional geometry to the so-called stability shear stress, that is, the total bottom shear stress necessary to maintain a null along channel gradient in net sediment transport. Therefore, on the basis of Jarrett’s “law,” we consider the cross-sectional area, , to be related to the flowing tidal prism, P, (and therefore to the landscape-forming tidal fluxes responsible for shaping network geometry) through the relationship = pPp, with p = 10–4 and p = 1.1. Such an assumption allows the model to describe the evolution of the channel network in response to changes in the tidal prism, P, possibly due to variations in the elevation of the marsh platform or in relative mean sea level (hereafter RMSL). (4) Once the cross-sectional area has been determined, channel width is assigned based on a fixed value of the width-to-depth ratio, (Marani et al., 2002; Lawrence et al., 2004), in which we summarize the complex morphodynamic processes responsible for channel cross-sectional shape. Because the flow field has now varied due to the inclusion in the network structure of newly channelized pixels, Equation (1) is solved again using the new boundary conditions reflecting the updated channel configuration and steps (1)–(4), which represent a model time step, are repeated iteratively. As the channel network extends into the intertidal area, the reference water surface and its gradients are progressively lowered and the procedure is repeated until the critical shear stress is nowhere exceeded.
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2.3. Model of marsh platform evolution Consistently with field observations we model marsh platform morphodynamic evolution by considering cohesive and nearly uniform bottom sediment, and furthermore assume that sediment particles are transported mainly in suspension. Evolution of bed topography is governed by the sediment continuity equation which reads ð1pÞ
@zb = Qd Qe @t
ð3Þ
here zb is the local bottom elevation with reference to the MSL, p is void fraction in the bed, and Qd and Qe are the local deposition and erosion fluxes, respectively, representing sediment volume exchange rates, per unit area, between the water column and the bed. We evaluate the erosion flux, Qe, by a relationship that can be applied when bed properties are relatively uniform over the depth and the bed is consolidated (Mehta, 1984). 0 Qe = Qe0 1 Hð 0 e Þ ð4Þ e where Qe0 is an empirical coefficient depending on sediment properties, water salinity, biological disturbance, or binding; 0 =|| is the absolute value of the local bottom shear stress evaluated through (2); e is the cohesive shear stress strength with respect to erosion; and H is the Heaviside step function. We assume that Qe vanishes as vegetation encroaches the marsh surface, in accordance with field observations emphasizing that tidal currents are unable to produce excess shear stress over vegetated marshes (Christiansen et al., 2000). As far as the deposition rate is concerned, various sedimentation mechanisms need to be considered. We estimate the deposition flux, Qd, as follows: Qd = fC0 ðQds þ Qdt Þ þ Qdb
ð5Þ
where Qds is the settling rate, Qdt is the trapping rate due to the effect of the plant canopy, Qdb is the net organic production due to the presence of vegetation, and fC0 is a coefficient accounting for the frequency with which a given value of suspended sediment concentration, C0, occurs, as described in detail in D’Alpaos et al. (2007a). If the marsh is not vegetated, both Qdt and Qdb are equal to zero, whereas Qds acts even when the marsh surface is not vegetated. According to the feedback mechanism existing between morphology, hydrodynamics, and sediment dynamics (French, 1993; Lanzoni and Seminara, 2002), the settling and trapping rates can be determined only once the equation for suspended sediment concentration (hereinafter SSC) has been solved. However, bottom topography evolves on a much longer timescale with respect to the hydrodynamic circulation, thus allowing one to decouple the solution of the hydrodynamic field from the morphological evolution. Under the assumption that the flow is fully turbulent, the equation for the conservation of sediment transported as a dilute suspension takes the form of a
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two-dimensional advection–diffusion equation for depth-averaged volumetric sediment concentration, C(x; t), which reads @ðCDÞ þ H ðUCD km DHCÞ = Qe Qds Qdt @t
ð6Þ
where D is the local water depth, U is the local depth-averaged velocity field, and km is a constant horizontal mixing coefficient which accounts for dispersive effects associated with vertical variations in both flow velocity and sediment concentration. We estimate the deposition due to settling, Qds, by the formula proposed by Einstein and Krone (1962). 0 Qds = ws C 1 Hð d 0 Þ ð7Þ d where ws is the settling velocity depending on the size of sediment flocs and d is a critical shear stress below which all initially suspended sediment eventually deposits. Such an empirical description, taking into account the probability that a floc will survive the near-bed turbulence, is usually employed to describe cohesive sediment deposition in coastal environments (Krone, 1987; Pritchard and Hogg, 2003; Temmerman et al., 2005). Vegetation encroachment at the surface, for emergent marsh platforms, increases sediment deposition rates as a consequence of trapping effects and of organic soil production. A number of previous analyses and studies indicate that the amount of trapped sediment is proportional to the concentration of suspended sediment and to the number of plant stems that can both reduce the turbulent energy and capture sediment particles (Leonard and Luther, 1995; Nepf, 1999; Leonard and Reed, 2002). In analogy with D’Alpaos et al. (2006, 2007a), we have expressed the trapping rate using the approach proposed by Palmer et al. (2004), which reads Qdt = CU"ds ns min ðhs ; DÞ
ð8Þ
where U=|U| is the absolute value of the local velocity field, " is a capture efficiency (Palmer et al., 2004) that gives the rate at which transported sediment particles are captured by plant stems, ds is the stem diameter, ns is the stem density, hs is the average height of the stems, and D is the local flow depth. Note that when vegetation protrudes above the water surface, hs is replaced with the local water depth, that is, the minimum between hs and D is chosen in order to determine the effective height, which contributes to the trapping of sediment particles. Finally, the production of organic matter, Qdb, is linked to annually averaged aboveground plant dry biomass, B, following the pioneering work of Randerson (1979). According to the formulation proposed by Mudd et al. (2004) and employed by D’Alpaos et al. (2006, 2007a), Qdb, can be expressed as follows: Qdb = Qdb0
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ð9Þ
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where Bmax is the maximum value of the biomass and Qdb0 is a typical deposition rate, which is derived empirically from field measurements (Blum and Christian, 2004). Aboveground biomass, B, is one of the main factors through which the control of vegetation on hydrodynamics and sediment deposition is exerted. Such relationships were thus required to couple geomorphic and ecological models. Generally, the aboveground storage of organic material in salt marshes is an extremely complex process, which depends on vegetation characteristics and involves root production, microbial decomposition, as well as edaphic factors such as nutrient availability and salinity (Blum and Christian, 2004; Silvestri and Marani, 2004). Although several biotic and abiotic factors may be relevant in determining plant productivity (Silvestri and Marani, 2004, and references therein), locally, biomass production can, however, be related mainly to the elevation of the marsh platform encroached by plants. Such a relationship is the result of differences in soil aeration resulting from marsh flooding by the tide, and its form fundamentally depends on the biodiversity typical of the tidal environment considered. We, therefore, consider two different vegetated scenarios. In salt marshes characterized by a prevailing presence of Spartina spp., such as those typical of many North European and North American marshes, plant biomass may be expressed as a linearly decreasing function of salt marsh elevation (Morris and Haskin, 1990; Morris et al., 2002). This is due to the fact that the increased pore water salinity caused by evapotranspiration (enhanced by the progressive reduction of the duration and frequency of inundation, as the platform elevation increases) can limit the growth of, or be fatal to, salt marsh macrophytes (Phleger, 1971). As soon as the local marsh elevation exceeds zmin, that is, the elevation at which Spartina encroaches the marsh surface, we assume (Morris et al., 2002) that its biomass, B, decreases linearly with the difference zmaxzb, where zmax is the maximum elevation withstood by Spartina and zb is the local marsh elevation. The progressive increase in zb toward zmax thus leads to a decrease in plant productivity, eventually causing the disappearance of Spartina plants. However, such a relationship does not hold in salt marshes hosting a variety of halophytic species, as in the case of the marshes in the Venice Lagoon, where a mosaic of vegetation patches is observed (Marani et al., 2004; Silvestri et al., 2005; Marani et al., 2006a,b). When soil elevation increases, because Spartina is not well adapted to more aerated soil conditions, it is outcompeted by other species (e.g. Sarcocornia or Limonium in the Venice Lagoon) which thus take over. Therefore, biomass B (as well as species richness) rather than decreasing exhibits an increase with soil elevation, due to the typically higher productivity of halophytic species adapted to more elevated soils (Marani et al., 2004; Silvestri et al., 2005). In order to investigate sedimentation patterns arising from this second scenario and compare the results with those obtained by applying the model to the single-species scenario, we have assumed Spartina biomass to increase linearly with soil elevation. More specifically, biomass B is proportional to the difference zbzmin as long as zb is in the range between zmin and zmax, while, when zb becomes greater than zmax, plant biomass remains constant and equal to its maximum value (Allen, 1997). Based on the data collected by Morris and Haskin (1990) at North Inlet Estuary (South Carolina, USA),
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we have also assumed that vegetation characteristics (e.g., ds, ns, and hs) can be expressed as a function of plant biomass in both vegetated scenarios (see D’Alpaos et al. 2006, 2007a for a detailed description). According to our formulation, vegetation biomass influences sediment dynamics, thus affecting the patterns of net deposition, which, in turn, determine the change in soil topography. The latter determines changes in the spatial distribution of biomass, thus closing the feedback which is fully described in the model.
3. RESULTS The model of channel network development (D’Alpaos et al., 2005) makes it possible to analyze both the initiation of a channel network over an undissected tidal embayment and the further elaboration of an already incised channel structure. A variety of experiments were performed starting from different initial conditions in order to analyze the effects related to the position of single or multiple inlets, the shape of the tidal basin, different values of the width-to-depth ratio, and different values of the critical shear stress for erosion, c, and of the temperature, T. The model was also applied to simulate the evolution of a channel network within an actual catchment, emphasizing its noteworthy capabilities to reproduce real-life features (D’Alpaos et al., 2005, 2007b). Here we present the results of numerical simulations aimed at studying the competition among tidal creeks to drain the marsh platform adjacent to a larger tidal channel. Figure 2 shows some snapshots portraying the progressive development of creek networks within an idealized rectangular domain, limited by impermeable boundaries except for the bottom, flanking a larger tidal channel. The marsh platform is characterized by an average elevation z0 = –0.20 m above MSL. Channel network formation is a result of the dynamics of the system. Creeks are initiated at sites along the bottom channel where the first incision, initially due to chance, further grows because of the progressive flux concentration caused by creek development. At the beginning of the simulation, the domain is entirely drained by the boundary channel on the lower side and all of the boundary channel points drain the same amount of the watershed area. As soon as the networks start to develop, the drainage area associated with each of the growing networks, as well as their width, are relatively small. When the networks further develop and dissect the unchanneled domain, their watersheds and tidal prisms increase, thus causing their cross-sectional areas to increase as well, in order to accommodate the swelling tidal prism. The dynamics of the system is characterized by a “competition” among developing networks to capture the available watershed area. Stages of incision and retreat are observed, as well as situations in which divides migrate as a consequence of channel competition (Figure 2). In order to verify the validity of the proposed modeling approach, we compare relevant geomorphic features of the synthetic networks to those observed in actual tidal networks. The geomorphic characterizations necessary to compare synthetic morphologies with observed ones are provided by previous theoretical and observational analyses of the drainage density in tidal networks, relying on the
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statistics of unchanneled flow lengths, ,, that is, unchanneled flow paths from any unchanneled site to the nearest channel (Marani et al., 2003). Such statistics make it possible to capture site-specific features of network development and important morphological differences, providing a dynamically based geomorphic description which proves distinctive of network aggregation features (contrary to traditional Hortonian measures).
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Indeed, the analysis of a great number of actual marsh systems in the Venice lagoon showed a clear tendency to develop watersheds characterized by exponential decays of the probability distributions of , and thereby a pointed absence of scale-free features (Marani et al., 2003). Interestingly, the probability distributions of , for the synthetic networks generated by the model display a linear semilog trend of the type observed in the case of actual tidal patterns. Figure 3 portrays the evolution in time of the probability distribution of unchanneled lengths, P(L > ,), as the synthetic networks cut through the undissected domain, moving from configuration (a) to configuration (d) in Figure 2. The distribution of unchanneled lengths changes considerably as the network develops. The initial stages of network development are associated with larger values of the mean unchanneled length and with probability distributions that are quite far from an exponential form. At later stages, the mean unchanneled length decreases and the probability distributions tend to become exponential. This emphasizes that the model of network development is capable of providing complex structures and reproducing distinctive geometrical properties of geomorphic relevance (D’Alpaos et al., 2005, 2007b). We then analyze the long-term morphological evolution of the marsh platform dissected by the (above-generated) synthetic creek networks by applying the model
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proposed by D’Alpaos et al. (2007a), which allows to investigate the response of tidal morphologies to different scenarios of sediment supply, colonization by halophytes, and changing sea level. To this end, we consider an idealized initial topographic configuration represented by a flat marsh surface with elevation equal to –0.20 m above MSL. Furthermore, we assume the volumetric SSC within the channel network to be constant in space and time and, for moment, neglect MSL-rise effects. Note that the initial network structure is in equilibrium with the tidal prism computed on the basis of the initial assigned topography. We considered fine cohesive and uniform sediments characterized by density s = 2,600 kg/m3; particle diameter d = 50 mm; settling velocity ws = 2.0 10–4 m/s; porosity p = 0.4; and erosion rate parameter Qe0 = 1/s 3.0 10–4 m/s. The critical bottom shear stress for erosion, e = 0.4 N/m2, and deposition, d = 0.1 N/m2, are characteristic of fully consolidated mud (D’Alpaos et al., 2007a). We have further assumed zmin = MSL; zmax = MHWL; the maximum value of the biomass Bmax = 2,000 g/m2; and the deposition rate parameter Qdb0 = 0.01 m/year (Blum and Christian, 2004). Figure 4 shows the bottom topographies obtained from the model, after 100 years of simulation, under different scenarios, namely, in the absence of vegetation (Figure 4a); by considering a Spartina-dominated marsh (e.g., North Inlet estuary, N.C. – USA), in which the biomass decreases with soil elevation (Figure 4b); and in the case of a marsh characterized by a large species diversity (e.g., Venice Lagoon), in which biomass increases with soil elevation (Figure 4c). The spatial representation of the bottom topography makes it possible to distinguish the different sedimentation patterns that characterize the marsh platform evolution. The three modeled scenarios present common evolutionary features, but important and interesting differences emerge. In the absence of vegetation (Figure 4a), the accretion rate is entirely due to the settling of inorganic sediment. The magnitude of deposition processes and, therefore, the local vertical growth of the platform are seen to decrease with distance from the creeks. This is due to the reduction in sediment concentration with distance from the creeks, owing to sediment settling and to a progressive decrease of advective transport as prescribed by the advection diffusion Equation (6). In fact, advective flux is maximum near the channel and tends to vanish near the watershed divide where, by definition, flow velocity is equal to zero. Moreover, a negative feedback is found to exist between local surface elevation and accretion rates, according to which the local settling flux progressively decreases as the bottom elevation increases. In fact, the reduction in hydroperiod leads to a progressive decrease in sediment deposition, which vanishes as soon as the local elevation tends to MHWL, that is, the long-term equilibrium elevation attained asymptotically in this scenario. Such a behavior agrees with observational evidence by Pethick (1981) and with a number of numerical models describing salt marsh vertical accretion within the tidal frame (Allen, 1990, 1995, 1997). As soon as vegetation starts populating the marsh surface, significant differences emerge. In the Spartina-dominated case (Figure 4b), in which biomass decreases with soil elevation, vegetation encroachment significantly accelerates the vertical growth of the marsh platform. In fact, the deposition of organic matter, together with the enhanced inorganic sediment deposition due to trapping by the canopy,
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Multiple vegetation-species scenario
Spartina-dominated scenario
Unvegetated scenario
(a)
(b)
(c)
Marsh elevation (m)
<–0.4
–0.2
0.0
0.2
0.4
0.6
Figure 4 Comparison of the evolution of marsh surface topographies after 100 years, according to different scenarios: (a) no vegetation; (b) spartina dominance; and (c) multiple vegetation species. In all cases, the initial uniform bottom elevation is zb = ^0.20 m above MSL, whereas the initial configuration of the tidal networks is that represented in Figure 2d. After 100 years, the mean marsh platform elevation is equal to 0.2 m, 0.29 m, and 0.26 m for cases a, b, and c, respectively.
produce an abrupt increase in the total accretion rate as soon as the local elevation of the marsh platform exceeds zmin(ffiMSL), that is, the elevation at which Spartina exhibits its maximum productivity. In subsequent stages, two negative feedback mechanisms progressively decrease the total accretion rate, which vanishes as the
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local bottom elevation asymptotically tends to MHWL(ffizmax). The first is due to the reduction in the settling rate, associated with the progressive reduction in the hydroperiod as the local elevation of the marsh increases. The second is due to the decrease in Spartina biomass production with bottom elevation. A more gradual acceleration in the vertical growth of the marsh surface, together with a smoother increase in the total accretion rate, characterizes the multiple vegetation species scenario (Figure 4c) in which biomass increases with soil elevation. In this case, a positive feedback holds between the bottom elevation and the production of plant biomass (which is minimum for zb = zmin and increases linearly to a maximum as zb approaches zmax) that counteracts the reduction in the settling rate with soil elevation. As a consequence, marsh soil elevation increases faster than in the previous scenario, and eventually exceeds MHWL and the tidal range (although this is also due to the fact that relative sea-level rise is neglected for the purpose of describing platform dynamics). The total accretion rate of the marsh platform increases up to a local maximum and then decreases to a constant value, dictated by the organic sediment production rate, when zb becomes larger than MHWL( ffizmax). The space-time dynamics of marsh surface topographies in the three modeled scenarios (Figure 4) indicate that vegetation largely contributes to drive sediment patterns on the marsh surface and that the coupled evolution of vegetation and morphology gives rise to different system properties. The model accounts for variations of the channel network structure, which controls sediment transport dynamics on the marsh surface, and of the related watershed divides, as the marsh surface evolves in the vertical direction. The progressive vertical growth of the marsh leads to a decrease in the tidal prism flowing within the channels, from which follows a reduction in channel cross-sectional areas (assumed to be in equilibrium with the instantaneous tidal prism), thus resulting in a retreat of the channel network. Although channel network structure strongly controls sediment transport dynamics on the marsh surface, the evolution of soil topography shows notably different features in the modeled scenarios. The Spartina-dominated marsh exhibits a faster vertical growth as soon as the marsh platform exceeds the threshold elevation, zmin, the following stages being on the contrary characterized by a progressive decrease in plant productivity and therefore by decreasing growth rates. The accretion patterns that arise in the Spartina-dominated scenario are, therefore, qualitatively similar to the ones developed in the absence of vegetation, although with much larger total deposition rates and much larger topographic gradients from the channels to the inner marsh, due to the formation of more pronounced levees than in the unvegetated case. Also in this case, the elevation of the marsh platform asymptotically tends to MHWL, which represents an equilibrium condition for the considered scenario. On the contrary, the evolution of the marsh platform in the multiple vegetation-species scenario is characterized by a smoother initial increase in bottom elevation when halophytes initially encroach the marsh surface, whereas in the subsequent stages the vertical growth of the marsh platform progressively accelerates because of biomass increase with soil elevation: in this case, the positive feedback between soil elevation and organic production counteracts the progressively decreasing settling rate, and the marsh can approach faster the MHWL and eventually make the evolutionary transition to upland.
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4. D ISCUSSION The theoretical framework and modeling described here appears to reproduce geomorphologically relevant features of the eco-morphodynamic evolution of tidal networks and of the adjacent marsh platform. The choice to decouple the process of network initiation and early development (which appears to be quite a rapid one) from the subsequent slower network elaboration and evolution of the marsh platform is supported by field evidence (Steers, 1960; Collins et al., 1987; Wallace et al., 2005; D’Alpaos et al., 2007b) and by conceptual and numerical models describing the evolution of channel marsh systems (Allen, 2000).Steers (1960) reported a channel headcut migration of up to 5–7 m/year, Collins et al. (1986) observed a headward erosion of more than 200 m in 130 years, and Wallace et al. (2005) related a mean extension rate of 6.2 m/year. D’Alpaos et al. (2007b) described the rapid development of a network of volunteer creeks, branching from an artificial channel within a newly restored salt marsh, characterized by mean and maximum annual headward growth rates of 11 m/year and 18 m/year, respectively. These observations and the comparison between actual and modeled geomorphic network features (D’Alpaos et al., 2005, 2007b) also substantiate the assumption concerning the strong control exerted by the water surface elevation gradients, and by the related bottom shear stresses, in driving the process of channel incision (see also Fagherazzi and Sun, 2004). It is worthwhile noting that our modeling approach does not contradict conceptual models of “depositional network development” (Redfield 1965, 1972; Hood, 2006), which describe network development as the consequence of the vertical accretion and horizontal progradation of the vegetated marsh platform. Feedbacks that shape the network are fundamentally similar in both cases. According to the picture provided by our modeling approach, network incision and development take place at locations where the threshold shear stress is exceeded. In the vegetated platform progradation case, the extension of the channels is determined by the location where the bottom shear stress at channel heads is greater than the critical one (Redfield 1965, 1972; Hood, 2006). Feedbacks between the evolution of the marsh platform and the channel network may lead to variations in network geometry due to the differential accretion between the platform and the channels: stronger tidal fluxes within the channel may promote erosion or maintenance, whereas weak fluxes on the marsh may allow deposition fluxes to overcome erosion lading to marsh vertical/ horizontal accretion. Such an observation supports the known concept of inheritance of the major features of channelized patterns from sand flat or mudflat to a salt marsh (Allen, 2000). Moreover, the positive feedback between channel incision and flux concentration has recently been described by a number of numerical models of tidal landform evolution (D’Alpaos et al., 2006, 2007b; Kirwan and Murray, 2007; Temmerman et al., 2007). Numerical modeling of the cross-sectional evolution of tidal channels (D’Alpaos et al., 2006) supports the assumption of instantaneously adapting cross-sectional area to the flowing tidal prism, on the basis of a deterministic power-law relation (O’Brien, 1969; Jarrett, 1976; Friedrichs, 1995; Rinaldo et al., 1999b;
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Lanzoni and Seminara, 2002). The use of a constant width-to-depth ratio, which summarizes the complex morphodynamic processes responsible for channel crosssectional shape, is supported by observational evidence (Marani et al., 2002; Lawrence et al., 2004) and modeling (D’Alpaos et al., 2006). As a note, we observe that tidal-flat and salt marsh channels display different values of the width-to-depth ratio, , since they are seen to respond to different erosional processes resulting in different types of incision (Marani et al., 2002). In fact, the presence of halophytic vegetation on the marsh platform is likely to strongly affect bank failure mechanisms, and therefore salt marsh creeks tend to be more deeply incised (5 < < 7) than tidal-flat channels (8 < < 20). Moreover, D’Alpaos et al. (2006) showed that the width-to-depth ratio, , decreases as the adjacent platform evolves from a tidal flat to a salt marsh. Changes in can be accounted for in the model of network evolution, as already shown in D’Alpaos et al. (2005). Results from the model of marsh platform evolution concerning the equilibrium elevation reached by the platform within the tidal frame are in accordance with observational evidence (Pethick, 1981) and with a number of numerical models describing salt marsh vertical accretion within the tidal frame (Allen, 1990, 1995). The formation of marsh levees paralleling channel banks, which later broaden toward the inner part of the platform that exhibits a typical concave-up profile, is in accordance with field observations and modeling (e.g., Mudd et al. 2004; Silvestri et al., 2005; Temmerman et al., 2005; D’Alpaos et al., 2007a). The vertical accretion of the marsh and the related reduction in the tidal prism lead to shrinking channel cross-sections and to a contraction of the network through knickpoint retreat, in accordance with previous observations (see e.g., Allen, 2000 and references therein). The model accounts for the enhanced accretion rates due to vegetation encroachment on the marsh, which is seen as the chief vegetation effect on the long-term eco-morphodynamic evolution of the platform. Changes in the flow field due to the presence of vegetation (e.g., Temmerman et al., 2005, 2007) are not accounted for in the model, because such a process could be properly described only by using complete hydrodynamic models (Defina, 2000) whose use for long-term modeling would be prohibitive, due to the numerical effort required.
5. C ONCLUSIONS We have shown here that long-term modeling of intertidal biogeomorphic systems is feasible by suitably simplifying the description of the governing processes and yet retaining their physically relevant features. We have also shown that biological–physical interactions are key in determining the observed spatial patterns both in the biological and in the geomorphic domain. The models presented generate network structures that are quantitatively close to observed ones, while the topographic and vegetation spatial patterns produced are realistic and qualitatively similar to observed ones. The quantitative validation of the spatial patterns obtained from a biogeomorphic model is, however, always difficult to achieve. This is due to a difficulty in obtaining observations of the
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time evolution of the system, the long characteristic timescales, and the problematic definition of objective landscape metrics. This problem has been overcome with reference to the spatial organization of a channel network by use of statistics of unchanneled path lengths, proving a quantitative characterization of the relationship between channels and the intertidal landscape. Spatial patterns of vegetation biomass also require a quantitative validation and are being compared to maps of halophytic vegetation species obtained from remote sensing (allowing observations on scales between tens of centimeters and several kilometers). A similar issue needs to be addressed when comparing intertidal platform topographies, in which case the full suite of tools developed for the analysis of fluvial topographies may be applied (e.g., hypsometric curves, fractal, and nonfractal roughness measures, scaling properties of isolines with varying elevations, etc.), for example, aided by the wealth of spatially detailed data afforded by current airborne laser scanner systems. Nevertheless, the field of bio-geomorphological modeling of intertidal systems is still in its infancy. Several issues remain to be tackled, including the clarification of the response of vegetation to changes in soil aeration, a more complete quantitative description of biological effects on sediment mechanical properties (e.g., physical factors regulating the onset and the time-space variability of microphytobenthos influence on sediment erosion), and the incorporation of the action of wind waves on the margins of intertidal platforms strengthened by plant roots. Many other important issues have not been listed or may have been overlooked here. However, the results presented certainly point to the fact that a predictive model of the evolution of intertidal landscapes must be based on the recognition that the system cannot be decomposed into its biological and physical components and that its dynamics is intrinsically a biogeomorphic one.
ACKNOWLEDGMENTS Funding from: PRIN 2006 “Fenomeni di trasporto idrologici: formulazioni lagrangiane, ruolo di eterogeneitma, effetti morfologici, processi stocastici a scala di bacino;” PRIN 2006 “Modelli dellevoluzione ecomorfologica di bassifondi e barene lagunari;” the University of Padova project: “Telerilevamento della Zonazione e della Biodiversita` della Vegetazione sulle Barene della Laguna di Venezia;” VECTOR-FISR Project, CLIVEN research line (MM), are gratefully acknowledged.
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Redfield, A.C., 1965. Ontogeny of a salt marsh estuary. Science 147, 50–55. Redfield, A.C., 1972. Development of a New England salt marsh. Ecol. Monogr. 42 (2), 201–237. Rigon, R., Rinaldo, A., Rodrı`guez-Iturbe, I., 1994. On landscape self organization. J. Geophys. Res. 99, 11971–11993. Rinaldo, A., Rodrı`guez-Iturbe, I., Rigon, R., Bras, R.L., Ijjasz-Vasquez, E., 1993. Self-organized fractal river networks. Phys. Rev. Lett. 70, 1222–1226. Rinaldo, A., Dietrich, W.E., Vogel, G., Rigon, R., Rodrı`guez-Iturbe, I., 1995. Geomorphological signatures of varying climate. Nature 374, 632–636. Rinaldo, A., Rigon, R., Rodrı`guez-Iturbe, I., 1998. Channel networks. Annu. Rev. Earth Planet. Sci. 26, 289–327. Rinaldo, A., Fagherazzi, S., Lanzoni, S., Marani, M., Dietrich, W.E., 1999a. Tidal networks 2. Watershed delineation and comparative network morphology. Water Resour. Res. 35, 3905–3917. Rinaldo, A., Fagherazzi, S., Lanzoni, S., Marani, M., Dietrich, W.E., 1999b. Tidal networks 3. Landscape-forming discharges and studies in empirical geomorphic relationships. Water Res. Res. 35, 3919–3929. Rinaldo, A., Belluco, E., D’ALpaos, A., Feola, A., Lanzoni, S., Marani, M., 2004. Tidal networks: form and function. In: Fagherazzi, S., Marani, M., Blum, L. (Eds.), Ecogeomorphology of Tidal Marshes, vol. 59.American Geophysical Union, Coastal and Estuarine Monograph Series, Washington, 266pp. Rodrı`guez-Iturbe, I., Rinaldo, A., 1997. Fractal River Basins: Chance and Self-Organization. Cambridge University Press, New York. Savenije, H.H.G., 2001. A simple analytical expression to describe tidal damping or amplification. J. Hydrol. 243 (3–4), 205–215. Savenije, H.H.G., Veling, E.J.M., 2005. Relation between tidal damping and wave celerity in estuaries. J. Geophys. Res. 110, C04007. doi:10.1029/2004JC002278. Schuttelaars, H.M., de Swart, H.E., 2000. Multiple morphodynamic equilibria in tidal embayments. J. Geophys. Res. 105 (24), 105–124,118. Silvestri, S., Marani, M., 2004. Salt marsh vegetation and morphology, modelling and remote sensing observations. In: Fagherazzi, S., Marani, M., Blum, L. (Eds.), Ecogeomorphology of Tidal Marshes, vol. 59.American Geophysical Union, Coastal and Estuarine Monograph Series, Washington, 266pp. Silvestri, S., Defina, A., Marani, M., 2005. Tidal regime, salinity and salt-marsh plant zonation. Estuar. Coast. Shelf Sci. 62, 119–130. Steel, T.J., Pye, K., 1997. The development of saltmarsh tidal creek networks: evidence from the UK. In: Proc. Canadian Coastal Conference, Can. Coastal Sci. and Eng. Assoc., Guelph, Ontario, vol. 1, 267–280. Steers, J.A., 1960. Physiography and evolution, in. In: Steers, J.A. (Ed.), Scolt Head Island, second ed.Heffer, Cambridge, pp. 12–26. Stoddart, D.R., Reed, D.J., French, J.R., 1989. Understanding saltmarsh accretion. Scolt Head Island, Norfolk, England. Estuaries 12, 228–236. Tambroni, N., Bolla Pittaluga, M., Seminara, G., 2005. Laboratory observations of the morphodynamic evolution of tidal channels and tidal inlets. J. Geophys. Res. 110, F04009. doi:10.1029/ 2004JF000243. Temmerman, S., Govers, G., Meire, P., Wartel, S., 2003. Modelling long-term tidal marsh growth under changing tidal conditions and suspended sediment concentrations, Scheldt estuary, Belgium. Mar. Geol. 193 (1–2), 151–169. Temmerman, S., Bouma, T.J., Govers, G., Wang, Z.B., De Vries, M.B., Herman, P.M.J., 2005. Impact of vegetation on flow routing and sedimentation patterns: Three-dimensional modeling for a tidal marsh. J. Geophys. Res. 110, F04019. doi:10.1029/2005JF000301. Temmerman, S., Bouma, T.J., van de Koppel, J., van der Wal, D., De Vries, M.B., Herman, P.M.J., 2007. Vegetation causes channel erosion in a tidal landscape. Geology 35 (7), 631–634. Ursino, N., Silvestri, S., Marani, M., 2004. Subsurface flow and vegetation patterns in tidal environments. Water Resour. Res. 40, W05115. doi:10.1029/2003WR002702. Wallace, K.J., Callaway, J.C., Zedler, J.B., 2005. Evolution of tidal creek networks in a high sedimentation environment: a 5-year experiment at Tijuana Estuary, California. Estuaries 28 (6), 795–811.
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C H A P T E R
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T IDAL C OURSES : C LASSIFICATION , O RIGIN AND F UNCTIONALITY Gerardo M.E. Perillo
Contents 1. Introduction 2. Proposed Tidal Course Classification 3. Geomorphology of Tidal Courses 4. Course Networks and Drainage Systems 5. Origin of Tidal Courses 6. Course Evolution 7. Summary Acknowledgements References
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1. INTRODUCTION Tidal courses (otherwise know as channels, creeks or gullies) are the most distinctive and important features of coastal environments even with a minimum tidal influence. They represent the basic circulatory system through which water, sediments, organic matter, nutrients as well as pollutants are transported in and out of these wetlands. The tide is the heart that pumps water through the arteries and veins of the system allowing the exchange with the open ocean and the exportation of materials (including pollutants) from it. As tide enters any of these environments, such as tidal flats, salt marshes or mangroves, it first becomes channelized until water level overflows channel banks and levees and develops a sheet flow. Ebbing is the reverse process, first water recedes as sheet flow but final drainage is through the courses. Without tidal courses, life in coastal wetlands could not be possible as they provide nourishment, protection for local fauna and a place for reproduction, the growth of juveniles and, finally, a way out, when mature, to the open ocean for numerous species. In fact, tidal courses are one of the first features that appear upon the formation of a coastal wetland either by modification of earlier fluvial networks or by the direct action of the tides, groundwater and precipitation. As the wetland evolves in time and space, courses follow up and, many times, set the pace of this Coastal Wetlands: An Integrated Ecosystem Approach
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evolution since they, being the most energetic environment, are most sensitive to possible changes on the external variables that influence the systems. Despite their importance, tidal courses have been taken for granted as a feature always present but little studied in comparison with the associated flat, marsh and mangrove areas. When studied, measurements focus on particular features such as biological or chemical composition and circulation (especially in meanders). Only recently, a relatively small group of researchers have started to address the geomorphologic issues regarding tidal courses intensively and, in most cases, their work is concentrated on specific wetlands mostly in western Europe (i.e., UK, Italy and the Netherlands), the Americas (i.e., Argentina, Canada and United States) and Australia, but their ecological relevance is undervalued as reflected by the lack of research on these systems compared to other parts of the wetlands. Tidal courses are widespread and abundant in estuarine ecosystems (Mallin, 2004). There have been some initial attempts to describe the geomorphologic characteristics of tidal courses and even some basic assumption about the mechanism by which they originate. Some recent reviews (Eisma, 1997; Allen, 2000) provide an integrated approach from the classical geomorphologic point of view of our present knowledge about tidal courses. Work by the Italian group working on the salt marshes of Venetia Lagoon (Fagherazzi et al., 1999, 2004; Marani et al., 2002, 2003; Rinaldo et al., 1999, 2004) have introduced new concepts on how to study tidal course features and evolution. Furthermore, knowledge of their geomorphologic and sedimentologic characteristics is essential for the interpretation of the stratigraphic record. Until recent, estuaries were seldom found in geological papers as there were no adequate correlations between the present-day conditions and what was observed in geological outcrops (Perillo, 1995). Estuaries in general and specifically tidal courses have relatively small area distributions which are difficult to find in stratigraphic outcrops. In many cases, their imprint may even be confused with unidirectional flows if other criteria (i.e., estuarine fossils) are not present. The mechanisms that originate the courses are largely unknown. Although this seems a very simple concept, there is no real agreement as to how and when they initiate, not even if there are only one or multiple processes that develop them. In most cases, course initiation is an underwater process occurring over the intertidal area which is commonly occluded from direct observation by suspended sediments. Even if water transparency were not a problem, there is no way to predict when and where a rill will form, or if the courses will persist in time when they are flooded during the next tide. Probably, many of the uncertainties arise from the simplistic although inaccurate idea that tidal courses are somewhat the bidirectional flow counterpart of fluvial streams. The original studies made by pioneers like Leopold et al. (1964) have specifically compared fluvial and tidal channels trying to apply various classification and statistical criteria employed in the terrestrial counterpart to the marine course networks. Preliminary studies, especially those related to drainage networks (Pethick, 1992; Pye and French, 1993) have considered fluvial terminology to describe tidal course networks although there are clear differences related to hydraulic and geomorphic considerations.
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Despite over 150 years of research on fluvial systems, there are still no adequate explanations for many of their basic problems, including the origin of gullies and rills (Tucker et al., 2006), which are very important features compared with tidal gullies and rills, still less is known from tidal courses when the length of time that they have been studied is only a minimum portion of the previous one and, for the most parts, the field study conditions are much more difficult. In this regard, the difficulties to study both origin and evolution of tidal courses in physical models are much greater than fluvial ones due to the lack of widespread laboratories that have tidal-emulating facilities and also the additional problem of resolving the sediment scaling. This is not a minor issue since the large majority of coastal wetlands are dominated by fine (silt and clays) sediments whose behavior varies depending on the local environmental factors and interaction with the surrounding material. Therefore, the objective of this chapter is to provide a review of the present knowledge of tidal course characteristics, networking and drainage systems as well as to present some ideas and examples of mechanism for course origin. We also provide a possible classification of the tidal courses as a way to simplify the existing confusion based on the indiscriminate use of various common names.
2. PROPOSED T IDAL C OURSE C LASSIFICATION Upon scanning the literature on wetland morphology, there is a remarkable confusion about the variety of names given to the different valleys that intersect wetland. Names such as tidal channels and tidal creeks are very common and used interchangeably even in the same publication. Other common terms are gullies, rills, canals, and so on. A similar confusion occurs in the fluvial literature, especially regarding the differentiation between rivers and creeks. In the latter case, the problem appears to be related, in most cases, to local denominations without consideration of geomorphologic features notably distorted by the variety of definitions that exist in dictionaries and encyclopedias. In the present context, we propose a basic definition and classification of tidal courses to provide a common descriptive ground based on size and persistence of water on the course during low tide conditions. We define a tidal course as any elongated indentation or valley in a wetland either originated by tidal processes or some other origin, through which water flows primarily driven by tidal influence. Tidal course is a general denomination that includes a series of indentations within a wide spectrum of sizes (width and depth) and with at least two levels of inundation (Table 1 and Figure 1). The classification based on size ranges proposed also provides a descriptive terminology for all tidal courses. Only depth and width of a cross section estimated to the bankfull level are considered since length could suffer large variations, especially if there are artificial or major geomorphologic constrains (i.e., dikes, cliffs, etc.). Tidal rills (Figure 1a – tr) are very small superficial indentations developing in the later ebb stages along the unvegetated, sloping margins of larger courses or marsh fronts. In order to form, they require a small veneer of fine sediments which
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Table 1 General classification and size range for tidal courses Name
Water in low tide
Depth (cm)
Width (cm)
Cross-section area (cm2)
Tidal Tidal Tidal Tidal Tidal
No No No Yes Yes
<1 1–5 5–100 10–200 >100
<2 2–10 10–100 10–200 >200
<2 <50 50–1,000 100–4,000 >2,000
rills grooves gullies creeks channels
Depth is the mean vertical distance from the thalweg to the bankfull border of the indentation. Width is the mean horizontal distance measured across the indentation between the bankfull borders.
(a)
(b)
(c)
(d)
Figure 1 Examples of end members in the tidal course classification. (a) Tidal rills (tr) and tidal grooves (tg), (b) tidal gullies, (c) tidal creeks and (d) tidal channels.
is dissected by very low flowing waters. Depending on the surface slope and sediment characteristics (i.e., size, cohesiveness and depth), rills vary in shape from linear to sinuous and may even develop braided conditions (both distributary and convergent). Another common mechanism of rill formation, as is also frequent on sandy beaches, is due to groundwater discharge resulting in a large number of rills having a variety of shapes. Rills may be considered as the first step in the initiation of a tidal course and their preservation and further evolution depends on
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the depth of the indentation, the soil characteristics and most important, the processes occurring when tides inundate the course margin back. If the indentation progresses in depth to more than 1 cm, the rill becomes a tidal groove. Grooves (Figure 1a – tg) tend to be 1–5 m long or exceptionally even longer courses, 1–5 cm deep and up to 10 cm wide. Normally, they are linear to sinuous in sectors. They develop along channel margins and marsh fronts with relatively high (3–7) slopes due to strong ebb flows or groundwater outflows. Tidal grooves are more prone to resist the following tidal inundation as they produce a much deeper indentation in the soil. It is common to observe parallel groves along the margin of typical tidal channels (Figure 5a). Concentration of the flow in some of those grooves may produce the formation of larger courses such as gullies or creeks. Tidal gullies (Figure 1b) are similar to those observed in continental drylands. They are deep indentations that may reach up to 1 m and much wider than the preceding courses. Gullies are preserved and enhanced by tidal inundation. They could easily superimpose in size with tidal creeks, the difference lies in the lack of tidal water during low tides although some flow may be observed, normally due to rainwater retained on the flats or marshes, or groundwater outflows. In all three cases, the larger relative relief is found near the head of the course becoming shallower as it approaches the mouth. Their course is also linear to sinuous and they seldom develop meanders, but if they do, those meanders appear at their mouth. The major dynamic structures in wetlands are tidal creeks and channels. Probably, the main difference with the other courses, as indicated, is the permanent inundation by tidally driven water in at least part of the course even in the lowermost tides. Creeks range in size from few tens of centimeters in depth to up to 2 m while width has similar values. They normally have water during low tide; however, water depth during this time varies from practically none at the head to about 10–30% of bankfull depth at the mouth. Creeks, as well as the previous courses, are the tributaries of the tidal channels and form at their banks but they are distributed over tidal flats and marshes. Rills and grooves never reach the level of the tidal flats and marshes and gullies seldom do. On the other hand, tidal channels are the largest features in most wetlands as they clearly stand out in maps and satellite images. Channels always have water along their whole course even during the lowest spring tides. Their depth is greater than 2 m and maximum values are highly dependent on the characteristics of the wetland, but on average they have as much as 10 m, in many cases reaching up to 30–40 m deep as in the case of coastal lagoon inlets. Regarding channel width, the degree of variability is much larger than in the case of depth. Tidal channel widths start at about 2 m and may reach several kilometers (i.e., Bahı´a Blanca Estuary and Ord Estuary).
3. GEOMORPHOLOGY OF TIDAL C OURSES At first sight, tidal courses resemble concentrated fluvial networks having also similar course shapes both in plan view and in cross section. However, when analyzed in further detail, numerous differences appear that set them apart.
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The most obvious ones are the flow characteristics (roughly unidirectional vs. bidirectional), relative relief, degree of course inundation, evolution pattern, among-course interactions and so on. Although there are wetlands developed in sandy substrates, most of them are dominated by fine sediments, therefore, braided patterns are rare and most valleys have single courses with roughly linear, sinuous or meandering patterns. Various authors have partially analyzed the cross-section shape of courses in wetlands (Pestrong, 1965; French and Stoddart, 1992; Collins et al., 1998; Allen, 2000; Fagherazzi et al., 2004). The cross section form depends on wetland stratigraphy, tidal range, their associated plant cover and sinuosity (Allen, 2000), and they can be classified as V-shaped (Figure 2a), U-shaped (Figure 2b), asymmetric (Figure 2c), complex (Figure 2d) or with one or both banks significantly overhanging (Figure 2e and f). Beyond a general description of the shapes, there are no actual studies dedicated to relate cross-section forms and the course conditions, evolution and size. Both V- and U-shaped (Figure 2a and b) are common in small courses and even in the linear portion of channels and creeks, whereas asymmetric cross sections are normally related to meandering reaches. Many linear channels also show clear asymmetries (Figure 2c), specially if they are subject to lateral migration (Ginsberg and Perillo, 2004). These types of cross sections appear during the initiation of the course, when the linear pattern is still present and most of the flow strength is used to deepen and lengthen the course. Complex cross sections (Figure 2d) appear when bars are present along the course. Kjerfve (1978) described them as bimodal with two channels separated by a shallow area or bar having contrasting residual circulation associated with differential depths (flood and ebb dominance). However, there is still no clear indication whether the differential circulation is due to the bathymetry or the bathymetry is a consequence of the change in direction of the residual circulation often found on vertically homogeneous estuaries (Dyer, 1998) as there are no reports describing the formation process. Overhanging banks (Figure 2e and f ) are the result of differential sedimentary characteristics as they are the remnant of flow undercutting by tidal currents within the course. The overhanging portion, due to higher sediment compaction (a)
(b)
(c)
(d)
(e)
(f)
Figure 2 Description of the various possible cross-sections of tidal courses. V- (a) and U-shaped (b) normally corresponds to longitudinal portions of courses. Asymmetric (c) and complex (d) are found commonly in long meanders, whereas overhanging simple (e) and complex (f ) represent differential erosion processes controlled by variations in sediment characteristics, water level in the course and/or differential material strength due to marsh plants.
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or resistance or sustained by plant roots, has been less affected by the currents and preserved while a lower portion of the bank is undercut. Preservation of overhanging banks is poor as they are intrinsically unstable features that will collapse eventually, and the material is transported by the tidal currents. In most cases, longitudinal profiles of tidal rills, grooves and gullies tend to be concave upward at the close-ended head becoming shallower and convex upward for most of their course (Figure 1a). At the head of these courses, there is normally fast erosion by headward retreat which induces a relatively high relief represented by a microcliff (Minkoff, 2007). Ebbing water cascades at the microcliff, resulting in a deeper thalweg at the head. In areas where sediment is compact, the head, flanks and bottom of the course at and near the head are marked by dislodged sediment clasts, flakes, crumbs and plant patches, remains of infauna burrows, and so on yielding a very irregular and even chaotic morphology (Figure 3a). Moving along the course toward the mouth, one will find that the relative relief diminishes in addition to the irregularities in the bottom and flanks, since the tidal flow tends to smooth them out by transporting and wearing out sediment blocks (Figure 3b). In recent deposited
(a)
(d)
(b)
(c)
(e)
Figure 3 Various examples of tidal course characteristics. (a) Complex structure of a gully head, (b) view of a tidal creek morphology, (c) marked differences in tributaries along a tidal channel, (d) typical meander in a tidal channel at low tide showing the formation of a point bar in the inner flank of the curve and the cliff on the outer flank and (e) variability of the meander pattern over a tidal flat showing that longer curvature radius occur on the flatter portion of the flat.
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sediments that have not achieved a certain level of compaction, preservation of small courses becomes difficult as tidal currents along creeks and channels resuspend and transport them and, as the courses are transversal to the flow, the potential to be filled up is high. In these cases, the flanks and bottom are smooth. Tidal creeks and channels, when established, have relatively smooth flanks and bottoms, worn out by currents, and irregularities appear related to erosional processes at the outer bend of meanders, infauna activities, rotational slumps and deformed microdeltas deposited at the mouth of tributaries. Longitudinal profiles show a progressively higher relative relief toward the mouth. Tributary inlets often show sediment deposits that range in shape from typical deltas to extended banks parallel to the main channel. These deposits result from the sudden drop of the suspended and bedload material being transported during the ebb, especially at or near low water slack. Depending on the strength of the currents, these deposits are either carried away or deformed, normally following the channel current dominance. Most examples found at the Bahı´a Blanca Estuary (Figure 3c) and several others along the Argentine coast show a continuity of the inlet along the main channel flank bordered by a parallel shoal which, in most cases, is ebb-directed. Potentially, the displacement of the mouth induces an initial meander that may affect the circulation within the tributary. Although there are no studies that assert it, this is clearly an instability on the course plan morphology that could displace headward, resulting in a meandering pattern developed inversely to accepted theories in fluvial systems (Leopold et al., 1964). Another interesting but often overlooked feature of tidal courses are the scour holes that form at the junction of two courses. Studies of scour holes have concentrated in fluvial junctions where their morphological features have been related to the angle of the junction, tributary/main river discharge ratio, sediment erodability, and so on (Best, 1987). Scour holes in tidal environments have only been mentioned by Shao (1977), Kjerfve et al. (1979) and Ginsberg and Perillo (1999). The results obtained demonstrate that scour holes, having large relative relief in excess of 2 m and up to 17 m, at tidal environments differ significantly in morphology from those found in fluvial systems. Those observed in the Bahı´a Blanca Estuary, for instance, although elongated in shape as in the case of fluvial holes, have the steeper side (3.5) at the mouth of the confluent channel and the gentler side (1.5) seaward. This structure is exactly opposite to scour holes found in river environments. Based on current measurements and sediment transport estimations, there is a clear flood dominance on the steep, inner face and ebb dominance over the gentler, outer flank (Ginsberg and Perillo, 1999). Another difference between holes in fluvial reaches and tidal scour holes is that the latter migrate headward. Courses in marshlands have an additional mechanism to deepen their channels: levee accretion. Restricted by surrounding vegetation, sediment deposition along the course increases the relative relief between channel bottom and marsh surface. This process also inhibits water exchange, resulting in water and sediment becoming trapped at high water and during surges. Both in marshes and mangroves, vegetation plays an important role in sediment trapping and stabilizing the wetland (Wolanski, 2007). Lateral and headward erosion activities also widen and lengthen
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the channel (Chapman, 1960). The type of wetland vegetation and the vertical distribution of plant species have significant influence on the degree and rate of stream channel migration and meandering (Garofalo, 1980). Although meandering flows have been extensively studied, especially in fluvial environments (Leopold and Wolman, 1960; Ikeda et al., 1981), an adequate theory that fully explains meanders is still lacking. The leading process in meander development in fluvial rivers is the redistribution of momentum due to channel curvature (Ikeda et al., 1981). Flow momentum concentrates along the outer bank originating its erosion. Secondary transversal flows driven by superelevation, transport the sediment which deposits on the inside bank resulting in the development of point bars. Furthermore, alternate bars can lead to the formation of meander bends at initial stages of meander development (Blondeaux and Seminara, 1985; Seminara and Tubino, 1992). Beyond the specific mechanism of meander formation, the main hypothesis as to why a course meanders is based on the concept that meanders elongate the water path thus enabling it to contain more of the water discharge within the same valley distance. Although this is true for rivers and other contained flows, it appears somewhat difficult to transfer this concept to tidal environments where the water is contained within the course only for a part of the time, even though it is the period with the highest velocities, whereas a large percentage of the flow spreads out over the contiguous wetland as the bankfull stage is overcome. The only correlation could be that once the water is concentrated in the course during the ebb, the total water volume may be larger than the actual volume that the course can discharge, thus causing a situation similar to that found in rivers. Tidal meanders have a cross-sectional morphology similar to fluvial meanders. The outer bank tends toward a steeper flank, sometimes a cliff develops (Figures 2d and 3d) and, eventually point bars evolve on the inner bank. Flow circulation on the meander appears, in general, similar to the circulation in river counterparts although this is a simplification that cannot go beyond the cases in which the meander is symmetric. The fact that in coastal wetlands tides control circulation implies that the flow along courses is bidirectional. Both the bottom topography and overflow are causes for common flow asymmetry which, eventually, could be enhanced by any river discharge. Therefore, flow around a bend in a tidal course is seldom symmetric and the resulting morphology of the bend suffers its consequences. Meander shape in coastal wetlands, as in fluvial systems too, is controlled by the sediment properties and erosion–sedimentation history of the environment. Wetlands developed over older fluvial terraces or deltas being eroded may have layers of sediment of variable consistencies outcropping and, therefore, they can control the direction and planform shape of the whole course but most importantly that of the meanders. That is the case of the Bahı´a Blanca Estuary and Anegada Bay (Argentina), which were both part of the late Pleistocene-Early Holocene Colorado River Delta (Perillo and Piccolo, 1999), and it is common to find meanders with a “straight loop” (Figure 3e) following the orientation of the original delta sets. In many cases, meanders are also initiated by rotational slumps or “erosion cusps” (Ginsberg and Perillo, 1990).
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Symmetrical currents may erode the banks differently resulting in meanders that are narrower at the bend apexes and wider between bends (Ahnert, 1960). Bend skewing, a normal feature in fluvial meanders due to their downflow asymmetry produces the “gooseneck loops” (Carson and LaPointe, 1983; Fagherazzi et al., 2004). However, due to the flow asymmetry, there are distinct planform morphologies depending on whether the course is ebb- or flood-dominated (Fagherazzi et al., 2004). As in rivers, wetland meanders also have point bars whose shape is controlled by the ebb-dominated tidal currents and the radius of curvature (Barwis, 1978). The theoretical concept that equilibrium courses should have an exponential increase of their cross section from head to mouth could only apply to the large tidal channels and creeks. The degree of strong meandering paths that most small courses clearly have shows that equilibrium cannot be easily achieved in coastal wetlands. In these environments, the dynamical processes are constantly changing, flow seldom stays concentrated within the course and flow discharge may vary, even with the same tidal amplitude, by a simple change in wind direction or speed. Especially, higher order tidal channels which conduct most of the tidal prism may approach this equilibrium idea. A model for young marsh environments where sedimentation is active demonstrates, in contrast with terrestrial rivers, that salt marsh creeks experience a strong increase in width–depth ratio seaward due to the short duration time of the peak discharge (Fagherazzi and Furbish, 2001). Course widening results from the imbalance between erosion versus deposition and the course cross section does not have time to adjust itself. Furthermore, autoconsolidation of the bottom cohesive sediments and the consequent vertical gradients in resistance properties obstruct the formation of deep courses, yielding the formation of shallow wide courses (Fagherazzi and Furbish, 2001). On the other extreme, course heads may have the same variability in shape than those observed in fluvial systems. Course heads can be defined as the upslope boundary of concentrated water flow and sediment transport between definable banks (Dietrich and Dunne, 1993). However, we follow the previous definition of course head given by Montgomery and Dietrich (1992) as the upstream limit of observable erosion and concentration of flow within definable banks. Minkoff and Perillo (2002) and Minkoff (2007) have differentiated three types of course heads: diffuse (with an undefined form, generally convergent with the surrounding land, Figure 4a), acute (ending in a point with a width <15 cm, Figure 4b) and open (wide, mostly round shaped with a width 15 cm, Figure 4c). In each of the head types, different processes act. In diffuse heads, the predominant process is the surface erosion of the sediment whereas other factors such as groundwater or headward erosion are of little consequence. In the case of acute heads, both subsurficial and headward retreat associated to water cascading are the most important factors, whereas subsurficial erosion and mass wasting are the major factors in the formation of open heads. Studies of the time and spatial evolution of creek and gully heads indicate that their shapes varied as the courses were growing (Minkoff, 2007). For this particular case, the effects of the burrowing crab Neohelice (formerly Chasmagnathus) granulatus play an important role, and landward advance of the heads are most significant
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(a)
(b)
(c)
Figure 4 Examples of the three types of gully heads described by Minkoff (2007). (a) Diffuse, (b) acute and (c) open.
(about 2–3 times faster) during the warmer (late spring, summer and early autumn) than colder months due to the high crab activity and population increase (Escapa et al., 2007). For instance, diffuse heads do not grow directly at the head, but there is surficial erosion some distance from the border and then a sudden connection with the thalweg occurs by a depression on the surface. In all cases, Minkoff (2007) has demonstrated, after surveying more than 130 creek heads over a period of 4 years, that creek and gully landward growth is a pulsating process. Basically, there is a period of preparation after which the course retreats a certain distance in one single movement. Minkoff (2007) seems to be the first detailed analysis of the retreat processes in coastal wetlands, except for the specific study made by May (2002) in which she related creek headward retreat on the level of mass wasting of the terrace between the head and the frontal marsh area incised by the creek. She divided it into strong terrace wasting (STW), weak terrace wasting (WTW) and no wasting (NW). The former occurs in sectors where there are thick deposits of organic-rich soil (at least 10 cm deep). The transition zone of these creeks had a steep slope and a large drop in elevation of the marsh surface (>15 cm). The transition zone presented various erosional features such as large holes (at least 5 cm in diameter)
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in the sediment surface that were coalescing, as well as undercutting and slumping of sod. Those of WTW appear in areas with thin deposits of organic-rich soil or overlying root mat with a smooth slope. The NW sectors have very little erosive features.
4. C OURSE NETWORKS AND D RAINAGE S YSTEMS At first approach, drainage patterns on coastal wetlands resemble fluvial networks. But once they are analyzed in more detail, the first difference that appears is that the number of branches and channel hierarchy (in the Hortonian sense) is closer to mountain basins rather than to lowlands to which coastal wetlands could be related. This factor by itself indicates that the network is draining large amounts of water in short time as opposed to flatland networks. As also occurs with their fluvial counterpart, courses at the head are blind-ended channels for marshes (Ashley and Zeff, 1988) and tidal flats. Blind-ended courses are a mechanism to show local processes that control the network (Stark, 1991). The second most striking difference is the lack of topographically defined basins. Although some geomorphologic control (likely to have developed after network inception) may exist, basin boundaries are controlled by hydrodynamic processes. Marani et al. (2003) found that the drainage density of a network, which was defined by Horton (1945), as proportional to the ratio of the basin’s total channelized length divided by the watershed area, must in fact be defined by the statistical distribution and correlation structure of the lengths of unchanneled pathways. By modeling the current directions on the wetland surface, basin boundaries can be drawn and a network structure defined. A situation that arises because, a large percentage of the time, water on the wetland is moving as overmarsh flow rather than being channelized. Flow directions under these conditions may be strongly affected by winds especially in areas where wind is a major factor, for instance, by having dominant directions or by storm surges. Also a wetland having different levels of inundation during the regular spring–neap cycle would be subject to different behavior for each sector resulting in modifications of the network. Therefore, drainage patterns at a high wetland could be rather different from that at the lower wetland. Novakowski et al. (2004), on the other hand, have geomorphologically defined a marsh island as a section of marsh circumscribed by tidal channels that are deeper than 1 m at low tide. Although this definition is somewhat artificial, it provides an objective way for a first-level estimation of hydraulic conditions in a wetland. Marsh course watershed was estimated by connecting end points of first-order courses. Network density could also be affected by anthropic actions such as ditching. Ditched marshes (a common practice in Europe and North America in the late 19th and early 20th centuries) have both less course density and pool density (Lathrop et al., 2000; Adamowicz and Roman, 2005). Pool density is an important factor in course initiation and development since pool interconnection may lead to course
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formation. Furthermore, some basin control may occur due to morphological/ artificial changes such as dikes or raised continental features. In some cases, the hydraulic control is based on the fact that tides, even during storm surges, cannot inundate the backmarsh. Changes in vegetation patterns can often be a good indicator of hydraulic patterns (Pethick, 1992). Figure 5 provides a series of examples of the various network patterns that can be observed in coastal wetlands. A variety of linear, sinuous, rectangular and dendritic patterns are common to all wetlands. Their diversity, unless artificially affected, is still a matter of discussion since there are no clear theories that can explain why a particular network pattern is present. Evidently sediment characteristics, previous sediment history and wetland slope all play a role, but to what extent each of them is actually responsible for a particular pattern is presently unknown. The degree of control that plants have over the development of tidal course networks can hardly be better shown than by the case of Caleta Brightman (southernmost channel of BBE) (Figure 5e). The main channel is borderer in most of its extension by relatively newly developed Spartina marshes, which are from a few hundred meters to 1 km wide, and from there to the continent there are muddy tidal flats about 1–2 km wide. Channels and creeks crossing the marsh are mostly
(a)
(b)
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Figure 5 Examples of drainage patterns on a coastal wetland. (a) Rectangular, (b) linear dendritic, (c) sinuous and (d) sinuous dendritic.
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linear or mildly sinuous with very few, parallel tributaries connected at right angles to the courses. As these courses are followed into the unvegetated areas, the network suddenly develops into a well-defined dendritic pattern with sinuous to meandering courses. The number of tributaries increases at least threefold, bearing in mind that the flats and marshes are subject to similar inundation regime.
5. O RIGIN OF T IDAL C OURSES After over 200 years of fluvial studies, detailed description of course initiation is still lacking (Montgomery and Dietrich, 1988, 1992, Istanbulluoglu et al., 2002; Kirkbya et al., 2003; Tucker et al., 2006). Obviously, there are many problems to reach consensus within the wetland community (Guilcher, 1957; Chapman, 1960; Pethick, 1969) but also in the well-documented fluvial literature. Based on Chapman (1960), Perillo et al. (1996) described the possible mechanism of tidal course formation as follows: during the ebb tide, sheet flows are guided by the topography, concentrating their discharge into depressions. These depressions become small courses by headward erosion. As the water velocity increases in the course during successive ebb tides due to the tendency to concentrate the retreating flow (Pestrong, 1965), these depressions become wider and deeper. Headward erosion is an important factor in further developing the course. As the course grows in size, meanders form, which are the final pattern that the channel acquires when fully developed. Headward erosion lengthens the course in combination with local annexation of nearby courses. The main incognita still remains as to why a course forms at the place it does and not somewhere else. For instance, a probabilistic approach was proposed associated with the flank slope and the erodability conditions of the sediment for continental situations (Istanbulluoglu et al., 2002). Following the original Horton (1945) idea that course inception occurs at a certain distance downslope of an incline, many grooves appear at the flanks of tidal creeks and channels during low tide. Horton’s idea is that overland flows may require certain distance to achieve enough bottom shear stress to overcome soil resistance. In other words, an erosion threshold controls the location of channel heads establishing that the same critical distance below a topographic divide is required for a sheet flow to erode sediment as for it to initiate a course. According to Horton’s theory, courses, such as erosion features, may expand rapidly upslope in response to changed climate and land use conditions and can even form during individual storm events. Although this is true for thin overflow layers, typical of continental processes, in coastal wetlands during ebb, the water depth along the flanks is at maximum and diminishes as low tide is reached. Nevertheless, at the initial stage of the ebbing phase of the tide, most of the flow may be concentrated along the main course and the flow cascading from the surrounding flats or marshes could be diverted toward the main course mouth and could not reach enough momentum to produce any erosion. As far as we were able to detect, there are no direct measurements of the flow dynamics very close to the surface along a channel margin during a tidal cycle;
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therefore, the following analysis is to be considered only as a hypothesis yet to be demonstrated. Course inception on channel flanks is produced only at the final stages of flow sheet as it moves downslope. At that moment, this sheet is only probably a few centimeter deep; therefore, the flow is both supercritical and most likely turbulent as well. Supercritical flows induce stationary waves that act damming the flow until enough depth is reached to reduce the Froude number at or below the critical value. Then the waves disappear and the flow is strongly accelerated which, in itself, can be strong enough to generate a bottom shear stress greater than the critical value for cohesive sediment erosion. As in most wetlands, this process occurs over fine sediments, which as they are eroded pass directly into suspension and thus increasing the flow density which further augments the erodability capacity of the flow. In the waters along channel borders, at low tide, one can commonly observe a 10 cm to up 1- to2-m wide streak (depending of the channel breadth) of high suspended sediment produced by this downslope erosion. The incipient grooves have a remarkably constant spacing of the order of 1–10 m which may depend on the slope of the segment of tidal flat where they develop (Perillo et al., 2005). Potentially, all the grooves may develop further into gullies, creeks and channels, but only a few of them actually upgrade to the next level. However, Horton’s general idea does not explain the presence of multiple, almost parallel grooves formed over a smooth muddy surface. Evidently, there must be other factors that induce course inception at specific points on the surface. The presence of the periodic grooves may be due to (1) periodic surface irregularities (i.e., along channel undulating surface); (2) presence of pebbles, infauna burrows or mounds; (3) nonuniform sediment (i.e., size, degree of compaction, mineralogy, bioturbation, etc.); (4) presence of vascular plants or roots; and (5) variability of the layer of microphytobentos and their excretions (i.e., EPS). Any or a sum of these items could be responsible for routing the flow, inducing hydraulic jumps. Eventhough all those conditions can be found individually or in groups acting together in helping the development of grooves, there is an important issue along channel banks that is often overlooked: groundwater seepage. Most studies of groundwater exchange between the flats and the marshes have been focused on the biogeochemical aspects (Findlay, 1995; Hollins et al., 2000; Osgood, 2000; Kelly and Moran, 2002; Duval and Hill, 2006). Gardner (2005) and Gardner and Wilson (2006) modeled the seepage from channel banks indicating that this process is a major mechanism for water and nutrient exchange between the course and the wetland. The process is driven mostly by the tide but affected by rain, plant roots and infauna burrows, resulting in complex and highly transient variations in boundary conditions along creek banks and the marsh platform (Gardner, 2005), and the alternate development and disappearance of water tables, seepage faces and zones of saturated and unsaturated flow. Tidal rills are one of the main consequences of seepage (Figure 1a) as they mostly develop from the groundwater discharge through the saturated zone on channel and creek banks. Rill initiation associated with groundwater discharge, as also occurs on sandy beaches or drylands, is due to sediment dislodging at the inception point most likely produced by the concentration of percolines resulting in an acceleration of the seepage at specific points (Perillo et al., 2005).
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In the case of rills, the longitudinal profile shows a maximum depth at the head but the course soon becomes shallower and only a minimum channelization is observed, being minimum at the water edge. The situation changes if several rills discharging groundwater converge into a single course, then course incision could be larger although discharge is seldom large enough to overcome a convex upward profile at the water edge. Grooves are formed in channel margins of the Bahı´a Blanca Estuary (Figure 1a) by groundwater discharge but enhanced by the percoline concentration due to crab burrows (Perillo et al., 2005). As groundwater fills the burrows, it spills out over the burrow border, often producing an indentation on the sill that helps concentrate the flow and generate the groove. As these crabs are an important food source of the White Croaker fish, on occasions the mouth of the burrow is enlarged (generating a crater) as the fish hits the surface when it traps a crab resulting in a greater groundwater concentration and larger spill over and, consequently, wider and deeper grooves. In this case, groundwater comes from the tidal filling of dense concentration of crab burrows at the associated marsh platform. While the various methods described are probably the basic mechanism for the initial development of courses on the lower tidal flats (hydroperiod = 360 days/ year), where tidal penetration occurs frequently, it is probably not the dominant mechanism of channel formation within salt marshes or mangroves at higher elevations. Higher (and older) areas may have reached an elevation that limits tidal penetration to only the highest tides (hydroperiod <120 days/year). As a consequence, sheet flows are rarer. For the case of courses in marshes, it has always been thought that they were inherited (and further modified) from the older tidal flats that they colonized. Some authors (Pethick, 1969, 1992; Pestrong, 1972; Pye, 1992) even suggested that channels were obliterated, rather than created in marshes. Yapp et al. (1917, in Allen, 2000) described the possible mechanism for the origin of courses in marshes based on the presence of hummocks formed by sediment retention due to the colonization of Glyceria (Puccinellia). Although Yapp et al. (1917) offered a valid explanation, it does not account for large gullies, creeks and channels. Perillo et al. (1996) proposed the first known mechanism for creek formation on a middle salt marsh in southern Argentina. Creeks develop from the interconnection of series of ponds enlarged by wave erosion of the pond walls as a consequence of the intense, direction-concentrated winds prevailing on Patagonia. Precreeks are finally connected to existing channels by cascading and seaward erosion as tides flood over the ponds. Further growth and maintenance of the creek depends on the intensity of the water exchange. If this exchange is poor, the reduction of soil salinity allows plant colonization which further reduces the exchange, enhancing sedimentation, and finally attaining the complete obliteration of the creek. A similar mechanism for gully and creek initiation was described for a freshwater marsh on the southern coast of the Rio de la Plata Estuary (Perillo and Iribarne, 2003b). In this case, the ponds are much smaller (Figure 6a and b) and formed by soil marsh depression both by organic matter compaction and by groundwater washing of interlayered sands. Interconnection between ponds and creek formation (Figure 6c) is produced by wall erosion by waves, water cascading and seaward
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(a)
(b)
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Figure 6 Formation of tidal creeks in a freshwater marsh. (a) View of the field of ponds along the southern coast of the Rio de la Plata, (b) view of a pond showing the variety of plants inside and outside the pond and (c) a threshold between two ponds. Cutting of the threshold interconnects the ponds as a step for the formation of a creek, (d) creek developed after the cutting of the threshold.
retreat. Once the ponds are connected among themselves and with the estuary, water flows through the creek driven by the microtide dominant in the estuary and also by storm surges that are frequent in the area. Although there is no mention in the literature of processes similar to those described in the previous paragraphs, there are examples observed by the author in other marshes, that is, Minas Basin (Figure 6d) that lead to the formation of creeks by pond connection. This idea is exactly the contrary of the development of channel pans proposed by Pethick (1974, 1992) where changes in the marsh conditions (i.e., lowering of the sea level, marsh disconnection, etc.) result in a
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smaller hydroperiod followed by vegetation growth and lateral erosion which may isolate parts of the course, specially those with meanders that become ponds. A fourth mechanism of creek and gully erosion found and extensively studied due to the interesting physical–biological interaction was observed in the Bahı´a Blanca Estuary (Perillo and Iribarne, 2003a,b). One of the most surprising and interesting findings is that creek formation can actually be a product of the intense action of crabs (Neohelice granulatus). In these settings, crabs first interact with a halophytic plant (Sarcocornia perennis) developing zones of high density of crab holes, which are then utilized by groundwater and tidal action to form tidal courses. When analyzing the spatial distribution of Sarcocornia, Perillo and Iribarne (2003a) discovered that the plants were mostly distributed in circles of up to 1.5–8 m in diameter, with the plants concentrated in a ring along the outer portion of the circle (Figure 7a). These rings vary in width from 0.5 to 1.5 m. The central part of the circle is an unvegetated salt pan, but it is densely excavated by the burrowing crab. Burrow density reaches from 40 to 60 holes/m2 (Figure 7b). The holes made by the crabs have a diameter of up to 12 cm (Figure 7c), and they reach up to 70–100 cm into the sediment (Iribarne et al., 1997; Bortolus and Iribarne, 1999). An interesting feature of the plant rings and their interaction with the crabs is the effect on the erosive process of the salt marsh in an estuary which is generally in an erosional stage. The formation of the Sarcocornia rings and the crab activity plays a major role in the erosion of the marsh transferring 1,380 m3 of sediment from a 270 ha marsh to the estuary (Minkoff et al., 2005, 2006; Escapa et al., 2007). In 2004 alone, more than 13% (183 m3) of the total sediment was exported indicating that the process keeps exporting more sediment every year in a pronounced exponential curve. The fact that the crab burrows are permanently inundated produces two effects. First, silty clay sediments (characteristic of this area) become partly loose and second, the water in the holes starts to migrate, breaking intercave walls and developing a groundwater stream that undermines soil. Under the pan surface, caves are formed where groundwater flows can be seen. A further stage in the development of the channel appears when soil surface fails. At this stage, the course has only a general structure filled with remnants of intercave walls (Figure 7c and d), around which water circulates (Perillo and Iribarne, 2003b). The final stage is reached when the walls are eroded after a number of tides and the channel presents smooth, low-slope banks (Figure 7d). Crabs can be found along the creek banks but no plants are observed. As a result, along the creeks, the original tidal flat moves landward at the expense of the salt marsh, and this mechanism has been described in the previous section. Application of a Cellular Automata model (Minkoff et al., 2006) that takes into account the various active interactions existing in the marsh even makes it possible to predict the changes in the geomorphology of the terrain and the formation and evolution of gullies and creeks. At the present time, a new mechanism is being studied in the Bahı´a Blanca Estuary related to the development of shallow ponds (Figure 7e) over tidal flats which may either be or not be connected to creeks. These ponds may result from a complex interaction originated in the differential resistance that biofilms may produce to the flat erosion by short-period wind waves.
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(a)
(b)
(c)
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Figure 7 (a) Aerial view of the Sarcocornia rings, (b) an example of a Sarcocornia ring showing the dense distribution of crab burrows at the center, (c) head of a creek previous to the collapse of the surface soil where the large number of burrows mark the area that will be collapsed, (d) head of a creek after the collapse, (e) creek partially flooded and (f ) aerial view of the distribution of small ponds over a tidal flat of the Bahı´ a Blanca Estuary.
6. C OURSE EVOLUTION The morphology of large rivers depends on many factors (i.e., tectonics, rock/sediment composition, slope, etc.). As size diminishes, local morphology plays an increasingly larger role in the valley and specifically in river channels.
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Tidal streams have similar influences. For instance, rills are controlled by surface sediment characteristics; the less compacted the material the deeper the rill. Also slope is important since rills are more lineal with higher slopes. In most cases, rills are deeper near their heads than at their mouth. In an evolving wetland, courses appear almost simultaneously with the deposition of sediments that forms a tidal flat (Wolanski, 2007). Rills, grooves and gullies are the first to appear and their persistence depends mostly on local factors such as some particular irregularity on the sediment surface or the depth of the incision. A deeper incision tends to occur in spring tides as low water level induces more erosion than neap low water. Once a course has been preserved through several tidal cycles, the possibility of it being erased by strong currents or wave activity diminishes proportionally with time. Thus, further evolution is now controlled by ebb and flood currents, the former being particularly important as they convey the water discharge from the overflat flooding. Although tidal flats can be preserved, soil stability commonly helps in establishing pioneer plants which, in a longer run, allows the formation of salt marshes and/ or mangroves. Levees tend to form along courses where sedimentation is predominant even in the case of tidal flats, being most common in marsh courses. However, they seldom occur along courses in erosive environments or when the concentration of suspended sediments is very low (<100 mg/l). The formation of levees further stabilizes the courses and allows their vertical growth. When there is no levee dissection, after some extraordinary tides or rainfalls, levees may act as dams allowing the formation of ponds even if there are no significant depressions. Ponds, when persistent in places, can be major controllers of the geomorphologic evolution of an intertidal area. Ponds are affected by evaporation and more prone to be colonized by benthic fauna (i.e., crabs, poliquetes, etc.), which prefers areas with higher water content as it is easier for them to make their burrows further deepening the pond by sediment subsidence. Also diatoms tend to concentrate there, and especially during warmer periods, their oxygen productivity can be large enough to dislodge sediment and keeps it floating (Figure 8). Wind action is a major factor in enhancing ponds although this process has been rather overlooked in the literature. Typical short-period waves form due to small fetch and depth have high steepness which makes them a very erosive process. In places where some wind directions are very frequent (i.e., Argentine Patagonia), ponds can be enlarged along the wind path. Wind, in these cases, could play an important role in defining water circulation over the tidal flat (the effect on marshes and mangroves should be smaller) and because of the relatively low water depth and velocities, wind shear transfer may easily change the pressure gradient induced by the tide or, at the beginning and end of the inundation time, by the topographic gradient. Creeks can be easily formed by the interplay of previous factors as described by Perillo et al. (1996). Levees are dissected by creeks and gullies but also by strong rain events and short-period, highly erosive waves. These processes help restore the exchange. Therefore, when considered as a whole, levees cannot grow indefinitely but must reach some kind of dynamic equilibrium based on tidal range, overflow velocities, suspended sediment concentration, plant species (involving all plant morphological characteristics affecting flow and sediment stability) and local geomorphology.
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Figure 8 Example of the mechanism in which diatoms, by strong oxygen production, can dislodge sediment on a tidal flat.
Moving further inland, courses are also elongated by the interaction with freshwater. Fine sediments react in a different way in freshwater than in salt water. In the former, mud tends to remain in suspension longer than in saltwater conditions, thus allowing better conditions for sediment erosion (Wolanski, personal communication). Several examples in Australia (Mulrennan and Woodroffe, 1998; Cobb et al., 2007) show that freshwater affects saline water mud and, during continental floods, erosion of the salt marshes and flats is larger than during regular tidal inundation (Figure 9). Freshwater is not only introduced into the system by
Figure 9 Aerial photograph showing the growth of saline tidal courses in the freshwater plains of the Mary River (Australia), seriously affecting the freshwater vegetation (courtesy of Eric Wolanski).
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continental drainage but also in the form of rain which, when intense, also produces strong surface sediment erosion (Mwamba and Torres, 2002) which when conducted by the geomorphology can easily erode the sediment surface further and generate or prolong tidal courses. Then, further evolution of these features becomes a tidal process.
7. SUMMARY Tidal courses are an essential part of coastal wetlands as they play a major role in water and nutrient exchange. However, their origin and evolution is still a matter of discussion due to the complexities of the dynamic processes associated with their initiation. Some factors such as the role of overland flows versus bankfull flow in the evolution of the courses are unknown. A major question is what actually controls course meandering on tidal flats. Contradictory examples about the meandering distribution can be given when the cases of the Anse d’Aiguillion (Eisma, 1997) and Bahı´a Blanca Estuary are compared. In the former, creeks in the higher flats are strongly meandering and become straight and sinuous in the low flats, whereas in Bahı´a Blanca Estuary, what occurs is exactly the opposite; sinuosity increases significantly for creeks and gullies along the margins of tidal channel while on the surface of the flats they are much less sinuous. A tidal course classification has been proposed to establish a unified description of these features and to avoid confusion. Based on geomorphologic information, this classification may evolve further by integrating other descriptors. Therefore, this classification is open to consideration and discussion.
ACKNOWLEDGMENTS Partial funding for researches that lead to the present review was provided by CONICET, Agencia de Promocio´n Cientifica y Tecnolo´gica and Universidad Nacional del Sur of Argentina and the National Geographic Society of USA. This chapter is also a contribution to the SCOR-LOICZ-IAPSO WG 122 “Mechanism of Sediment Retention in Estuaries.” I thank the constructive comments by Eric Wolanski.
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Barwis, J.H., 1978. Sedimentology of some South Carolina tidal-creek point bars, and a comparison with their fluvial counterparts. In: Miall, A.D. (Ed.), Fluvial Sedimentology. Canadian Society of Petroleum Geologists, pp. 129–160. Best, J.L., 1987. Flow dynamics at river channel confluences: implications for sediment transport and bed morphology. In: Recent Developments in Fluvial Sedimentology. The Society of Economic Paleontologists and Mineralogists. Tulsa, Special Publication 39, 389pp. Blondeaux, P., Seminara, G., 1985. A unified bar-bend theory of river meanders. J. Fluid Mech. 157, 449–470. Bortolus, A., lribarne, O.O., 1999. Effects of the burrowing crab Chasmagnathus granulata on a Spartina salt marsh. Mar. Ecol. Prog. Ser. 178, 78–88. Carson, M.A., LaPointe, M.F., 1983. The inherent asymmetry of river meander planform. J. Geol. 91, 41–55. Chapman, V.J., 1960. Salt Marshes and Salt Deserts of the World. Leonard Hill, London. Cobb, S.M., Saynor, M.J., Eliot, M., Eliot, I., Hall, R., 2007. Saltwater intrusion and mangrove encroachment of coastal wetlands in the Alligator Rivers Region, Northern Territory, Australia. Supervising Scientist Report 191. Supervising Scientist, Darwin NT. Collins, M.B., Ke, X., Gao, S., 1998. Tidally-induced flow structure over intertidal flats. Estuar. Coast. Shelf Sci. 46, 233–250. Dietrich, W.E., Dunne, T., 1993. The channel head. In: Beven, K., Kirkby, M.J. (Eds.), Channel Network Hydrology. John Wiley and Sons, Chichester, pp. 175–219. Duval, T.P., Hill, A.R., 2006. Influence of stream bank seepage during low-flow conditions on riparian zone hydrology. Water Resour. Res. 42, W10425. doi:10.1029/2006WR004861. Dyer, K.R., 1998. Estuaries: A Physical Introduction. John Wiley & Sons, Chichester, 195pp. Eisma, D., 1997. Intertidal Deposits: River Mouths, Tidal Flats, and Coastal Lagoons. CRC Press, Boca Raton, 525pp. Escapa, C.M., Minkoff, D.R., Perillo, G.M.E., Iribarne, O.O., 2007. Direct and indirect effects of burrowing crab activities on erosion of SW Atlantic Sarcocornia-dominated marshes. Limnol. Oceanogr. 52, 2340–2349. Fagherazzi, S., Bortoluzzi, A., Dietrich, W.E., Adami, A., Lanzoni, S., Marani, M., Rinaldo, A., 1999. Tidal networks 1. Automatic network extraction and preliminary scaling features from DTMs. Water Resour. Res. 35, 3891–3904. Fagherazzi, S., Furbish, D.J., 2001. On the shape and widening of salt marsh creeks. J. Geophys. Res. Oceans 106, 991–1005. Fagherazzi, S., Gabet, E.J., Furbish, D.J., 2004. The effect of bidirectional flow on tidal channel planforms. Earth Surf. Process. Landforms 29, 295–309. Findlay, S., 1995. Importance of surface-subsurface exchange in stream ecosystems: the hyporheic zone. Limnol. Oceanogr. 40, 159–164. French, J.R., Stoddart, D.R., 1992. Hydrodynamics of salt marsh creek systems: implications for marsh morphological development and material exchange. Earth Surf. Process. Landforms 17, 235–252. Gardner, L.R., 2005. A modeling study of the dynamics of pore water seepage from marsh sediments. Estuar. Coast. Shelf Sci. 62, 691–698. Gardner, L.R., Wilson, A.M., 2006. Comparison of four numerical models for simulating seepage from salt marsh sediments. Estuar. Coast. Shelf Sci. 69, 427–437. Garofalo, D., 1980. The influence of wetland vegetation on the tidal stream channel migration and morphology. Estuaries 3, 258–270. Ginsberg, S.S., Perillo, G.M.E., 1990. Channel bank recession in the Bahı´a Blanca Estuary, Argentina. J. Coast. Res. 6, 999–1010. Ginsberg, S.S., Perillo, G.M.E., 1999. Deep scour holes at the confluence of tidal channels in the Bahia Blanca Estuary, Argentina. Mar. Geol. 160, 171–182. Ginsberg, S.S., Perillo, G.M.E., 2004. Characteristics of tidal channels in a mesotidal estuary of Argentina. J. Coast. Res. 20, 489–497. Guilcher A., 1957. Morfologı´a litoral y submarina. Ediciones Omega, Barcelona. Hollins, S.E., Ridd, P.V., Read, W.W., 2000. Measurement of the diffusion coefficient for salt in salt flat and mangrove soils. Wetlands Ecol. Manage. 8, 257–262.
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Horton, R.E., 1945. Erosional development of streams and their drainage basins; hydrophysical approach to quantitative morphology. Bull. Geol. Soc. Am. 56, 275–370. Ikeda, I., Parker, G., Sawai, K., 1981. Bend theory of river meanders, part I, Linear development. J. Fluid Mech. 112, 363–377. Iribarne, O.O., Bortolus, A., Botto, F., 1997. Between-habitat differences in burrow characteristics and trophic modes in the southwestern Atlantic burrowing crab Chasmagnathus granulatus. Mar. Ecol. Prog. Ser. 155, 132–145. Istanbulluoglu, E., Tarboton, D.G., Pack, R.T., Luce, C., 2002. A probabilistic approach for channel initiation. Water Resour. Res. 38, 1325. doi:10.1029/2001WR000782. Kelly, R.P., Moran, S.B., 2002. Seasonal changes in groundwater input to a well-mixed estuary estimated using radium isotopes and implications for coastal nutrient budgets. Limnol. Oceanogr. 47, 1796–1807. Kirkbya, M.J., Bullb, L.J., Poesenc, J., Nachtergaelec, J., Vandekerckhove, L., 2003. Observed and modelled distributions of channel and gully heads – with examples from SE Spain and Belgium. Catena 50, 415–434. Kjerfve, B., 1978. Bathymetry as an indicator of net circulation in well mixed estuaries. Limnol. Oceanogr. 23, 816–821. Kjerfve, B., Shao, C.-C., Stapor Jr., F.W., 1979. Formation of deep scour holes at the junction of tidal creeks: an hypothesis. Mar. Geol. 33, M9–M14. Lathrop, R.G., Cole, M.B., Showalter, R.D., 2000. Quantifying the habitat structure and spatial pattern of New Jersey (U.S.A) salt marshes under different management regimes. Wetlands Ecol. Management 8, 163–172. Leopold, L.B., Wolman, M.G., 1960. River meanders. Geol. Soc. Am. Bull. 71, 769–793. Leopold, L.B., Wolman, M.G., Miller, J.P., 1964. Fluvial Processes in Geomorphology. W.H. Freeman, New York. Mallin, M.A., 2004. The importance of tidal creek ecosystems. J. Exp. Mar. Biol. Ecol. 298, 145–149. Marani, M., Lanzoni, S., Zandolin, D., Seminara, G., Rinaldo, A., 2002. Tidal meanders, Water Resour. Res. 38, 1225. doi:10.1029/2001WR000404. Marani, M., Belluco, E., D’Alpaos, A., Defina, A., Lanzoni, S., Rinaldo, A., 2003. On the drainage density of tidal networks. Water Resour. Res. 39(2), 1040. doi:10.1029/2001WR001051. May, M.K., 2002. Pattern and process of headward erosion in Salt Marsh Tidal Creeks. M.D. Thesis. Faculty of the Department of Biology East Carolina University. Minkoff, D.R., 2007. Geomorfologı´a y dina´mica de canales de mareas en ambientes intermareales. PhD Dissertation, Departamento de Ingenierı´a, Universidad Nacional del Sur, Bahia Blanca, 188pp. Minkoff, D.R., Escapa, C.M., Ferramola, F.E., Maraschin, S., Pierini, J.O., Perillo, G.M.E., Delrieux, C., 2006. A Cellular Automata model for study of the interaction between the crab Chasmagnathus granulatus and the halophyte plant Sarcocornia perennis in the evolution of tidal creeks in salt marshes. Estuar. Coast. Shelf Sci. 69, 403–413. Minkoff, D.R., Escapa, C.M., Ferramola, F.E., Perillo, G.M.E., 2005. Erosive processes due to physical – biological interactions based in a cellular automata model. Lat. Am. J. Sedimentol. Basin Anal. 12, 25–34. Minkoff, D.R., Perillo, G.M.E., 2002. Evolucio´n de canales de marea en una marisma de Bahı´a Blanca. III Taller de Sedimentologı´a y Medio Ambiente, Buenos Aires (abstract). Montgomery, D.R., Dietrich, W.E., 1992, Channel initiation and the problem of landscape scale. Science 255, 826–830. Montgomery, D.R., Dietrich, W.E., 1988. Where do channels begin? Nature 336, 232–234. Mulrennan, M.E., Woodroffe, C.D., 1998. Saltwater intrusion into the coastal plains of the Lower Mary River, Northern Territory, Australia. J. Environ. Manage. 54, 169–188. Mwamba, M.J., Torres, R., 2002. Rainfall effects on marsh sediment redistribution, North Inlet, South Carolina, USA. Mar. Geol. 189, 267–289. Novakowski, K.I., Torres, R., Gardner, L.R., Voulgaris, G., 2004. Geomorphic analysis of tidal creek networks, Water Resour. Res. 40, W05401. doi:10.1029/2003WR002722. Osgood, D.T., 2000. Subsurface hydrology and nutrient export from barrier island marshes at different tidal ranges. Wetlands Ecol. Manage. 8, 133–146.
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Perillo, G.M.E., 1995. Geomorphology and sedimentology of estuaries: An introduction. In: Perillo, G.M.E. (Ed.), Geomorphology and Sedimentology of Estuaries, Development in Sedimentology Vol. 53, Elsevier Science BV, Amsterdam, 1–16. Perillo, G.M.E., Iribarne, O.O., 2003a. New mechanisms studied for creek formation in tidal flats: from crabs to tidal channels. EOS Am. Geophys. Union Trans. 84, 1–5. Perillo, G.M.E., Iribarne, O.O., 2003b. Processes of tidal channels develop in salt and freshwater marshes. Earth Surf. Process. Landforms 28, 1473–1482. Perillo, G.M.E., Minkoff, D.R., Piccolo, M.C., 2005. Novel mechanism of stream formation in coastal wetlands by crab–fish–groundwater interaction. Geo. Mar. Lett. 25, 124–220. Perillo, G.M.E., Piccolo, M.C., 1999. Geomorphologic and physical characteristics of the Bahı´a Blanca Estuary, Argentina. In: Perillo, G.M.E., Piccolo, M.C., Pino Quivira, M. (Eds.), Estuaries of South America: Their Geomorphology and Dynamics.Springer-Verlag, Berlı´n, pp. 195–216. Perillo, G.M.E., Ripley, M.D., Piccolo, M.C., Dyer, K.R., 1996. The formation of tidal creeks in a salt marsh: new evidence from the Loyola Bay Salt Marsh, Rio Gallegos Estuary, Argentina. Mangroves Salt Marshes 1, 37–46. Pestrong, R., 1965. The development of drainage patterns in tidal marshes. Stanford Univ. Publ. Earth Sci. 10, 1–87. Pestrong, R., 1972. Tidal-flat-sedimentation at Cooley Landing, southwest San Francisco Bay. Sediment. Geol. 8, 251–288. Pethick, J.S., 1969. Drainage in tidal marshes. In: Steers, J.R. (Ed.), The Coastline of England and Wales, third ed. Cambridge University Press, Cambridge, pp. 725–730. Pethick, J.R., 1974. The distribution of salt pans on tidal salt marshes. J. Biogeogr. 1, 57–62. Pethick, J.S., 1992. Saltmarsh geomorphology. In: Allen, J.R.L., Pye, K. (Eds.), Saltmarshes, Morphodynamics, Conservation and Engineering Significance.Cambridge University Press, Cambridge, pp. 41–62. Pye, K., 1992. Saltmarshes on the barrier coastline of North Norfolk, eastern England. In: Allen, J.R.L., Pye, K. (Eds.), Saltmarshes, Morphodynamics, Conservation and Engineering Significance. Cambridge University Press, Cambridge, pp. 148–181. Pye, K., French, P.W., 1993. Erosion and Accretion Processes on British Salt Marshes, Vol. 1. Introduction: Saltmarsh Processes and Morphology. Cambridge Environmental Research Consultants, Cambridge. Rinaldo, A., Belluco, E., D’Alpaos, A., Feola, A., Lanzonni, S., Marani, M., 2004. Tidal networks: form and function. In: Fagherazzi, S., Marani, M., Blum, L.K. (Eds.), The Ecogemorphology of Tidal Marshes.American Geophysical Union, Washington, DC, pp. 75–91. Rinaldo, A., Fagherazzi, S., Lanzoni, S., Marani, M., Dietrich, W.E., 1999. Tidal networks 2. Watershed delineation and comparative network morphology. Water Resour. Res. 35, 3905–3917. Seminara, G., Tubino, M., 1992. Weakly nonlinear theory of regular meanders. J. Fluid Mech. 244, 257–288. Shao, C,-C., 1977. On the existence of Deep Holes at Tidal Creek Junctions. MS Thesis, University of South Carolina, Columbia, SC, 31pp. Stark, C.P., 1991. An invasion percolation model of drainage network evolution. Nature 352, 423–425. Tucker, G.E., Arnold, L., Bras, R.L., Flores, H., Istanbulluoglu, E., So´lyom, P., 2006. Headwater channel dynamics in semiarid rangelands, Colorado High Plains, USA. GSA Bulletin 118, 959–974. Wolanski, E., 2007. Estuarine Ecohydrology. Elsevier, Amsterdam, 157pp. Yapp, R.H., Johns, D., Jones, O.T., 1917. The salt marshes of the Dovey Estuary. Part II. The salt marshes. J. Ecol. 5, 65–103.
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C H A P T E R
7
H EAT E NERGY B ALANCE IN C OASTAL W ETLANDS Marı´a Cintia Piccolo
Contents 1. Introduction 2. Mid-Latitudes 3. Low Latitudes 4. High Latitudes 5. Summary Acknowledgments References
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1. INTRODUCTION By virtue of its unusually high specific and latent heat, water stands apart from other substances in its thermal properties for supporting the life cycles of a great number of plant and animal species that depend on wetlands. Wetland plants alone exhibit a wide diversity of growth forms including emergent plants, submerged plants, floating-leafed plants as well as a combination of these leaf forms within the same species (Sculthorpe, 1967). Given this variety, it is useful not only to understand how these organisms have adapted to energy budgets, but also their role in contributing to the distribution of energy flows at the ecosystem level. Relatively few studies have been conducted on heat exchanges in coastal wetlands. Yet, an understanding of how heat is transferred across air–sea–soil interfaces is fundamental to predicting, for example, how coastal wetlands will respond to global climate change. Coastal wetlands develop particularly steep chemical and hydrological gradients as a result of their position between continents and the ocean. Unlike vertical fluxes that dominate exchanges in upland ecosystems, the exchange of matter and energy in coastal wetlands is complicated by strong horizontal fluxes, particularly those driven by water movement. This chapter reviews studies that have contributed to understanding of how heat energy is transferred across the different levels of these fluid and solid interfaces, and how this energy affects different biological and chemical processes. Coastal Wetlands: An Integrated Ecosystem Approach
2009 Published by Elsevier B.V.
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Studies of the energy budget over an earth surface elucidate how solar energy is locally redistributed to create a particular microclimate (Kjerfve, 1978). The heat budget equation (Oke, 1978) R N = Q H þ QB þ Q E þ Q A
ð1Þ
establishes that at any moment in time the available energy at the earth’s surface (net radiation, RN) must be equivalent to a combination of convective exchange to or from the atmosphere (sensible, QH, and latent heat, QE), conductive flux to or from the soil (QB), and incoming or outgoing advective flux (QA). Figure 1 shows the components of the heat budget equation in an idealized bare tidal flat. The interface or boundary between the water and the air is dynamic. Matter and energy are continuously being transferred across the air–sea interface in both directions. Air either gains or loses heat from the water depending on the temperature difference between the water surface and the overlying air. Water evaporates to contribute to atmospheric moisture and atmospheric water vapor condenses to form fog and clouds, and eventually precipitation. Vegetated wetlands also receive and lose energy by radiation, conduction, convection, and evaporation. However, water lost by plant transpiration must be separated from salts in seawater, a process that comes at great cost to the plant (Teal and Kanwisher, 1970). Few species are well adapted to do this, which explains in part the low species richness of vascular plants in coastal wetlands. Horizontal exchanges of matter and energy are affected by tides and wind. Of these, tides behave more predictably (Perillo and Piccolo, 1991). They are responsible for the ebb and flow of water in all of the major groups of wetlands (seagrass meadows, mudflats, marshes, and mangroves). Winds, on the other hand, are less predictable (except for sea breezes), and few studies focus on how wind affects the energy budget of coastal areas (Leal and Lavı´n, 1998; Castro et al., 2003). In any case, advective fluxes may originate with tides and/or with wind (Figure 1b). Furthermore, winds may serve to amplify or dampen the effects of astronomic tides. (b)
Atmosphere
(a) RN
QH
Soil
QE
QA
QB
RN
Soil
Atmosphere QH
QE
Water
QA
QA QB
Figure 1 Components of the heat balance equation at a bare tidal flat: (a) low tide and (b) high tide.
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Annual heat flux cycles in water bodies usually follow the seasonal fluctuations of incident solar radiation. Annual mean radiation is used to express the main source of heat energy, and has been measured in several studies to determine both the magnitude and the seasonality in coastal wetlands (Hovel and Morgan, 1999; Bianciotto et al., 2003; El-Metwally, 2004; Jacobs et al., 2004; Paulescu et al., 2006). Radiation on tidal flats is of particular interest because of its significance to ecological processes. In a study of sediment temperatures on the Bay of Fundy (Canada), Piccolo et al. (1993) confirmed that radiation and tides are the drivers of thermal behavior over and within tidal flats. Gould and Hess (2005) focused on environmental radiation related to tides. They measured sediment exposure rate from environmental radiation on tidal flats using a high-pressure ion chamber so the shielding effects of the tidal cycle could be evaluated. They derived a theoretical model to predict the behavior of exposure rate as a function of time. In addition, they developed an empirical formula to calculate the total exposure on a tidal flat that requires measurements of only the slope of the tidal flat and the exposure rate when no shielding occurs (Gould and Hess, 2005). The formula and the model can be applied to biological studies where radiation exposure is needed. Radiation can be used to estimate other parameters of the heat balance equation. For example, Geostationary Operational Environmental Satellitederived estimates of radiation to predict daily evapotranspiration (ET) in Florida wetlands (USA) with the Penman–Monteith, Turc, Hargreaves, and Makkink models (Jacobs et al., 2004). These estimates agreed well with ET measured with an eddy correlation system. Also incoming radiation plays a major role in chlorophyll production in plants in general; higher radiation induces a reduction in the chlorophyll production rate for each of the studied species. With regard to incoming short-wave radiation (Qsw), linear relationships have been found between daytime incoming radiation and both net radiation [RN = 0.73, Qsw 13.45 (W/m2)] and reflected radiation [QR = 0.079, Qsw þ 3.3 (W/m2)] over a Spartina alterniflora salt marsh during the summer (Crabtree and Kjerfve, 1978). They found that, on the average, net radiation was 70% and reflected radiation was 9% of incoming radiation. In spite of potentially damaging effects of Qsw, Costa et al. (2006) found no evidence of differential sensitivity or resilience to UV-B radiation between Salicornia species from low-mid-latitudes and a high-latitude population in the Americas. For many wetlands, ET is the major component of water loss. When considered as its energy equivalent, latent heat flux, evaporation is a major energy sink (Wessel and Rouse, 1993; Souch et al., 1996). Yet despite numerous studies, evaporation from wetlands is little understood (Lafleur, 1990) and detailed studies of the physical processes involved are geographically restricted (Souch et al., 1996). The latent heat flux term in Equation (1) is second only to the radiation flux. Most of the radiation energy is used for evaporation. In spite of that, the hydrologic implications may be minor because water loss from coastal wetland sediment is typically and often replaced by precipitation or by infiltration of flooding estuarine water (Harvey and Nuttle, 1995).
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Soil properties play an important role in plant composition, productivity, and zonation in coastal ecosystems because plant species are differentially tolerant to salinity and soil saturation (Adams, 1963; Pennings et al., 2005; Wang et al., 2007). Variations in salinity and temperature in the intertidal zone restrict the distribution and abundance of marine organisms (Johnson, 1967). In tidal flats, species richness is higher where rates of salinity change are low (Sanders et al., 1965). In sandy intertidal areas where salinities fluctuate widely, the maximum number of species and individuals is small. This feature is not caused by sediment salinity alone because temperature, oxygen content, and other environmental factors also vary by intertidal zone. There is a strong relationship between evaporation and sediment salinity. Removal of plant cover, for example, induces higher soil temperature, increases pore water salinities, and lowers water content, most likely due to the greater sun exposure and higher evaporation (Whitcraft and Levin, 2007). On sandy beaches, Johnson (1967) found that an increase in evaporation caused an increase in the soil salinity in the first 20 cm of depth. Many of the effects of salt on plants occur as a result of water stress. Tolerance of plants to saline soils is due in part to biophysical, morphological, and biochemical adaptations. The narrow leaves characteristic of high marsh species may be an adaptation to help regulate leaf temperature in times of low latent cooling (Maricle et al., 2007). Latent heat fluxes of coastal wetlands play a significant role in plant zonation. Surface sensible and latent heat fluxes can be used to predict coastal storm development and precipitation. Both affect wetland development. Surface latent flux provides a direct source of moisture needed for precipitation, while sensible heat flux can affect the stability of the storm environment, thereby modulating the timing and amount of precipitation (Persson et al., 1999). Data from any meteorological station placed almost anywhere across a wetland provide representative estimates of evaporation when atmospheric conditions are relatively homogeneous. Individual patches of vegetation in a wetland do not influence overlying atmospheric conditions significantly and evaporation can be estimated using the well-known formula of Penman–Monteith (Gavin and Agnew, 2003). However, Acreman et al. (2003), using the eddy correlation method to calculate evaporation from two types of wetlands, wet grassland and reed beds, in southwest England, demonstrated that the evaporation calculated by the Penman potential method did not represent actual evaporation. General agreement as to the best method of estimating evaporation on different coastal wetlands is still to be achieved. Because different climates generate different types of coastal wetlands and, therefore, are the product of different heat energy balances, this chapter reviews heat balance studies on a diversity of wetland types, and with a variety of methods. Two facts must be highlighted. First, there are very few relevant studies and second, most studies were performed on tidal flats in estuaries, with very little work on other types (i.e., salt marshes and mangroves). The review is presented in three sections to emphasize climate difference in the wetland studied. Since most of the investigations were carried out at mid-latitudes, the review starts in this region.
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2. MID -LATITUDES Tidal flats are affected by tides, winds, and sun radiation exposure. Heat balance researches made at temperate tidal flats are generally based on the bulk formulation (Hsu, 1978; Smith, 1981; Smith and Kierspe, 1981; Vugts and Zimmerman, 1985; Harrison and Phizacklea, 1985; Piccolo and Da´vila, 1993; Piccolo et al., 1993, 1999; Beigt et al., 2003; Beigt and Piccolo, 2003). Two different methods to estimate the heat balance on tidal flats will be presented here. Both methods highlight the significant influence of the tidal forces on heat fluxes. For instance, the annual heat exchanges that occur at an estuarine tidal flat in the Bahı´a Blanca Estuary, Argentina were studied (Beigt, 2007; Beigt et al., 2008). Heat fluxes were analyzed across the water–atmosphere and the sediment–atmosphere interfaces at high and low tide, respectively. Different bulk aerodynamic formulas were used to estimate the radiative and turbulent fluxes from available meteorological and oceanographic data. Net radiation (RN) was determined from incident solar radiation and temperature data using (Evett, 2002) RN = Rsi ð1 Þ L " þ L #ðW=m2 Þ
ð2Þ
where Rsi is the incident solar radiation, is the albedo, L" is the terrestrial longwave radiation [L" = "s T4s ], L# is the atmospheric long-wave radiation, "s is the surface emissivity, is the Stefan–Boltzmann’s constant (W/m2/K4), and Ts is the surface temperature (K) (water or sediment temperature, depending on tidal stage). Atmospheric long-wave radiation (L#) is assessed using two different equations depending on temperature data. The Swinbank (1963) equation is used when temperatures are over 0C, while the Monteith (1973) equation is used for temperatures lower than 0C and higher than 5C. Soil heat flux (QB) across the surface layer was determined from temperature data using the usual Fourier equation (QB = (DT/Dz), Oke, 1978), where T is the sediment temperature (K), z is the depth (m), is the thermal conductivity [ = KSC (W/m/K)], KS is the thermal diffusivity (m2/s), and C is the heat capacity (J/m3/K). Sensible heat flux was estimated by means of two different equations depending on the tidal stage. During tidal flat inundation, sensible heat flux is assessed using (Kantha and Clayson, 2000; Zaker, 2003) QH = cp ðUa Us ÞCH ðTw Ta ÞðW=m2 Þ
ð3Þ
where is the air density (kg/m3), cp is the specific heat of the air (J/kg/C), Ua is the wind speed at height z, Us is the wind speed at the water surface (m/s) (zero for a stationary surface), CH is the heat exchange coefficient (dimensionless), Tw is the temperature at the water surface (C), and Ta is the air temperature (C). The heat exchange coefficient (CH) for the water–atmosphere interface was taken from Friehe and Schmitt (1976).
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During sediment exposure to atmospheric conditions, sensible heat flux was estimated by (Evett et al., 1994; Evett, 2002) QH = cp ðTs Ta ÞDH ðW=m2 Þ
ð4Þ
where DH(= k2U[ln(z/z0H)]2) is the heat exchange coefficient (Kreith and Sellers, 1975; Ma et al., 2003) (m/s), k is the von Ka´rma´ n’s constant, and z0H is the roughness length for sensible heat flux (m) taken from Kreith and Sellers (1975). Latent heat flux across the sediment–atmosphere and water–atmosphere interfaces was estimated by the Penman–Monteith equation that is the potential evaporation rate, namely, the evaporation rate that occurs when water availability is not limiting (Wallace and Holwill, 1997). Although it is usually applied in agronomical studies of vegetated and bare soils, it has also been applied to salt marshes (Hughes et al., 2001) and tidal flats (Harrison and Phizacklea, 1985). Taking into account that the sediments of the studied tidal flat are permanently saturated, assessing the potential evaporation was considered correct (Beigt, 2007). The authors estimated that annually, nearly 5,978 MJ/m2 of heat entered the tidal flat as incident solar radiation. Of this, only 2,954 MJ/m2 (49.4%) remained as available energy (RN). Winds and tides helped to add heat to the ecosystem (1,301 MJ/m2). The annual budgets of sensible and soil heat fluxes showed that both processes provided heat energy to the tidal flat surface. Indeed, an amount of 947 MJ/m2 was transferred to the surface as sensible heat, while the annual budget of soil heat flux indicated an upward heat transfer of 25.2 MJ/m2. The total energy that entered the tidal flat was balanced by an equal heat loss (5,227 MJ/m2 = 2,127 mm) as evaporation (Beigt, 2007; Beigt et al., 2008). Surface heat fluxes through the air–sea interface for the coastal water of Kuwait was estimated by Sultan and Ahmad (1994) using the bulk formulas such as Beigt (2007) and showed similar annual behavior. Even though Kuwait is representative of an arid region, both investigations show that a nocturnal inundation generally heats the tidal flat sediment (previously cooled by long-wave emission), causing an upward circulation of sensible heat. On the contrary, a tidal inundation at midday or early afternoon usually cools the sediment, with the resultant flow of sensible heat from the air to the tidal flat. The horizontal transport of heat or advection causes the addition (or subtraction) of energy to (from) an ecosystem. The most common agent of this process is wind; however, tide must also be considered when studying a tidal flat. Tidal energy generally acts as an “energy subsidy” to the coastal ecosystem, increasing its productivity as it increases the amount of energy which is capable of being converted to production (Odum, 1975). Advective heat flux is estimated as the residual energy from the heat budget equation [Equation (1)]. The total advective flux is then divided according to the tidal height into two different fluxes (“advective flux at low tide” and “advective flux at high tide”). The former is developed by winds, while tide is considered to be the main agent during tidal inundation. Atmospheric and tidal conditions regulate the heat exchanges. Tidal
217
Heat Energy Balance in Coastal Wetlands
inundation affects the direction and magnitude of sensible and soil heat fluxes. The 2003 annual heat budget showed that net radiation, advective, sensible and soil heat fluxes provided heat to the tidal flat surface and the most important heat fluxes were net radiation and latent heat. They are followed (in order of magnitude) by advective and sensible heat fluxes and finally by the soil heat flux (Beigt et al., 2008). Few heat balance studies were carried out on salt marshes. Teal and Kanwisher (1970) calculated the energy balance for the plants growing in a marsh on Cape Cod. They found that leaf temperature was well coupled to air temperature. If there were no evaporation of water from the leaves, they would have been from 3.6C to 9.2C above air temperature when heat gain equalled heat loss. Some of their results are shown in Table 1, where is the Bowen ratio (= QS/QE). Latent heat fluxes were always greater than sensible heat ones. The same behavior is found in most of the heat balances in coastal wetlands (Vugts and Zimmerman, 1985; Rouse, 2000; Beigt et al., 2008). The net balance between ET and rainfall infiltration is believed to be important in controlling soil salinity, particularly in the less frequently flooded high marsh zone. To elucidate the biophysical effects of drought and salinity on the interception and dissipation of solar energy in estuarine grasses, Maricle et al. (2007) studied leaf energy budget of 13 species. They found that latent heat loss decreased by as much as 65% under decreasing water potential (a measure of the ability of a substance to absorb or release water relative to another substance), causing an increase in leaf temperature of up to 4C. Consequently, radiative and sensible heat losses increased under decreasing water potential. Sensible heat flux increased as much as 336% under decreasing water potential. Latent heat loss appeared to be an important mode of temperature regulation in all species and sensible heat loss appeared to be more important in high marsh species than in low marsh ones (Maricle et al., 2007). In general, heat budget estimates showed similar annual patterns (Sultan and Ahmad, 1994; Roads and Betts, 2000; Hughes et al., 2001; Rutgersson et al., 2001; Finch and Gash, 2002). Another method to calculate the heat budget of a tidal flat area indirectly from downstream observations of temperature and horizontal velocity in a tidal course was presented by Onken et al. (2007). The advective heat flux (Qa) in tidal channels is monitored and then the heat excess or deficit for the catchment area is calculated by integral methods. Instead of using the bulk formulation, a relationship between the velocity and the volume flux is established. The heat budget of the upstream region is then determined by integrating the heat flux over one tide (Qtide). The Table 1 Heat fluxes (W/m2) measured in a Spartina alterniflora salt marsh on cape code Date 30 30 28 10
August August September November
RN
QE
QH
= QH/ QE
425.353 439.299 320.758 355.623
599.678 522.975 278.920 355.623
104.595 97.622 83.676 104.595
0.17 0.19 0.30 0.29
Source: Modified from Teal and Kanwisher (1970).
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heat budget of a water volume fixed in space is simply defined as the sum of the heat input through its boundaries and the production of thermal energy inside the volume (Onken et al., 2007). The direction of the heat flux components between the tidal flat and the associated channels during all the different tidal stages is described in Figure 2. At low tide, the bottom and the atmosphere exchange heat in terms of the following vertical heat flux Qat = Qsw þ Qlw þ QE þ QH
ð5Þ
where Qsw is the short-wave radiative heat flux and Qlw is the long-wave radiative heat flux. The bottom heat content changes due to atmosphere–soil interaction (Figure 2d). The advective flux in the channel is zero because there are no currents (slack water). During flood tide, there is a positive flux of heat toward the tidal flat by means of the advective flux Qa (onshore flow). At the same time, the heat content of the water column is modified by the interaction with the atmospheric flux Qat, and the heat exchange Qb between the substrate and the water (Onken et al., 2007). The high water condition (Figure 2b) is also characterized by zero heat advection but the water temperature will change due to Qat and Qb fluxes. The heat gain (or loss) Qtide of the water over one tidal cycle from low tide to low tide was determined by measuring the terms in Equation (6) Z tþ Z tþ low low Qtide = Qa dt = ðQat þ Qb Þdt ð6Þ tlow
tlow
(a)
Q at
Qa Qb (b)
Ebb Q at
Qb (c)
High Q at
Qa Qb (d)
Flood Q at
Low
Figure 2
Heat flux components on different tide stages (modified from Onken et al., 2007).
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where t is the time, tlow indicated the time of the low tide, and tþlow the time of the subsequent one. Onken et al. (2007) developed in addition a simple model which can be used to determine the integral bottom heat flux of the tidal flats. An analytical estimate suggests that the sign of the budget is controlled by the tidal prism and the length of the drying period of the flats in the upstream region. This method presents another option for a good estimate of the heat balance equation without using the bulk formulation. Other studies, such as heat budget investigations performed in Pauatahanui Inlet, a small (3.5 km2) New Zealand coastal inlet, indicates that the balance is essentially between solar and long-wave radiation, evaporation and advective heat exchange with the coastal waters (Heath, 1977). The role of the mudflat in the heat balance is only minor. The temperature at the entrance to the inlet exhibits strong tidal fluctuations resulting from exchange with the coastal waters and a diurnal inequality produced by interaction of solar heating with the tidally controlled surface area and volume. The simulated temperature record for a month, calculated from a heat budget equation exhibits the effect of the tidal/solar interaction in producing a 14.75-day pulse, variable diurnal inequality, and the generation of high-frequency components (Heath, 1977). Several studies in different coastal environments related to heat balance equation focus on temperature fluctuations and/or applied numerical models. The thermal behavior of the air, water, and sediments over a tidal flat was studied at Starr Point, Minas Basin, Bay of Fundy, Canada in July 1989 (Piccolo et al., 1993). Temperature in the intertidal sediments showed rapid changes which occur principally during tidal inundation. Vertical gradients of 0.5 102C/m were found in the upper 0.25 m layer. The presence of large populations of Corophium volutator increased thermal diffusivity because of their vertical migration. Therefore, the heat was distributed faster and through a greater depth. Vugts and Zimmerman (1985) predicted daily mean water temperatures with heat balance calculations of the tidal flat areas of the Dutch Wadden Sea. The daily heat balance interacts with the tidal cycle, resulting in a 15-day periodicity in the water temperature as well in the bottom temperature. They showed that with a simple model and some measured bulk parameters, it is possible to predict daily mean water temperatures from simple weather data measured at a nearby coastal station. The traditional formulation of the SWIFT2D model has been applied to numerous estuaries, bays, and harbors throughout the world. Swain (2005) made modifications to expand SWIFT2D for applicability to shallow coastal wetlands. These modifications include the representation of spatially and temporally varying rainfall and ET, wind sheltering owing to effects of emergent vegetation, and changes in frictional resistance with depth. Dietrich et al. (2004) presented a model-based method of determining the surface fluxes of heat and freshwater in the near-shore coastal waters. The new method determines the fluxes as a residual within the framework of physically realistic and natural boundary conditions on the sea surface temperature and sea surface salinity. On the basis of a balance model of the energy by surface waves in a coastal zone and experimental data about surface flows in shallow and deepwater zones, Panin et al. (2006) developed a model of the heat–mass exchange of a coastal zone
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reservoir with the atmosphere. The model allows the calculation of the values of the energy–mass exchange in the atmosphere boundary layer and the corresponding interaction characteristics based on standard micrometeorological information. For the construction of the model and its verifications they used the direct measurement data (eddy covariance) of the momentum, heat and humidity turbulent fluxes, as well as the surface wave characteristics and the main microcharacteristics of air and water. The model shows the intensification of the processes in a coastal zone in comparison with an open sea and allows the determination of the size intensification of flows at different distances from the coast.
3. LOW LATITUDES Estimation of the heat budget in low latitudes (tropical and subtropical) is performed as it is in mid-latitudes. Tropical wetlands are characterized by an increase in radiation energy and dense vegetation. Salt marshes are ideal to study plant community patterns. Species interactions during colonization of bare patches are different than those found in dense vegetation. The metabolic process of the plant communities exhibit different rhythms of intensity. They are regulated by variations in environmental factors such as light and temperature. Another important vegetation parameter such as photosynthesis is known to be temperature sensitive (Hargrave, 1969; Gallagher and Daiber, 1973). Since temperature affects respiration rates, an exogenous daily rhythm in respiration in salt marshes and mangroves would be expected in response to temperature cycles (Gallagher and Daiber, 1973). The presence of plants affects ecosystem-level processes such as hydrology, sedimentation rate and nutrient cycling (Bertness, 1988; Whitcraft and Levin, 2007). Plant cover is a fundamental feature of many coastal marine and terrestrial systems and controls the structure of associated animal communities. Studying the impact of shading in salt marshes, a relationship between temperature, salinity, water content and macrofaunal density and diversity was determined by Whitcraft and Levin (2007). Increases in temperature and salinity and decreases in water content for Salicornia virginica were correlated with decreased macrofaunal density. Although heat balance studies are important to determine these temperature variations in vegetated ecosystems, no specific measurements were found in the literature. On the other hand, estuaries in arid tropical regions differ significantly from their temperate and wet tropical counterparts. First, river discharge into the estuaries is often highly seasonal with very large flows in the wet season being followed by 5–10 months of negligible discharge. The second difference is that large areas of salt marshes, mangrove swamps and salt flats (where annual evaporation greatly exceeds annual precipitation) often fringe these tropical arid estuaries (Ridd and Stieglitz, 2002; de Silva Samarasinghe and Lennon, 2004) and they usually becomes hypersaline for much of the year. Evaporation plays an important role in concentrating salts and nutrients in soils and groundwater in estuarine wetlands. This is particularly true in zones
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of less frequent tidal inundation where the soil salinity depends on the balance between “evapoconcentration” of tidally supplied salts and rainfall or groundwater flushing (Hughes et al., 2001). The more arid the climate, the more extreme this effect becomes. For example, large areas of mangrove-fringed salt flats occur in the dry or mainly dry tropics. In contrast, salt marshes in temperate Australia seldom have unvegetated zones (Saenger et al., 1977). ET estimates are often the weak link in wetland water and solute balance modeling (Hughes et al., 2001). Latent heat flux is a major variable to calculate for salinity changes estimates over the mangroves. The plant zonation in salt marshes is a consequence of local variation in soil patch salinities (Bertness, 1991). The different soil salinities is due to tidal flooding, annual variation in rainfall, ET and small-scale topographic features which influence the drainage. Evaporation and ET increase the salinity of the swamp soils. A typical value of evaporation over open water in tropical areas is 5 mm/day. Wolanski et al. (1980) and Wolanski and Ridd (1986) calculated evaporation rate over mangroves at a rate of 2 mm/day. For salt flats, Hollins and Ridd (1997) estimated a monthly average evaporation rate of 2 mm/day with peak rates of 4–5 mm/day during spring tides (when the flats are saturated) falling to less than 1 mm/day when the salt flats form a hard surface crust. The salinity rate of change due to this effective evaporation rate E (Ridd and Stieglitz, 2002) is given by @S ES = @t h
ð7Þ
where S is the salinity, E is the evaporation and h is the depth. Field data from five arid estuaries fringed by mangroves and salt flats indicate that where a large area of salt flats and mangroves extends over the whole length of an estuary, the estuary becomes completely inverse with salinity rising up to 55 within a couple of months (Ridd and Stieglitz, 2002). The estuarine evaporation rates due to the presence of salt flats and mangroves cause a rapid increase in salinity. The persistence of fresher water in the upper reaches of this type of estuaries is likely to affect mangrove species assemblages (Ridd and Stieglitz, 2002). Precipitation has a significant importance in determining the salinity of the soil in coastal wetlands. Mondal et al. (2001) developed a multiple linear and nonlinear regression model to predict topsoil salinity (S) for both moderately saline and saline soils by using daily rainfall (P) and evaporation (E) as independent variables. The prediction level was not significantly improved with a nonlinear model; therefore, they suggest using the following linear one: S = 1:29077 0:49831P þ 1:31230E
ð8Þ
An important contribution to estimate the factors controlling the surface energy budget in coastal wetlands was made by Shoemaker et al. (2005). Changes in heat energy stored within a column of wetland surface water was calculated because this variable is a considerable component of the surface energy budget; a feature that is demonstrated by comparing changes in stored heat energy with net radiation at
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seven sites in the wetland areas of southern Florida, including the Everglades. Using the following simplified surface energy budget equation for wetlands RN ðW þ Gveg Þ = QE þ QH
ð9Þ
where Gveg is biomass storage (heat energy storaged in the vegetation) and W is the change in heat energy stored in wetland surface water. The difference between Equations (1) and (9) is that the advection and soil heat flux (QA þ QB) terms were replaced by the energy stored term in wetland surface. The biomass energy storage term was included to add the effect of the vegetation. This method estimates changes in stored heat energy that overcome an important data limitation, namely, the limited spatial and temporal availability of water temperature measurements, because it is assumed that a change in surface water temperature reflects a change in stored heat energy (Edinger et al., 1968). The new method was based on readily available air temperature measurements and relies on the convolution of air temperature changes with a regression-defined transfer function to estimate changes in water temperature. The convolution-computed water temperature changes are used with water depths and heat capacity to estimate changes in stored heat energy within the Everglades wetland areas. These results can probably be adapted to other humid subtropical wetlands characterized by open water, seagrass and several vegetation community types. The final discrete form of the transfer function and convolution integral, used to compute mean vertical water temperature changes and, ultimately, changes in heat energy stored in a column of wetland surface water, takes the form (Shoemaker et al., 2005) DTwi =
M X kex t j=0
hij
kex t
ehij DTaij ; j = 1; 2; 3; . . . M
ð10Þ
where kex is a proportionality constant that describes the rate at which water temperature responds to heat exchange processes (Edinger et al., 1968). The coefficient represents the fraction of air temperature change (DTa) that eventually causes an equivalent water temperature change (DTw) and is less than or equal to 1.0, and M is the historical time step discretizing the surface water’s thermal memory. It is important to point out that some authors incorporate not only the characteristic wetland heat energy stored in the vegetation, but also that stored in the fauna. In the literature several studies on the energy flow or trophic level production of coastal wetlands discuss the subject (Smalley, 1960; Teal, 1962; Wolff et al., 2000).
4. H IGH LATITUDES High-latitude (subpolar and polar) wetlands are underlain by ice-rich permafrost, which helps maintain wetland systems and also imparts special characteristics to their energy and water balances. In North America, components of the radiation balance decrease poleward (1.8 W/m2/latitude), whereas the poleward rate of decrease of temperature (i.e., between 50latitude and 65latitude: 1C/latitude)
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and precipitation lessens (i.e., between 50latitude and 65latitude: 22 mm/latitude) (Rouse, 2000). High latitudes are characterized by large annual changes in solar input. Albedo decreases strongly from winter, when the surface is snow-covered, to summer, especially in nonforested regions such as Arctic tundra and boreal wetlands. A primary characteristic of wetlands, in permafrost regions, is that the ice-rich frozen layer inhibits vertical water losses so that ponded water can persist through much of the summer. Components of the radiation balance tend to decrease with increasing latitude (Rouse, 2000; Rouse et al., 2004). During the 4-month summer of a high sub-Arctic wetland, net radiation is large and the latent heat flux dominates the energy cycle (Rouse, 2000). In winter, which typically lasts a minimum of 7 months, there is almost no evaporation but there is sublimation loss. In winter, heat loss from the ground approximately balances negative net radiation. After the final departure of the snow in spring, there is a change in net radiation, evaporation and ground heat flux. A major requirement in high latitudes is documentation of the magnitudes of the forcing parameters, of which the most important is precipitation, in all its forms. Synoptic weather systems play a major role in day-to-day energy and water responses to climate forcing. Presented in Table 2 is an example of the influence of weather conditions on the energy balance of Hudson Bay coastal wetland. RN is similar for both warm and cold overlying air masses. However, all other energy balance components are different. Under warm air mass conditions latent heat flux is greater than sensible heat flux, as expected in low and mid-latitudes, but in the cold air condition, sensible heat flux is greatly enhanced (Rouse, 2000). Harazono et al. (1996) found that energy partitioning at a coastal site near Prudhoe Bay, Alaska, was strongly controlled by cold and warm air advection as observed near the Hudson Bay coast (Lynch et al., 1999). Onshore winds advected cold and humid air masses from the Arctic Ocean resulting in low air temperature, a large temperature gradient between the land surface and the air and, therefore, a high sensible heat flux and low evaporative flux. Conversely, when offshore winds advected warm and dry continental air to the site, the temperature gradient between the land surface and air was small, resulting in low sensible heat flux, but only slightly higher evaporative flux than during onshore wind conditions (Lynch et al., 1999). The difference in energy partitioning was primarily due to larger heat gain of the open water ponds during the offshore conditions and to a minor increase in ground heat flux (Table 3). Yoshimoto et al. (1996) found a similar behaviour in energy partitioning at Barrow, Alaska, during their 1993 field season. Lynch et al. Table 2 Comparative energy balance under a warm air mass and a cold air mass at Hudson Bay Fluxes (W/m2) RN QE QH QB Source: Rouse (2000).
Warm air mass
Cold air mass
Warm/cold
161 85 47 19
161 64 80 17
1.00 1.33 0.59 1.13
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Table 3 Monthly mean surface energy balance (W/m2) at Betty Pingo coastal site (70180 N, 149550 W) for summer
June 1995 July 1995 August 1995
RN
QH
QE
QB
128.1 122.6 75.7
39.6 57.2 14.7
61.0 38.1 48.6
27.5 27.3 12.4
Source: Lynch et al. (1999).
(1999) used a regional climate model in their study (ARCSyM) to simulated fluxes that were within the range of measured fluxes (Table 3), but overestimated both net radiation and latent heat fluxes. The regional model also captured site to site variations quite well, which appear to be more sensitive to mesoscale meteorological conditions than to site characteristics. A study on the effect of advection from the cold polar sea to a warmer terrestrial surface in the coastal area of Hudson Bay was carried out by Weick and Rouse (1991). Three objectives were pursued: (i) to investigate the changes in the surface energy balance along a transect perpendicular to the coast line of the bay, (ii) to identify local and regional effects on the energy balance along the transect and (iii) to document the use of a box model and the divergence and convergence of energy mass in the coastal boundary layer during onshore and offshore winds. With Equation (1) the different fluxes and the Bowen ratio were estimated. The authors demonstrated that the largest advective influence on the turbulent fluxes occurs within 10 km of the coast, with a 2.7-fold downwind decrease in the Bowen ratio and a 1.8-fold decrease during offshore winds (Table 4). This decrease is due both to boundary layer adjustments to a new surface under onshore winds and to horizontal and vertical convergence and divergence in the atmosphere under all wind conditions. Different ET models applied to an Arctic coastal wetland near Prudhoe Bay, Alaska, during the summers between 1994 and 1996 were compared by Mendez et al. (1998). The objective was to gain a better understanding of ET in Arctic wetlands. Evaporation after spring snowmelt averaged 3.11 mm/day (obtained via the WB). Latent heat flux was the dominant heat sink in wetlands, whereas sensible
Table 4 Seasonal Bowen ratios () at four microclimate stations located at 0, 2.5, 9.4 and 12.4 km from the coast for all wind conditions and for onshore and offshore winds Wind conditions
Site 1 at the coast
All directions Onshore winds Offshore winds Source: Weick and Rouse (1991).
0.95 1.10 0.71
Site 2 (2.5 km)
Site 3 (9.4 km)
Site 4 (12.4 km)
0.79 0.93 0.58
0.58 0.64 0.48
0.58 0.67 0.43
Heat Energy Balance in Coastal Wetlands
225
heat flux dominated in the drier upland area. Differences between the formulations were not significant.
5. SUMMARY Heat budget studies in coastal wetlands although significant for the health of these ecosystems have been relatively poorly studied. The heat balance depends on the site conditions, the latitude and the climate. While the latitude determines the amount of incoming radiation entering the ecosystem, the resultant net radiation also depends on the site conditions that characterize the ecosystem. The net radiation represents the available energy that different processes, which characterize any energy budget, will have. Net radiation energy is used by evaporation (latent heat flux) and by heat conduction between the air–water and/or air–substrate (sensible heat flux). Naturally, the wetland region climate determines the magnitude of these fluxes (i.e., low latitudes wetlands receive more radiative energy). The typical vegetation of each ecosystem use part of that energy in specific biological processes. Therefore, the remaining heat is transported as advective heat flux either by wind or tides. Temperate wetlands, especially tidal flats, are by far the sectors that have received more attention. Plant cover influences the microhabitat of the sediment by controlling the amount of light reaching the sediment surface. Then significant differences might be found in heat balances between bare and vegetated marshes. These changes may induce changes in the sediment biotic community (Whitcraft and Levin, 2007). In low latitudes, besides the net radiation, the water balance and the evaporation are the most significant processes that define the heat balance, and then the temperature variations in the soils, coastal waters, and lower layers of the atmosphere. Temperature affects respiration rates and photosynthesis, influences the distribution and movements of fish, and also affects many important biological processes (number of eggs laid, incubation time, etc.). Therefore, heat balance studies might help to understand temperature variations of wetlands and some of the processes that generate distribution patterns in coastal natural flora and fauna communities. In the arid climate estuaries of low latitudes, the relationship between salinity and evaporation is important. Although some formulas that relate both parameters are presented in this chapter, more measurements and experiments in diverse environments should be made. On the other hand, in high latitudes, because of the cold climate, numerical models are the most powerful tool to study the heat balance. Despite the described research on heat budget in coastal wetlands, many points remain to be investigated such as the monitoring of the heat balance components in vegetated and bare coastal wetlands and its influence in plant community, the effects of shading in temperature variation of wetlands soils and the effect of the heat balance in plant zonation and animal interactions. The results of such studies would provide some insight in flora and fauna distribution, biodiversity and species
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behavior in wetlands. Scientists are working with the bulk formulas and/or numerical models in several sites, time and space scales. But even though some work has been done on the interaction between biological and physical processes, an exhaustive analysis of the basic interactions among flora, fauna and heat stored in coastal wetlands is still lacking. Future studies should investigate these interactions.
ACKNOWLEDGMENTS Partial support for the work dealing to this paper was provided by grants by CONICET, Agencia Nacional de Promocio´n Cientı´fica e Innovacio´n Tecnolo´gica, and Universidad Nacional del Sur. I would to thank the comments and suggestions by the Eric Wolanski, Mark M. Brinson, Bjo¨rn Kjerfve, W. Barclay Shoemaker, and Reiner Onken.
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C H A P T E R
8
H YDRODYNAMICS AND M ODELING OF W ATER F LOW IN M ANGROVE A REAS Yoshihiro Mazda and Eric Wolanski
Contents 1. Introduction 1.1. Peculiar hydrodynamics in mangrove area 1.2. Material dispersion 1.3. Holistic system 2. Physical Characteristics of Mangrove Topography and Vegetation 2.1. Classification of mangrove topography 2.2. Bottom condition of mangrove swamps 2.3. Influence of the vegetation on the hydrodynamics 3. Peculiar Hydrodynamics in Mangrove Areas 3.1. General equations that control water flow 3.2. Timescales of flow system 4. Modeling 4.1. Hydraulic model 4.2. Material dispersion model 4.3. Ecosystem model as the holistic system 5. Summary Acknowledgments References
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1. INTRODUCTION Mangrove forests cover wide tropical and subtropical intertidal areas, and they are very important for their role in maintaining biodiversity, for sustainable livelihood (e.g., wood and food resources) and for coastal protection (Robertson and Alongi, 1992; Wolanski et al., 2001, 2004; Mazda et al., 2002; Wolanski, 2006a). Human activities since the late 19th century have led to the reduction of mangrove forests around the globe (Spalding et al., 1997). This degradation seriously threatens the sustainability of mangrove ecosystems worldwide, and has also adversely affected human populations (Hong and San, 1993; Mazda et al., 2002; Hong, 2006). Coastal Wetlands: An Integrated Ecosystem Approach
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Managing mangroves requires understanding the natural mechanisms that form and maintain this environment. This requires using quantitative, process-based, models. In temperate coastal environments, such models have been developed based on a two-step procedure, namely, Step 1: A hydrodynamic model is used to calculate water flows, which transport and disperses chemical/biological materials. Step 2: Based on the hydrodynamic model, an ecosystem model is driven to calculate the flows of biomass and energy in the food web. This is a one-way process, whereby the physics drive the biology. This one-way methodology may not be adequate for mangrove areas, because in mangroves there are strong feedback processes as shown in Figure 1 (Mazda et al., 2007a), compared to the above temperate environment. The physiology of mangrove vegetation, the tidal motion in mangrove swamps, and the bathymetry of the mangrove-fringed tidal estuary have been formed through the feedback processes between themselves during many decades. Once the bathymetry is changed as a result of natural or human actions, this change leads directly to changes in the intensity and pattern of water flows. The change in the water flow leads to changes in the transport processes of water-born materials such as mangrove seeds and nutrients, and to changes in the distribution of mangrove trees/benthos. These changes in turn generate additional changes in the bathymetry. Accordingly, the mangrove ecosystem is a system for which feedback processes between the biology and the physics cannot be neglected. Modeling mangrove ecosystems following the above close connections between physics and biology requires linking separate models of biotic actions, water flows, geomorphology, and the atmosphere. The total ecosystem model is established through the connections or feedbacks between these individual models. Such a holistic ecosystem model is still being developed based on the following consideration.
Atmosphere
Biota
Water flow
Landform
Natural environment (ecosystem)
Figure 1 Feedbacks in mangrove environment (Mazda et al., 2007a).
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1.1. Peculiar hydrodynamics in mangrove area Mangrove areas constitute a very wide intertidal area, which is aquatic at high tide and terrestrial at low tide; thus, water movement cannot be treated as a continuous flow throughout the cycle of a tidal period. Densely vegetated mangrove trees, prop roots, leaves, and pneumatophores affect the horizontal and vertical hydrodynamics. Further, since the nature of water flow within mangrove areas depends on the timescale, it is necessary to develop different flow models due to tidal flow, sea waves, groundwater, and tsunamis, individually.
1.2. Material dispersion The fate of water-born material in the mangrove areas is controlled by the peculiar dispersion processes, which depend on the unique topography and spatial characteristics of mangrove vegetation. The fate of cohesive, fine suspended sediments influences the inflow of sediment-laden waters into the swamp and the settling of a fraction of this sediment in the swamp, thereby modifying the bathymetry and expanding the wetlands.
1.3. Holistic system The mangrove ecosystem is maintained via strong feedbacks between many factors as mentioned above. Each of these factors operates at different timescales. The total ecosystem is established by nonlinear interactions between these factors with contrasting timescales. Further, the mangrove environment should be understood as the ecohydrology, composed of the river basin, the river, the estuary, and coastal waters, through which not only water and dissolved materials but also biotic actions are strongly connected (Wolanski et al., 2004; Wolanski, 2006a). In this chapter, physical characteristics of mangrove topography and vegetation and then peculiar hydrodynamics in mangrove areas are introduced. Finally, based on these environmental characteristics in mangrove areas, methodologies for modeling mangrove environments are discussed particularly from the hydrodynamic viewpoint.
2. PHYSICAL C HARACTERISTICS OF MANGROVE T OPOGRAPHY AND VEGETATION Water flows in mangrove areas are strongly influenced by the mangrove trees and roots emerging from the substrate. Mangrove trees have built their own substrate (biogeomorphology) in intertidal areas and developed their colonies under the physical characteristics of the local tides and waves (Wolanski et al., 1992a; Mazda et al., 1999).
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2.1. Classification of mangrove topography From hydrodynamic viewpoints, mangrove forests are classified into three types based on topographic features, namely, fringe forest, riverine forest, and basin forest, as shown in Figure 2 (Cintron and Novelli, 1984).
(a)
H.W.
L.W.
(b)
H.W.
L.W.
(c)
H.W.
L.W.
Figure 2 Classification of mangrove topography (after Cintron and Novelli, 1984): (a) riverine forest type; (b) fringe forest type; and (c) basin forest type.
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2.1.1. Riverine forest (R-type) This type of landform is defined as floodplains alongside river drainage channels or tidal creeks, which are inundated by the highest tides and exposed during low tides (Figure 3). Tidal creeks commonly run perpendicular to the coastline or to the estuary banks, and are highly sinuous, intertwining with other creeks inside the forest. Wind-driven waves and swell (period less than typically 20 s) rarely propagate into the swamps because of the dissipation of wave energy along the long tidal creeks. Swamp water within a few meters from the tidal creek is dragged by tidal flow in the creek, thus it flows parallel to the creek. Further inside the swamp the flow is predominantly perpendicular to the creek due to the high vegetationinduced friction and the water surface gradient between the swamp and the tidal creek (Kobashi and Mazda, 2005; Mazda et al., 2005). 2.1.2. Fringe forest (F-type) This landform comprises swamps along shorelines that face the open sea and are directly exposed to the action of both tidal water and sea waves. Sea waves are mitigated in swamps because of the resistance of thick mangrove trees and emergent roots (Mazda et al., 1997a; Massel et al., 1999). 2.1.3. Basin forest (B-type) This landform comprises partially impounded depressions that are seldom inundated by high tides during the dry season, but are inundated by spring high tides
10°30′ N
Mul Na l River
N
0
Long
Can Gio Hoa South China Sea 10 km
107°00′ E
Figure 3 Map of Can Gio district, Vietnam, highlighting the presence of numerous tidal creeks, as well as small and local, tidal rivers, and their tributaries in a mangrove forest.
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during the wet season (John and Lawson, 1990). During the dry season, the water level in the depressions continues to fall slowly because of groundwater flow to the open sea driven by the difference in water level between the depression and the open sea. The groundwater that slowly leaves and enters the swamp with tidal action exchanges various water-born materials between the swamp and the open sea and vertically mixes the bottom water in the depression (Mazda et al., 1990a,b). The water flows are different between these three systems. The flushing characteristics vary between these types. Indeed, it is generally observed for R-type forests that after the tidal water above the substrate has ebbed away from the swamp to the creek, the surface soil at sites near the creek dries rapidly at low tide. In F-type the surface soil remains wet at low tide (Mazda and Ikeda, 2006).
2.2. Bottom condition of mangrove swamps The bottom slopes of mangrove swamps are very small (typically 1/1,000) although there are innumerable local depressions or small holes formed by bioturbation and water eddies that develop behind mangrove roots and trunks. The bathymetry is extremely complex because mangrove trees and their roots occupy a large proportion of the bottom area of swamps (Figure 5). Below the substrate there are many macropores due to animal burrows and decaying roots. For instance, approximately 10% of the total volume of the bottom sediment to a depth of approximately 1 m in a Rhizophora forest at Gordon Creek, Townsville, Australia, is composed of animal burrows that are intermingled but not connected (Stieglitz et al., 2000). Soils within mangrove swamps are organic rich. For instance, the soil within the top 1 m within a mangrove swamp at Cocoa Creek, Cleveland Bay, Australia, is mainly composed of organic clay made up of particles smaller than 10 mm (Susilo, 2004). Fine sediment is trapped and accumulates in mangrove swamps compared to the open coast (Sato, 2003). For instance, the particle size of sediment within the mangrove swamp that fringes the MairaGawa River, Iriomote Island, Japan, is an order of magnitude smaller than that at the adjacent open coast (Mazda and Ikeda, 2006).
2.3. Influence of the vegetation on the hydrodynamics Mangrove roots are very dense in swamps. They emerge vertically from the soil (Sonneratia, Avicennia, Laguncularia, Bruguiera, and Lumnitzera) or take the form of prop roots (Rhizophora). In addition, thickly vegetated mangrove canopies cover swamps. These vertical configurations perform a variety of roles, for example, the canopies moderate atmospheric influences and the roots resist water flow although mangroves are influenced physiologically by atmospheric and hydrological conditions such as sunlight, humidity, tidal flow, sea waves, and salinity (Snedaker, 1989). The vegetation density is greater in the vicinity of the bottom substrate of the swamp, though each mangrove species has a unique configuration. As shown in Figure 4a, the control volume in a mangrove swamp V (a hatched rectangular element) with a horizontal area S (Dx Dy) and a depth H is divided into two parts,
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(a)
H Δx
Δy
(b)
Mangrove swamp
F T,x2
Δy y
F T,y1
v
F D,x
u
F D,y
F T,y2
Δx
F T,x1 x
Figure 4 (a) Sketch of the control volume in a mangrove swamp and (b) a schematic plan view of the hydrodynamics.
V = V M þ VW
ð1Þ
where VM is the total volume of obstacles, which is composed of submerged tree trunks and roots over the substrate, and VW is the volume of water in V. VM is not very small compared to VW and cannot be disregarded, particularly at small depth (Figure 5a; Mazda et al., 1997b). Mazda et al. (2005) modified a proposal of Wolanski et al. (1980) and defined the characteristics of the vertical configuration of mangroves by the representative length scale in mangroves. L=
VW A
ð2Þ
where A is the total projected area of the obstacles (i.e., trees and roots) to the flow in the control volume V. L has a dimension of length and includes information on the spacing between vegetations such as mangrove trunks and roots in the swamp. Mazda et al. (1997b) suggested that L varies significantly with vegetation species and tidal elevation. Several examples of L that varies with tidal level are shown in Figure 5b.
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(a) 50 Water depth (cm)
a 40 30 b
c
20 10 0
0
0.1
0.2
0.3 VM / V
0.4
0.5
(b) 50 a
b
Water depth (cm)
40
c
d
f e
g
30 20 10 0
0
10
20
50 60 70 30 40 Representative length in mangroves (cm)
80
90
Figure 5 (a) Dependence of VM /V and (b) the representative length L on the water depth for different mangrove species. In Figure 5a, a = Coral Creek, Hinchinbrook Island, Australia (Rhizophora sp.), b = Nakama-Gawa, Iriomote Island, Japan (Bruguiera sp.), c = Nakama-Gawa, Iriomote Island, Japan (Rhizophora sp.), (Mazda et al., 1997b). In Figure 5b, a = Coral Creek, Hinchinbrook Island, Australia (Rhizophora sp.), b = Can Gio, Ho-Chi-Minh, Vietnam (Rhizophora sp.), c = Maera-Gawa, Iriomote Island, Japan (Rhizophora sp.), d = Chone River, Manabi, Ecuador (Rhizophora sp.), e = Aira-Gawa, Iriomote Island, Japan (Bruguiera sp.), f = Nakama-Gawa, Iriomote Island, Japan (Bruguiera sp.), g = Maera-Gawa, Iriomote Island, Japan (Bruguiera sp.), (Mazda et al., 2005).
The tidal hydrodynamics are strongly dependent on L (Mazda et al., 1997b, 2005). Because L is defined at the macroscopic scale, the application of L for hydrodynamics is not valid for the analysis of small-scale water motions such as that due to sea waves. Further, Wu et al. (2001) proposed the blockage effect, i.e. porosity effect of mangrove vegetation on the flow structure in the swamp system, which plays an important role when the vegetation density is high. From the viewpoint of land protection, Sato (1978) statistically analyzed the morphological characteristics of Rhizophora mucronata and discussed the relationship
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between its spatial distribution and sedimentation. He inferred that the root system, which concentrates near the substrate, contributes to determining the particle size distribution of the bottom sediment in the marginal and central parts of swamps. Based on field observations within mangrove forests of Rhizophora spp., Furukawa and Wolanski (1996) and Furukawa et al. (1997) emphasized that sedimentation associated with short-period waves such as sea waves depends upon the detailed flow pattern around the root matrix (Figure 6). The belowground roots are generally confined to a thin layer (1 m) below the substrate (Komiyama et al., 1989). Notwithstanding, the radial extent of these roots and their density are comparable to that of the aboveground roots and trunks (Komiyama et al., 2000). For swell waves, from field observations of forests dominated by Kandelia sp. and Sonneratia sp., Mazda et al. (1997a) and Mazda et al. (2006), respectively, quantified the role of the vegetation in reducing the wave energy as the wave propagates in mangroves. Mangrove branches and leaves in the canopy are generally located above the height of the water surface, even at spring high tide. However, when the water level at spring high tide is further raised by an increase in mean sea level during the rainy season or a typhoon accompanied by large sea waves (Section 3.2), the leaves are submerged and contribute significantly to reducing sea waves by applying a drag force to the water flow. Mangrove vegetation, further, buffers tsunamis, particularly when the trees fall into the water (Mazda et al., 2007b). The great volume of the vegetation, including (a) Trunk 10 cm/s 20 cm
(b)
Rock
Root
Figure 6 Observed 1-min mean water velocities around mangroves for (a) Ceriops sp. and (b) Rhizophora sp. (Furukawa and Wolanski, 1996).
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not only tree trunks but also thick vegetated branches and leaves of fallen trees and intricate roots uprooted and exposed in the water, acts to dissipate the kinetic energy of the tsunami. The wind decreases through the height of the tree canopy. This diminishes the momentum of the air exiting the canopy and thus shelters the area downstream by a wake effect. As a result, the airflow downwind of trees creates less turbulent zones, where the suspended salt particles deposit, resulted in the significant protection against salt spray (Wolanski, 2006b). Additionally, the mangrove canopy decreases the wind inside the forest and solar radiation.
3. PECULIAR HYDRODYNAMICS IN M ANGROVE A REAS 3.1. General equations that control water flow The behaviors of hydraulic factors, such as tides, sea waves, and groundwater flows in mangrove areas, are very different from those of temperate estuaries that have been studied in detail, though the physical mechanisms are basically the same between both areas. One of the reasons for this is that mangrove estuaries generally constitute a very wide intertidal area compared to the temperate estuaries, resulting in the condition that the water movement cannot be treated as continuous flow throughout the tidal cycle. Furthermore, the mangrove estuaries differ from the temperate estuaries by the presence of numerous obstacles to water flow, including mangrove trees, prop roots, and pneumatophores. The flow around these obstacles forms wakes as well as turbulence with various spatial (horizontal and vertical) and temporal scales, as mentioned in Section 3.2.7. As demonstrated for R-type forests that are composed of mangrove swamps and tidal creeks, water motions in the swamp and the tidal creek strongly interact with each other (Wolanski et al., 1992a), notwithstanding the contrasting topographies of the wide-flat swamps and the narrow-deep creeks. The water flowing through a mangrove swamp is resisted by the drag force due to mangrove trees and their roots, by the bottom friction on the uneven mud floor, and by the eddy viscosity due to turbulent motions of water through gaps between trees and roots (Figure 6; Furukawa and Wolanski, 1996). Following the concept of Raupach and Thom (1981) and Shimizu et al. (1992) who modeled the drag force and the eddy viscosity separately, the momentum and mass conservation equations for the aboveground flow are, respectively @u @u @u @ þu þv VW = g VW þ FD;x þ ðFT;x1 þ FT;x2 Þ þ FB;x ð3Þ @t @x @y @x @v @v @v @ ð4Þ þu þv VW = g VW þ FD;y þ ðFT;y1 þ FT;y2 Þ þ FB;y @t @x @y @y @ @u @v = þ H @t @x @y
ð5Þ
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where x and y are the horizontal axes (Figure 4b); t the time; u and v the depthaveraged current velocities in the x- and y-directions, respectively; the water surface elevation; g the acceleration of gravity; FD,x and FD,y the x- and y-components of the drag force due to submerged mangrove trees/roots; FT,x1 and FT,x2 the shear stresses (eddy viscosities) on two vertical planes along x-axis; FT,y1 and FT,y2 the shear stresses on two vertical planes along y-axis; and FB,x and FB,y are the x- and y-components of the bottom friction, respectively. Wu et al. (2001) proposed that these equations should be modified using the porosity factor (see Section 2.3) that is composed of the diameter and density of mangrove trees when the vegetation density is high. The shear stress on the water surface can be ignored because the trees decrease the wind through the canopy. FD, FT, and FB depend on the vegetation density and the bottom roughness. Further, FD and FT may vary with the tidal stage because the submerged portion of mangrove vegetation varies with the tidal stage. Each of these terms has to be formulated for the local vegetation condition and the water flow type such as tidal flow or sea waves.
3.2. Timescales of flow system In mangrove areas, there are various horizontal and vertical water motions with different timescales, from changes in water level with a seasonal period, tidal periods, sea waves at periods less than typically 20 s, to turbulence at timescales of less than 1 s. When modeling the system, the dominant mechanisms controlling or at least influencing the system should be identified. Water motions with different timescales are described in the following sections. 3.2.1. Seasonal change Mangrove ecosystems are influenced by seasonal changes in climate, even though they are within tropical to subtropical areas. Although seasonal variations in rainfall are well known, seasonal changes in sea level have often been neglected. Seasonal changes in sea level result from a number of factors: (1) variations in wind direction and speed (often monsoonal) upon the coastal ocean (Ridd et al., 1988), (2) changes in water temperature that bring about an expansion in water volume, (3) changes in atmospheric pressure, and (4) changes in river runoff due to rainfall (Kjerfve, 1990). For example, at the Can Gio coast, southern Vietnam, the annual range in mean sea level between the summertime low and the wintertime high is approximately 0.6 m (Figure 7c). At Ishigaki Harbor, southern Japan, the range between the springtime low and autumn high is approximately 0.4 m (Figure 7a), whereas the annual range at the mouth of the Chone River, central Ecuador, is not apparent (Figure 7b). Further, there is a semiannual change in the tidal range (Figure 7c). Thus, some mangrove swamps are dry for several months, and this influences not only the growth of mangrove trees but also the survival of benthos (Mazda and Kamiyama, 2007). 3.2.2. Fortnightly change in tidal regime Variations in tidal forcing lead to the spring–neap cycle, which is associated with the lunar cycle and often causes the tidal range to change significantly over
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(a) 200
100
0
200
(b)
100
Tidal elevation (cm)
0
–100
(c) 400
300
200
100
0 1 Jan 20 Feb
11 Apr
31 May
20 Jul Date
8 Sep
28 Oct
17 Dec 2001
Figure 7 Time series plots of tidal elevation (a) at the Ishigaki Harbor, Okinawa in Japan, (b) at the mouth of Chone River in middle Ecuador, and (c) at the Long Hoa coast, Can Gio in Vietnam, during 1 year. Grey lines show the mean sea levels. Adapted from Mazda and Kamiyama (2007).
fortnightly timescales (Figure 7). The volume of water that enters a mangrove swamp changes markedly during the spring–neap cycle (Mazda et al., 1995). The magnitude of groundwater flux also varies markedly between the spring and neap tides (Mazda and Ikeda, 2006). Tidal inundation in the innermost parts of
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mangrove swamps is particularly affected by this cycle; the soil in these innermost parts may last to be exposed for several days during neap tide, and their ecosystems depend strongly on this environmental change (Snedaker, 1989; Kjerfve, 1990), as well as the seasonal change. 3.2.3. Daily change in atmospheric variables The diurnal variation in solar radiation causes changes in air temperature and soil temperature, particularly in coastal tidal flats adjoining a mangrove swamp. These variations result in daily changes in the temperature of the water that inundates mangrove swamps. Dissolved oxygen in coastal water, which is transported to mangrove swamps by tidal action, also varies with the daily cycle in solar radiation and biotic activity such as photosynthesis and respiration (Mazda et al., 1990b). 3.2.4. Tidal change with a diurnal or semidiurnal period The tides induce changes in water depth and in horizontal currents in the mangroves. Tidal fluctuations are due to the summation of sinusoidal tidal harmonic components. The diurnal constituents are K1, O1, P1, Q1, and S1, which have periods of 23.93, 25.82, 24.04, 26.87, and 24.00 h, respectively. The semidiurnal constituents are M2 (12.42 h), S2 (12.00 h), N2 (12.66 h), and K2 (11.97 h). Among these components, K1, O1, M2, and S2 are termed the dominant tidal constituents. The tidal range changes fortnightly with the progression of spring and neap tides. The amplitude and phase of each tidal component depend on the location of the observation site. Generally, the levels of high and low tide change daily, as does the interval between consecutive high tides; these fluctuations are known as the tidal inequality. Differences in the periods of the dominant tidal components mean that the timing of high tide shifts by approximately 50 min each day (Ippen, 1966). Thus, the phase of tides shifts from that of solar radiation by approximately 50 min every day. In mangrove areas, the tide commonly behaves like a solitary wave. At low tide the bottom substrate is exposed. The water begins to inundate the swamp at around the middle of the flood tide period (at about mean sea level offshore). The water flow through the vegetation stops at about high tidal level and reverses direction as the tide ebbs. Even if the tidal oscillations offshore are symmetrical, the tide becomes highly asymmetrical in mangrove swamps due to the effect of mangrove vegetation and local topography (Mazda et al., 1995; Aucan and Ridd, 2000). The modification of the tidal signal is apparent at the time when the water begins to inundate the swamp, and is especially pronounced at around the time when the bottom substrate dries up at ebb tide (Mazda and Kamiyama, 2007). 3.2.5. Resonant oscillation In many mangrove swamps, water oscillations with a period of 10–30 min are present in addition to the usual astronomical tidal oscillations (Mazda et al., 2007a). These are free oscillations with a period that is dependent on the horizontal dimensions and the water depth of the area. This water oscillation is called to a seiche or a resonant oscillation (Ippen, 1966). The wavelength of this oscillation is
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of the same order of magnitude as the horizontal scale of the area. This water motion can transport floating leaves and suspended and bottom sediments between the swamp and the coastal area. 3.2.6. Sea waves Sea waves that penetrate mangrove areas vary from short-period wind waves (<1 s) to longer-period swells with periods of up to 20 s that are induced by tropical depressions in the open sea. In R-type forests, the waves are attenuated, as mentioned in Section 2.1.1. Consequently, the water surface in this swamp remains calm. In F-type forests, however, sea waves intrude into the swamp directly, then being attenuated due to the drag force of mangroves within the region of the swamp close to the boundary with the open coast. The rate of wave reduction depends on the vegetation conditions and the characteristics of the waves in the open sea (Mazda et al., 1997a, 2007b; Massel et al., 1999). 3.2.7. Water turbulence Mangrove prop roots and pneumatophores are densely intertwined above the bottom substrate. Because of interaction between mangrove roots and tidal flows or sea waves, water turbulence or eddies with periods smaller than those of sea waves occurs primarily in the region of the swamp near the boundary of the open coast. These turbulent interactions act to mix and diffuse water and materials, and contribute to form and maintain the distribution of material within the swamp (Wolanski et al., 1992b, 1996; Wolanski, 1995; Furukawa and Wolanski, 1996; Furukawa et al., 1997). In particular, Wolanski et al. (1998) proposed that the sedimentation is enhanced by the turbulence around the vegetation and results in the formation of new land. 3.2.8. Damaging events Tropical depressions or typhoons can result in an increase in water levels by several meters, and this effect can last several days in coastal areas (Dean and Dalrymple, 2002). Superposition of this elevated water level on high water at spring tide leads to destructive conditions in coastal areas, because sea waves move across such deep water without significant reduction in energy (Mazda et al., 1997a, 2007b). When tsunamis (seismic sea waves) occur, the bores penetrate into mangrove swamps with a period of between 10 min and 1 h. In tidal creeks, wakes generated by boats and ships produce waves with a period of several seconds. These waves erode sediment from the creek bottom and the creek bank made of the fine and often unconsolidated nature of mangrove sediments (Mazda et al., 2006). The sea level may rise up to 0.5 m by the year 2100 due to global warming. This will facilitate the intrusion of sea waves in the mangroves and accelerate erosion at the seaward fringe of the mangrove. If the landward side of the mangrove swamp is not dyked, the mangrove area may migrate landward. Otherwise it may disappear. The response of coastal ecosystems to a sea level rise remains unknown (Kikuchi, 1995; Miyagi, 1998).
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3.2.9. Interaction between processes with different timescales In addition to the mangrove vegetation, the benthos and the mangrove topography also result from various feedback systems with different timescales. The succession of the physiology and ecology of biota occurs at timescales of several decades or more. In contrast, the tidal regime has diurnal or semidiurnal periods that repeat with a period of 1 year in a statistical sense (Section 3.2.4). Mazda and Kamiyama (2007) described the nonlinear interaction or feedback between these biota and tidal regimes with different timescales; biota in mangrove areas are unable to respond instantly to a tidal regime with diurnal or semidiurnal periods, but follow the long-term tidal inundation characteristics over many decades, often in a logarithmic manner (Mazda, 1984).
4. M ODELING 4.1. Hydraulic model As mentioned in Section 2.1, the flow mechanisms in mangrove areas are very different between R-, F-, and B-types. Further, the fate of water flow in their systems is strongly restricted by various timescales, as mentioned in Section 3.2. Thus, when modeling the mangrove system, the dominant mechanism controlling the system should be identified spatially and in timescale. 4.1.1. Tidal circulation model The hydrodynamics in mangrove areas were initially studied by Wolanski et al. (1980) who proposed a numerical model for an R-type mangrove area in Coral Creek, Hinchinbrook Island, Australia. As defined in Section 2.1.1, R-type mangrove forests are composed of tidal creeks and mangrove-vegetated swamps, which are topographically very different with each other. Therefore, the model combines a 1-D model for the tidal creek and a 2-D model for the mangrove-vegetated flood plain (Figure 8). To explain this model simply by the hydrodynamics shown in Section 3.1, we assume here that the creek in this model is along x-direction. Since there is no mangrove vegetation in the creek, both the drag force and the eddy viscosities in Equation (3) can be neglected. The bottom friction (FB,x) was parameterized using the Manning roughness coefficient. For the swamp, inertia terms were neglected because generally water currents in mangrove swamps are small; however, the drag forces due to vegetation [FD,x and FD,y in Equations (3) and (4)] were parameterized using the fraction of cross-sectional area of flow that is blocked by vegetation per unit length of flow, where the fraction decreases rapidly with elevation. The eddy viscosities in Equation (3), that is FT,x1 þ FT,x2, were dealt with by the following simplified parameterization. Water leaves the creek to enter the mangrove swamp with its momentum that is dissipated by friction around vegetation in the first few meters from the banks of the creek. Thus, at rising tides, the momentum was parameterized using the velocity in the creek and the lateral exchange rate between the creek and the mangrove swamp. At falling tides, water
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(a) 26 cells for the creek
(b) 96 cells for the swamp
Figure 8 (a) Location of cross-sections in Coral Creek in Hinchinbrook Island, Australia, and (b) division of mangrove swamps into cells used in the mathematical model (Wolanski et al., 1980).
flowing from the swamp into the creek was assumed to bring a negligible contribution to longitudinal momentum flux in the creek. The eddy viscosities in Equation (4), that is, FT,y1 and FT,y2, were assumed to balance each other, because the tidal wave length along the creek is sufficiently long compared to the grid scale for calculation. The 1-D model and the 2-D model were linked through the lateral exchange rate and the water level at the boundary between the swamp and the creek. On the other hand, Mazda et al. (1995) discussed the hydrodynamics of a tidal creek-mangrove swamp system using a two-dimensional numerical mesh model. In this model, the eddy viscosity (FT) was assumed to be negligible compared to the drag force (FD). The drag force was parameterized using a constant drag coefficient. In this model analysis, they explained the tidal flow mechanism that generates a large tidal inequality characterized by the peak ebb tidal currents being nearly twice as large as the flood tidal currents. Further, based on this model, Mazda et al. (2002) pointed out that the flooded volume of a mangrove swamp results in a huge increase in water flux within the creek, particularly at the creek mouth, compared to the case of a similar creek without swamps. Both models mentioned above highlighted the importance of the interaction between tidal creeks and fringing mangrove swamps in shaping the tidal flow distribution not only in mangrove swamps but also in the tidal creeks. This strong interaction between tidal creeks and fringing mangrove swamps controls material dispersion and siltation in this area, and indirectly the coastal erosion (Mazda et al., 2002). Conclusively, this hydraulic interaction relates to form the topography and to maintain the mangrove ecosystem.
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The spatial scales of creeks and fringing mangrove swamps are quite different with each other, as mentioned above. To accurately simulate the interaction between these creeks and mangrove swamps, Nihei et al. (2001, 2004) proposed to apply a nesting procedure, in which the computational results in the larger-scale domain are given as the boundary conditions for the smaller-scale computation. Applying this model to an R-type mangrove area, Fukido-River in Ishigaki Island, Japan, they adopted a one-way nesting procedure. In this computation, to realize the highly resolution, three computational domains, the larger-, intermediate-, and smaller-scale domains with different grid resolutions were used. As a result, they could reproduce not only the flow distribution in the swamp and the creek but also the horizontal eddies around the creek bank. The model thus suggested the presence of mechanisms forming and stabilizing the R-type topography. Mazda et al. (1995) initially assumed that the eddy viscosity is negligible, as mentioned above. However, based on field data, Kobashi and Mazda (2005) demonstrated its importance in mangrove swamps, particularly near the creek bank. Mazda et al. (2005) used their observations at a number of mangrove sites to show that both the drag force (FD) and the eddy viscosity (FT) play dominant roles in the tidal-scale hydrodynamics in mangrove swamps. They also parameterized the drag coefficient and the coefficient of dynamic eddy viscosity as functions of the Reynolds number defined by the representative length of the vegetation [L; Equation (2)]. The representative length varies greatly with vegetation species, vegetation density and tidal elevation. Both these coefficients decrease with increasing values of the Reynolds number (Figure 9). At the low range of the Reynolds number both these coefficients reach much higher values than those typical of vegetation-poor estuaries and rivers. Consequently, the tidal flow within mangrove areas depends to a large degree upon the submerged vegetation density, and this varies with the tidal stage. As mentioned previously, Wu et al. (2001) introduced an idea of the porosity depending on the tree diameter and the vegetation density in mangrove swamps, Based on this idea, Wu et al. (2001) and Struve et al. (2003) demonstrated the blockage effect on the hydrodynamics in mangrove areas through experiments in a hydraulic flume and depth-integrated 2-D numerical modeling. However, their porosity factor is defined to be independent of tidal elevation. It is expected in future that the porosity factor in their mathematical model will be modified to depend on tidal elevation. Except for the interaction between mangrove swamps and tidal creeks, the tidal flow mechanism in F-type forests is the same as that described above. 4.1.2. Sea waves model F-type mangroves are directly exposed to the action not only of tides but also of sea waves, though in R-type mangrove areas sea waves are not important. Hong and Dao (2003) confirmed that, when a typhoon hit an area in the Thai Thuy district, northern Vietnam in 1996, shrimp and crab pond-embankments in this area were well protected by mangroves growing offshore, while those in a nearby coast without the protection of mangroves were eroded and destroyed by sea waves of the typhoon. The mangrove forest not only protects the coast from coastal erosion but also provides habitats and nursery grounds for aquatic lives. Thus, along many
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Drag coefficient CD
(a) 100
10
1
×104 10
0.1 0
4
2
6
8
Reynolds number Re (vL/ν )
Coefficient of dynamic eddy viscosity f (cm2/s)
(b)
×103 1000
100
10
1
0.1
×104 0
1
2
3
4
5
6
7
Reynolds number Re (vL/ ν )
Figure 9 Relationships (a) between the drag coefficient CD and the Reynolds number Re (Mazda et al., 1997b), and (b) between the coefficient of dynamic eddy viscosity f and Re (Mazda et al., 2005).The marks in the figures show the different observation sites.
coastal areas earnest efforts for reforestation have been executed on an extensive scale. However, excessive planting, that is, excessive vegetation density and/or excessive width of the vegetation area, may inhibit the water movement in the recesses of the forest, and prevent the exchange of water and organic/inorganic materials between the forest and the open sea. This increases the residence time for suspended materials and may increase pollution. In turn, this results in degradation of water/sediment qualities and consequently of the ecosystem in the forest. Wolanski (2006c) reported pollution problems in mangrove-fringed harbors in the Asia Pacific region. This reforestation paradox should be noticed. To avoid this paradox, appropriate assessments should be executed through quantitative models before attempting remedial measures.
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Drag coefficient: CD
0.4
0.3
Area C
0.2 Area B
0.1 Area A 0
0
20
40 60 Water depth: h (cm)
80
100
Figure 10 Variation of the wave-induced drag coefficient CD with the water depth h in the mangrove coast along Thuy Hai, north Vietnam (Mazda et al., 1997a). Areas A, B, and C are composed of a 0.5 (seedlings), 2^3, and 5^6 -year-old trees, respectively.
Figure 10 shows a strong relationship between the water depth, the growth level of mangrove trees and the drag coefficient due to the submerged mangrove vegetation in a Kandelia candel forest planted with approximately 1 m intervals (Mazda et al., 1997a). The vegetation-induced wave reduction depends not only on the water depth but also strongly on the vegetation condition such as the mangrove species, growth level, and vegetation density. Further, Mazda et al. (2006) showed from their field data in an F-type mangrove area that all the vegetation contributes to this wave reduction, including the mangrove tree trunks, the leaves, and the roots emerging from the soil. In the above discussions, the significant wave, which was defined as the average of the highest one-third of the waves, was used for the practical application of coastal protection. On the other hand, Massel et al. (1999) analyzed wave energy dissipation in mangrove forests in the frequency domain from theoretical standpoint. Their analytical model pointed out that the wave energy attenuation depends not only on the vegetation conditions but also on the spectral characteristics of the incident waves. The fate of the resonant oscillation in mangrove areas has hardly been discussed (Mazda et al., 2007a), though this oscillation may be important in material exchange between the swamp and the coastal area, as mentioned in Section 3.2.5. In consideration of spatial- and time-scales, the hydrodynamic model for the resonant oscillation in mangrove areas may be analogous to the tidal model mentioned above rather than the sea waves model.
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4.1.3. Tsunami model Mangroves buffer tsunamis (Mazda et al., 2007a,b). The period of tsunamis, typically 10 min to 1 h, is very different from both of sea waves (<20 s) and tides (12 h). In comparison between Figures 9a and 10, it is noted that the formulation of the drag force for sea waves is very different from that for tidal waves, because of the difference of timescale between them. Considering the above facts, the hydrodynamic effects of a tsunami wave in a mangrove forests should not be estimated by interpolation between sea waves and tides. Tsunamis destroy mangrove forests by their huge inertial force at the bore-like wave front (Imai and Matsutomi, 2005), while the inertial force is neglected both for tidal and sea waves (Mazda et al., 2007b). Harada and Imamura (2005) summarized the effects of coastal forest width, vegetation density, and wave period on the reduction of tsunamis, and proposed the criteria to identify quantitatively the relation of the tsunami intensity to the disaster, which can be used as a quantitative standard to design a coastal forest as a tsunami countermeasure. These studies, however, have not been specifically designed for mangrove areas. Dam break models suggest that the tsunami bore becomes a smoothly rising flood wave when penetrating 500 m inside a continuous wall of mangrove vegetation. However, the wave can be amplified at the mouth of a tidal creek, breaking the continuous wall of vegetation, and the tsunami wave can then propagate landward along the tidal creek and impact areas upstream (Wolanski, 2006b). Considerably more research is needed to quantify these processes (Mazda et al., 2006). Some mangrove areas are adjacent to 5–10 km wide tidal flats. In these conditions, tsunamis may behave with peculiar characteristics. The tsunami wave breaks first offshore and propagates shoreward along the tidal flat as a bore. When arriving in the mangrove swamp, the bore scours off the loose muddy bottom around mangrove roots with approximately 1 m depth under the bottom substrate (Section 2.3) so that the mangrove trees fall down. The volume of mangrove trees above the substrate, which fall down into the water, and the volume of uprooted mangrove roots, which is comparable as the total volume above the substrate (Komiyama et al., 2000), act together to absorb some of the kinetic energy of the tsunami. If the mangrove trees remain standing, both of their underground roots and leaves in the air would have no effect as obstacles to the flow. Thus, though the mangrove natural environments in this area are destroyed by tsunamis, they protect human lives behind them by sacrificing themselves. However, we have little quantitative information (Mazda et al., 2007a,b). To design a coastal mangrove forest as a protection against a tsunami the following topics should be studied: 1. The mechanism of hydraulic resistance in mangrove forests in relation to the timescale of tsunami waves, under situations not only of standing trees but also of fallen trees. 2. The mechanisms of tsunami waves scouring the bottom and scooping up underground roots in mangrove forests, as well as snapping mangrove trunks. 3. The mechanism of deformation/attenuation of tsunami waves at reef edges and over a wide shallow tidal flat. 4. The hydraulic criteria, which can be used as a quantitative standard for designing and establishing tsunami control forests, based on the results in the above points 1–3.
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4.1.4. Groundwater model In mangrove swamps, groundwater seepage is often ignored by modelers because of its small magnitude compared to aboveground tidal flows. However, the groundwater plays an important role in the mangrove ecosystem. Ridd and Sam (1996) stated that the salinity of groundwater in mangrove swamps controls the growth and distribution of the trees. Provided the one-dimensional, homogeneous bottom soil and negligible inertial term, the macroscopic momentum and the continuity equations for groundwater are 1 @u @ u = g g @t @x k
@ @u = Hb @t @x
ð6Þ
ð7Þ
where u is the macroscopic groundwater velocity at a site x and time t, the groundwater level, k the hydraulic conductivity, the effective porosity, and Hb the height of permeable layer. Based on Equations (6) and (7), Mazda et al. (1990a) simulated the groundwater flow in mangrove swamps both of F-type and B-type. They found out, particularly for the B-type, that the underground flow has three components: (1) a quasi-steady flow toward the open sea due to the tidal mean pressure gradient between the water levels in the depressed swamp and the open sea; (2) a tidally reversing flow with exponentially damped amplitude and linearly delayed phase toward the swamp; and (3) a residual flow toward the swamp caused by the exponentially damped tidal flow. In Bashita-Minato, Iriomote Island, Japan, these groundwater flows played an important role in determining the water properties and the sediment condition. On the other hand, Susilo (2004) developed an analytical model in which the Laplace equation defined by the hydraulic potential was solved under the conditions that the soil is homogeneous and isotropic, and simulated accurately the flow pattern of groundwater in Gordon Creek and Cocoa Creek, Australia. The soil is, however, not isotropic in practice, there are abundant burrows produced by crabs and other organisms (Ridd, 1996). These burrows are hydraulic conduits that greatly facilitate groundwater flows (Heron and Ridd, 2001, 2003). Susilo and Ridd (2005) proposed a simple mathematical model of animal burrows to infer the resulting the hydraulic conductivity [k in Equation (6)] of mangrove sediment. Mazda and Ikeda (2006) showed that the hydraulic conductivity in mangrove swamps is two to three orders of magnitude larger than that measured in a laboratory using small scale sediment core samples collected in the swamp, because large animal burrows as well as humus-rich sediment layers increase the permeability of the mangrove soils (Section 2.2).
4.2. Material dispersion model Dissolved, floating, or suspended particles in/on water such as mangrove seeds, fish eggs, prawn larvae, nutrients, and fine sediments are dispersed within mangrove
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Ck
Conn
Conn
Ck
Hinchinbrook Channel
areas, and may then be flushed out to the adjacent coastal sea due to tidal and wave actions. This dispersion is important for sustaining mangrove colonies and their ecosystems (Boto and Bunt, 1981). Applying the concept of material trapping in an embayment, which was proposed by Okubo (1973), Wolanski and Ridd (1986) and Ridd et al. (1990) showed that the material dispersion in R-type forests and estuaries is controlled primarily by the tidal trapping in the swamps. The tidal trapping is the process of temporary water storage in the swamp at rising tide when swift tidal currents prevail in the creek. On returning to the creek at ebb tide, the trapped water mixes with the creek water. Through this process, the water-born material disperses longitudinally along the creek. Wolanski et al. (1990) computed the fate of contaminants around a coastal zone adjacent to R-type mangrove forests (Figure 11). As a result, the residence time and the fate of water-born material were found to be controlled by the tidal trapping effect in mangrove swamps. Coastal waters move back and forth between the coastal boundary layer and the mangrove swamp; thus mixing between estuarine and offshore waters was strongly inhibited. This ensured a strong dynamic link between mangroves and coastal waters, and the importance of the tidal trapping effect within mangrove swamps. Wolanski et al. (1998) demonstrated that the above process also traps fine sediment in the intertidal wetlands and builds new mud banks. As mentioned in Section 4.1.1, the mangrove swamp in R-type forests generates a tidal asymmetry with larger ebb than flood tidal currents in the tidal creek. The tidal creek is thus
5 km
Figure 11 Predicted, synoptic distribution of the cloud of particles in Hinchinbrook Channel, North Queensland, Australia, at high tide (left) and low tide (right) 10 days after 48,000 particles were released near the coast in the channel near Conn Creek. The cloud of particles after 10 days is still highly concentrated near the original point of release (Wolanski et al., 1990).
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self-scouring. This result appears analogous to that in areas with wide intertidal areas demonstrated by Uchiyama (2004) who incorporated a wetting and drying scheme into the Princeton Ocean Model (POM model; Blumberg and Mellor, 1983) and adopted this model to San Francisco Bay (Uchiyama, 2005), which has extensive intertidal area including mudflats and deeper channels, though there is no mangrove. The model demonstrated that cohesive sediments are suspended dominantly in the deeper channels while being transported and deposited on intertidal areas, thereby self-scouring the deep channel and silting the intertidal areas. This model is useful for the R-type mangrove areas. Mazda et al. (1999) conducted numerical experiments based on the tidal trapping model to demonstrate that the material dispersion in an R-type forest depends on vegetation density (Figure 12). The water-born particles reside in the creek at
00:00 Low tide
03:00
15:00
06:00 High tide
18:00 High tide
09:00
21:00
12:00 Low tide
24:00 Low tide
10 cm/s
Figure 12 Tidal changes in current velocity (arrows) and tracer distribution within two tidal periods. At low tide (00:00) 9,000 tracers were released (Mazda et al., 1999).
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Dispersion
b a.b a
A B Vegetation density (drag force)
Figure 13 Sketch of the variation of the dispersion in the creek according to the vegetation density in the swamp. The dashed line (a) in the figure indicates the magnitude of material dispersion caused by a decrease in the inundation of water volume from the creek into the swamp with increasing vegetation density (volume effect). The dashed line (b) indicates the effect of an increase in the delay time of water discharging to the creek at ebb tide with increasing vegetation density (delay effect). The solid line ða bÞ is the combined effect due to the volume effect (a) and the delay effect (b), which shows the nonlinear relationship between vegetation density and material dispersion. The solid line records a minimum level of dispersion at Point A and a maximum level at Point B (Mazda et al., 1999).
low tide and disperse in the forest at rising tide. The magnitude of the dispersion depends on the vegetation density, that is, the drag force of vegetation, as shown in the solid curve in Figure 13. Considering the above nonlinear relationship between vegetation density and material dispersion, Mazda et al. (1999) proposed the following basic guideline regarding deforestation or thinning within mangrove areas. In natural forests, the vegetation density is maintained by the balance between the death of trees, the production of seeds and the dispersion of the seeds over many generations of mangrove trees. Artificial thinning disturbs this natural balance. Thinning leads not only to a reduction in vegetation density, which is accompanied by a reduction in the production of seeds (propagules), but also to a change in the dispersion of seeds. When the vegetation density after thinning lies within the range to the left of Point A in Figure 13 or in the range to the right of Point B, the dispersion of seeds is enhanced compared with that before thinning. In contrast, when the vegetation density after thinning lies within the range between Point A and Point B, a reduction in seed dispersion occurs. These considerations suggest that the question of whether the mangrove colony will progressively degenerate from generation to generation following thinning or whether it will recover to the prethinning vegetation density depends not only on the vegetation density at thinning but also on the dispersion characteristics of seeds following thinning. Mangrove trees shed thousands of seeds to the surrounding area; however, the fact that it is difficult to locate these seeds within the forest after a period of time is an indication of the effectiveness of the dispersion. Thus, when thinning is planned, the characteristics of seed dispersion should be taken into consideration. However, Figure 13 is just a schematic representation based on a simple model simulation, and the above discussion is purely conceptual. The actual positions of Points A and B
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for a real forest are unknown, because although the relationship between vegetation density and the drag force is presumed to depend on the species of mangrove, the details of the relationship are unknown. The above discussion is based on well-mixed flow conditions. Other processes of material dispersion can occur in density (salinity) stratified flows (Turrell and Simpson, 1988). Kuwabara (2002) emphasized the importance of physical characteristics for seed dispersion in tidal creeks, including the position of the centroid of floating seeds and the density of seeds relative to the surrounding water. These processes merit further study in modeling in view of their importance to mangrove ecology and their implications for environmental management.
4.3. Ecosystem model as the holistic system Mangrove ecosystems are formed by strong interactions or feedback relations shown in Figure 1. Shokita (1988) showed the interactions mainly between fishes, benthos, and detritus caused by mangrove litter. Robertson et al. (1992) proposed a qualitative model of food chains and carbon fluxes in mangrove forests and between mangrove forests and nearshore regions. In their models, the relationships between the biology and the physics such as water flow and sedimentation were described conceptually only, not quantitatively. To construct the ecosystem model as the holistic system, the hydraulic model and the material dispersion model mentioned in Sections 4.1 and 4.2, respectively, are linked. When linking these models, the following terms should be taken into account particularly from the physical viewpoint. 4.3.1. Feedback relation As shown in Figure 1, the four factors, namely, (1) biota which are composed of mangrove trees themselves, benthos such as mad crabs and algae, (2) sediment topography, (3) water flow such as tidal flow and sea waves, and (4) the atmosphere, play important roles individually for forming and maintaining the mangrove ecosystem. Further, every factor interacts one of other factors, as Mazda et al. (2007a) stated in detail. For example, the biology drives the physics of mangroves. The amount of water that inundates mangrove swamps depends on vegetation density in mangrove swamps, because the vegetation resists water inundation (Mazda et al., 1999). Further, the tide in mangrove swamps is measurably modified from that offshore due to resistance of mangrove vegetation (Mazda and Kamiyama, 2007). On the other hand, the physics drives the biology in mangroves. Watson (1928) proposed a simplified classification model to explain that the growth of mangrove trees and species zonation patterns depend strongly on the hydrological conditions such as the tides and the elevation of the substrate, as these factors control the flooding frequency, the duration of inundation and the depth of inundation (Bunt et al., 1985). Australian Institute of Marine Science is conducting laboratory studies about the dependence of the growth of mangrove seedlings on the inundation duration and salinity, and about the importance of bioturbation to ventilate the soils and prevent hypersalinity.
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Further, it should be understood that all of these interactions between arbitrary two factors construct the feedback system for maintaining the mangrove ecosystem. For example, the water flow associated with tides and rainfall helps to supply nutrients to mangrove trees. The mangrove trees, which grow with the help of solar radiation, accumulate their decayed leaves around their bottom substrate as sediments, which lead to the establishment of landforms. The landform or topography feeds back to modify the water flow by the drag force of the mangrove trees and roots. 4.3.2. Nonlinear interaction between factors with different timescales As mentioned in Section 3.2, the timescales of change in the biotic succession and the physical factors such as sedimentation and water flow are very different from each other. Thus, the relationship between the biology and the physics is nonlinear. In consideration of water inundation condition that undulates tidally and seasonally but with an annual period (Section 3.2), Snedaker (1989, personal communication) pointed out the importance of the statistical tidal information such as the frequency of tidal inundation, the tidal prism volume, the inundation duration and salinity range for the growth of mangrove forests. Lewis (2005) also described the importance of these statistical factors in controlling mangrove zonation. Mazda and Kamiyama (2007) discussed the nonlinear relationship between tidal deformation in mangrove swamps, which varies with seasonally changed tidal regime, and vegetation condition of mangroves, which changes in timescale of many decades. 4.3.3. Indirect and remote action We have experienced frequently that the human actions in a remote place or remote in time (e.g., damming rivers, embankment, and coastal reclamation) caused indirect influence on the natural environment such as the siltation of river mouth, coastal erosion, and hypersalinity (Mazda, 1984; Wolanski et al., 2001). Mazda et al. (2002) introduced that long-term human actions in an R-type mangrove forest have caused large-scale coastal erosion in a remote place and remote in time. Hong and San (1993) experienced that the embankments and barrages in the upstream regions have led to saltwater intrusion in the agricultural land, and the freshwater inflow into the mangrove system has resulted in mortality of trees along the seaward belt. However, these indirect effects have hardly been foreseen quantitatively. 4.3.4. Ecohydrology The mangrove environment should be understood as the total ecosystem composed of the river basin, the river, the estuary, and coastal waters forming an ecohydrology system (Wolanski et al., 2004). Mazda et al. (1990b) analyzed that the biota in the Bashita-Minato mangrove swamp, Iriomote Island, Japan, survive by the help of tidally supplying dissolved oxygen from the neighboring coral reef. Based on a model experiment applied to Darwin Harbor, Australia, Wolanski (2006a) suggested that the sustainable development of mangrove forests requires the adoption of this ecohydrological approach.
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The above findings need to be integrated into a mangrove ecosystem model. To construct this ecosystem model composed of four factors, biota, landform, water flow, and atmosphere (Figure 1), first we need to form models for each factor, secondly to form models of the interaction between every two factor, lastly to merge these models in a holistic ecosystem model. The problem is complex because these factors have different timescales and feedback with each other nonlinearly and indirectly through neighboring systems.
5. SUMMARY In this chapter, the present state and suggested future developments of modeling mangrove environments were summarized based mainly on hydrodynamic viewpoints. In comparison with the environment in temperate regions, the following peculiarities in mangrove areas should be noted, particularly when modeling the environment: 1. Mangrove areas consist of a very wide intertidal area, which is aquatic at high tide and terrestrial at low tide; thus, water movement in such an area cannot be treated as a continuous flow throughout the cycle of a tidal period. 2. The bottom substrates of mangrove swamps are formed by very loose sediment, which is readily eroded by water currents and sea waves. 3. Mangrove swamps present a large number of obstacles to water flow, including trees, prop roots, leaves, and pneumatophores, which affect the horizontal and vertical hydrodynamics. 4. Below ground, there are many macropores related to animal burrows and decayed roots, through which groundwater permeates and enables a belowground interaction between the swamp, the estuary, and the open sea. 5. In mangrove areas, there are various water motions with different timescales; these include seasonal changes in mean sea level (at timescales of months), tides (12 h), sea waves (<20 s), tsunami waves (10 min–1 h), and turbulence (<1 s). 6. The three mangrove forest types, R-, F-, and B-types, have different dominant water movements with one another. 7. In the R-type forests, the hydrodynamics is formed by the strong interaction between mangrove swamps and tidal creeks. 8. The mangrove environment or the ecosystem is sustained through intricate and tight feedback processes between biotic actions surrounding mangrove trees, landform with peculiar three-dimensional topographies, water flows over and under the ground and the surrounding atmosphere. 9. The biotic transitions in mangrove areas are formed through long term and nonlinear interactions with physical factors that operate at different timescales greatly from those of biotic life. If human actions impact on any part of the above feedback processes, the ecosystem may be destroyed or at least deteriorate. To recover and sustainably manage the mangrove environment, the mechanisms to sustain the environment
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must be understood scientifically and modeled quantitatively, in consideration of the above peculiarities in mangrove areas. In addition, the mangrove environment should be understood as the total ecosystem composed of the river basin, the river, the estuary, and coastal waters, affecting with each other directly and indirectly, further remotely both in points of space and time. In other words, for human needs at the river catchment scale an ecohydrological approach is required. To achieve this goal, the model needs the input of oceanographers, sedimentologists, chemists, biologists, ecologists, hydrologists, pedologists, dendrologists, entomologists, and geneticists.
ACKNOWLEDGMENTS We thank Mr. Keita Furukawa who permitted to use an original figure of his article and gave us useful information for numerical modeling in tidal flats. Our thanks also go to Dr. Yasuo Nihei for his kind help to collect references about nesting models.
REFERENCES Aucan, J., Ridd, P.V., 2000. Tidal asymmetry in creek surrounded by saltflats and mangroves with small swamp slopes. Wetlands Ecol. Manag. 8, 223–231. Blumberg, A.F., Mellor, G.L., 1983. Diagnostic and prognostic numerical circulation studies of the South Atlantic Bight. J. Geophys. Res. 88, 4579–4593. Boto, K.G., Bunt, J.S., 1981. Tidal export of particulate organic matter from a northern Australian mangrove system. Estuar. Coast. Shelf Sci. 13, 247–255. Bunt, J.S., Williams, W.T., Bunt, E.D., 1985. Mangrove species distribution in relation to tide at the seafront and up rivers. Aust. J. Mar. Freshwater Res. 36, 481–492. Cintron, G., Novelli, Y.S., 1984. Methods for studying mangrove structure. In: Snedaker, S.C., Snedaker, J.G. (Eds.), The Mangrove Ecosystem: Research Methods, UNESCO, Paris, pp. 91–113. Dean, R.G., Dalrymple, R.A., 2002. Coastal Processes with Engineering Applications. Cambridge University Press, Cambridge, 475pp. Furukawa, K., Wolanski, E., 1996. Sedimentation in mangrove forests. Mangroves Salt Marshes 1, 3–10. Furukawa, K., Wolanski, E., Mueller, H., 1997. Currents and sediment transport in mangrove forests. Estuar. Coast. Shelf Sci. 44, 301–310. Harada, K., Imamura, F., 2005. Effects of coastal forest on tsunami hazard mitigation – a preliminary investigation. In: Satake, K. (Ed.), Tsunamis: Case Studies and Recent Developments, Springer, The Netherlands, pp.279–292. Heron, S.F., Ridd, P.V., 2001. The use of computational fluid dynamics in predicting the tidal flushing of animal burrows. Estuar. Coast. Shelf Sci. 52, 411–421. Heron, S.F., Ridd, P.V., 2003. The effect of water density variations on the tidal flushing of animal burrows. Estuar. Coast. Shelf Sci. 58, 137–145. Hong, P.N., 2006. The Role of Mangrove and Coral Reef Ecosystems in Natural Disaster Mitigation and Coastal Life Improvement. Agricultural Publishing House, Hanoi, 385pp. Hong, P.N., Dao, Q.T.Q., 2003. Mangrove reforestation in Vietnam – achievements and challenges. In: Proceedings of National Scientific Workshop: Evaluation of Effects of Mangrove Reforestation
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Mazda, Y., Yokochi, H., Sato, Y., 1990a. Groundwater flow in the Bashita-Minato mangrove area, and its influence on water and bottom mud properties. Estuar. Coast. Shelf Sci. 31, 621–638. Miyagi, T., 1998. Mangrove Habitat Dynamics and Sea-Level Change. Tohoku Gakuin University, Sendai, 109pp. Nihei, Y., Nadaoka, K., Kumano, R., Sato, K., Machida, Y., 2001. A new multi-nesting approach for coastal ocean model. In: Proceedings of the 11th PAMS/JECSS, 365–368. Nihei, Y., Sato, K., Aoki, Y., Nishimura, T., Nadaoka, K., 2004. An application of a nesting procedure to a highly resolved current simulation in a mangrove area. APAC2003, CD-ROM, 1–8. Okubo, A., 1973. Effect of shoreline irregularities on streamwise dispersion in estuaries and other embayments. Neth. J. Sea Res. 6, 213–224. Raupach, M.R., Thom, A.S., 1981. Turbulence in and above plant canopies. Annu. Rev. Fluid Mech. 13, 97–129. Ridd, P.V., 1996. Flow through animal burrows in mangrove swamps. Estuar. Coast. Shelf Sci. 43, 617–625. Ridd, P.V., Sam, R., 1996. Profiling groundwater salt concentrations in mangrove swamps and tropical salt flats. Estuar. Coast. Shelf Sci. 43, 627–635. Ridd, P.V., Sandstrom, M.W., Wolanski, E., 1988. Outwelling from tropical tidal saltflats. Estuar. Coast. Shelf Sci. 26, 243–253. Ridd, P.V., Wolanski, E., Mazda, Y., 1990. Longitudinal diffusion in mangrove fringed tidal creeks. Estuar. Coast. Shelf Sci. 31, 541–554. Robertson, A.I., Alongi, D.M., 1992. Tropical Mangrove Ecosystems. Coastal and Estuarine Studies 41. American Geophysical Union, Washington, DC, 329pp. Robertson, A.I., Alongi, D.M., Boto, K.G., 1992. Food chains and carbon fluxes. In: Robertson, A.I., Alongi, D.M. (Eds.), Tropical Mangrove Ecosystems. Coastal and Estuarine Studies 41. American Geophysical Union, Washington, DC, pp.293–326. Sato, K., 1978. Studies on the protective functions of the mangrove forest against erosion and destruction. (1) The morphological characteristics of the root system of Yaeyamahirugi (Rhizophora mucronata Lamk.). Sci. Bull. Coll. Agric. Univ. Ryukyus 25, 615–630. Sato, K., 2003. Reality of sedimentation in mangrove forest by the tide and discharge – investigation on trapped amount of deposit in a serial of high tides. A summary on the Mangrove Study in Okinawa (FY2000-2002), Research Institute for Subtropics, 58–59. Shimizu, Y., Tsujimoto, T., Nakagawa, H., 1992. Numerical study on turbulent flow over rigid vegetation-covered bed in open channels. Proc. Jpn. Soc. Civ. Eng. 447/II-19, 35–44. Shokita, S., 1988. Aquaculture in Tropical Areas [in Japanese]. Midori Shobou, Naha, Japan. Snedaker, S.C., 1989. Overview of ecology of mangroves and information needs for Florida Bay. Bull. Mar. Sci. 44, 341–347. Spalding, M., Blasco, F., Field, C., 1997. World Mangrove Atlas. The International Society for Mangrove Ecosystems, Okinawa, 178pp. Stieglitz, T., Ridd, P.V., Muller, P., 2000. Passive irrigation and functional morphology of crustacean burrows in a tropical mangrove swamps. Hydrobiologia 421, 69–76. Struve, J., Falconer, R.A., Wu, Y., 2003. Influence of model mangrove trees on the hydrodynamics in a flume. Estuar. Coast. Shelf Sci. 58, 163–171. Susilo, A., 2004. Groundwater Flow in Arid Tropical Tidal Wetlands and Estuaries. PhD thesis, School of Mathematical and Physical Sciences, James Cook University, pp. 152. Susilo, A., Ridd, P.V., 2005. The bulk hydraulic conductivity of mangrove soil perforated with animal burrows. Wetlands Ecol. Manag. 13, 123–133. Turrell, W.R., Simpson, J.H., 1988. The measurement and modelling of axial convergence in shallow well-mixed estuaries. In: Dronkers, J., van Leussen, W. (Eds.), Physical Processes in Estuaries, Springer, Berlin, pp.130–145. Uchiyama, Y., 2004. Modeling wetting and drying scheme based on an extended logarithmic law for a three-dimensional sigma-coordinate coastal ocean model. Rep. Port Airport Res. Inst. 43 (4), 3–21. Uchiyama, Y., 2005. Modeling three-dimensional cohesive sediment transport and associated morphological variation in estuarine intertidal mudflats. Rep. Port Airport Res. Inst. 44 (1), 3–21.
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C H A P T E R
9
M ATHEMATICAL M ODELING OF T IDAL F LOW OVER S ALT M ARSHES AND T IDAL F LATS WITH A PPLICATIONS TO THE V ENICE L AGOON Luigi D’Alpaos, Luca Carniello, and Andrea Defina
Contents 1. Introduction 2. Wetting and Drying, and the Dynamics of Very Shallow Flows 3. Wind and Wind Waves 4. Salt Marsh Vegetation 5. Salt Marshes and Tidal Flats Morphodynamics 6. Conclusions Acknowledgments References
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1. INTRODUCTION In this chapter, attention is focused on the hydrodynamics and morphodynamics of salt marshes and tidal flats in shallow tidal lagoons. Shallow tidal basins are often characterized by extensive tidal flats and marshes dissected by an intricate network of channels (Fagherazzi et al., 1999; Rinaldo et al., 1999a,b; Defina, 2000a; Marani et al., 2003b). Both tidal flats and salt marshes are prevalently flat landforms located in the intertidal zone. Salt marshes have an elevation higher than the mean sea level and are periodically flooded by high tide. They are characterized by a very irregular surface and exhibit a dendritic and meandering structure of channels of varying sizes. These channels perform a drainage function, often continuing to flow long after the tide has receded and the marshes are exposed. They usually sustain a dense vegetation canopy of halophyte plants that withstand the relative infrequent flooding periods. Besides the biological processes related to plant colonization, such as biostabilization or soil production, vegetation sensitively affects the local hydrodynamics, reduces the bottom erosion, and increases sediment deposition. Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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Tidal flats are intertidal landforms usually flat or characterized by smooth sand waves. Occasionally, wide and shallow meandering channel are present. For the sake of simplicity, the term “tidal flat” is herein used in its broadest sense to include the “shallow sub-tidal flats,” that is, the muddy platforms that do not emerge during ordinary low tide. Salt marsh elevation is controlled by mineral and organogenic sediment accumulation (Pethick, 1981), sea-level variations, stabilizing effects of halophyte vegetation on its platform (Morris et al., 2002; Mudd et al., 2004; Silvestri et al., 2005), and the interaction between flora and fauna (Perillo et al., 2005; Minkoff et al., 2006). Tidal flats stem from a delicate balance between sediment deposition and erosion by wind waves and tidal currents (Allen and Duffy, 1998). Indeed, biology plays a nonnegligible role on the morphology of tidal flats (for a thorough review see Uncles, 2002). In particular, microphytobenthos and other organisms which colonize shallower areas affect the bottom shear stress threshold for sediment resuspension (Amos et al., 2004), and thus the morphological equilibrium of shallow tidal flats (Marani et al., 2007). However, besides the many studies which demonstrate the important interplay between biology and morphology on salt marshes and tidal flats, reliable mathematical models predicting the biological impact on flow dynamics and sediment processes are still lacking (Dietrich and Perron, 2006). The distinctive characters of tidal flats and salt marshes reflect on flow, wave field, transport and diffusion processes, and morphologic evolution as well. Therefore, different strategies must be followed when modeling the local hydrodynamics and morphology to maximize the accuracy and minimize the computational effort. Although considerable progress has been made in the application of twodimensional (2D) and three-dimensional (3D) models to simulate flow, waves and sediment transport in estuaries and coastal lagoons, a number of outstanding problems still remain in this branch of computational fluid dynamics. These problems mainly stem from the need to accurately model the key physical processes when dealing with very shallow flows, time-dependent flow domains, and complex topography. Adequate solutions cannot be just given by adopting accurate and well-structured numerical schemes, or else extremely refined computational grids. An important effort should instead be addressed toward modeling the relevant physical phenomena, which are neglected or drastically filtered by the numerical solution. This can be accomplished through the construction of suitable subgrid models, that is, by setting up a phenomenological representation of the overall processes which ensures a statistically equivalent description of the actual physics. Among others, wetting and drying of salt marshes and tidal flats, and the hydrodynamics of the small-scale drainage networks dissecting salt marshes are discussed in Section 2. Section 3 presents a simplified, computationally efficient wind wave model: the model generates and propagates a monochromatic wave inside tidal lagoons by solving the wave action conservation equation on an unstructured triangular mesh of arbitrary shape with a first-order
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land salt marsh tidal flat channel
0 1 2 3 4 5 6 7 8 km
N
Malamocco inlet
Lido inlet
Chioggia inlet
Adriatic Sea
Figure 1 Map of the Venice Lagoon. Symbols show the location of sites in the lagoon referred to in the text. Adapted from Defina (2000a).
finite volume explicit scheme. The problem of evaluating the bottom shear stress distribution due to the combined action of tidal currents and wind waves is also addressed in this section. Impact of salt marsh vegetation on tidal currents and wind waves is shortly discussed in Section 4. Morphodynamics modeling of salt marshes and tidal flats is discussed in Section 5, distinguishing between long- and short-term approaches. Finally, the main conclusions are summarized in Section 6. The examples presented in this chapter use the Venice Lagoon as typical irregular and shallow tidal basin. The Venice Lagoon is a wide tidal basin crossed by a network of deep channels departing from three inlets, namely Lido, Malamocco, and Chioggia (Figure 1). The lagoon is also characterized by the presence of wide tidal flats, small islands and salt marshes that exhibit a dendritic structure of channels of varying sizes (Rinaldo et al., 1999a,b; Defina, 2000a; Marani et al., 2003b).
2. WETTING AND D RYING, AND THE D YNAMICS OF VERY SHALLOW F LOWS The wetting and drying problem has received considerable attention during the last two decades. Recent reviews, mainly concerned with numerical aspects of this problem, can be found in the works of Balzano (1998), Bates and Hervouet (1999), and Bates and Horritt (2005). The wetting and drying problem can be handled either by adapting the numerical grid at each time step to follow the deforming flow domain (Lynch and Gray, 1980; Kawahara and Umetsu, 1986; Akanbi and Katopodes, 1988) or by retaining a fixed computational grid and
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utilizing some additional algorithms to deal with the hydrodynamics of partially wet elements. Due to the great difficulty of developing efficient deformable grid techniques, a fixed grid approach is by far preferable. In this case, a whole range of algorithms is available to identify wet elements and to control the flow over these (King and Roig, 1988; Leclerc et al., 1990; Falconer and Chen, 1991; Bates et al., 1992; Braschi et al., 1994; Defina et al., 1994; Hervouet and Janin, 1994; Defina and Zovatto, 1995; Ji et al., 2001; Oey, 2005). These algorithms are often intimately related to a particular numerical scheme and their application to a different numerical model is not straightforward (Balzano, 1998). Moreover, when dealing with very small water depths and wetting/drying of large areas, the major source of inaccuracy comes from the fact that numerical models approximate the bottom with a piecewise homogeneous plane surface. In this way they do not properly account for the effects due to the local variations of the flow field produced by small-scale topography (Defina et al., 1994; Bates and Hervouet, 1999; Defina, 2000a), thus yielding to approximate distributions of velocity and depth. The above problems can be partially overcome by setting up a phenomenological representation of the overall processes to supply a statistically equivalent description of the physics. To deal with partially wet and very irregular domains, an effective subgrid model of ground topography was developed by the authors in the early 1990s and improved over the years (Defina et al., 1994; Defina and Zovatto, 1995; Defina, 2000a). On considering bottom irregularities from a statistical point of view, and assuming the hydrostatic approximation, the 3D Reynolds equations have been phase averaged over a representative elementary area (REA) and then integrated over the depth. The averaged equations read (Defina, 2000a): Hh þ
1 d q þ J H Re = 0 g dt Y
ð1Þ
@h þH q = 0 @t
ð2Þ
ðhÞ
where h is the free surface elevation, g is the gravity, t is the time, q = (qx, qy) is the flow rate per unit width, Y the equivalent water depth, defined as the volume of water per unit area actually ponding the bottom, the local fraction of wetted domain and accounts for the actual area that can be wetted or dried during the tidal cycle, Re accounts for the horizontal turbulent stresses, and J = ( Jx, Jy) is energy dissipation per unit length due to bottom shear stress and vegetation, and energy gain due to wind shear stress acting on the free surface. J=
b þv w gY
ð3Þ
where b is bottom shear stress, v is an equivalent shear stress accounting for vegetation resistance, w is wind shear stress (see Section 3), and is fluid density.
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The bottom topography within a REA is assumed to be irregular with bottom elevations distributed according to a Gaussian probability density function. In this case the functions and Y are found to be (Defina, 2000a): 1 2D = 1 þ erf 2 ar ( " #) D 1 D 2 þ pffiffiffi exp 4 Y = ar ar ar 4 p
ð4Þ
ð5Þ
where erf() is the error function, ar is the typical height of bottom irregularities (i.e., the amplitude of bottom irregularities or, approximately twice the standard deviation of bottom elevations), and D = h – zb is the average water depth, zb being the average bottom elevation within a REA (Figure 2). For the case of turbulent flow over a rough wall, the energy dissipations due to bed shear stress can be written as (Defina, 2000a): 2 b n jqj q ð6Þ = gY H 10=3 where n is the Manning bed roughness coefficient and H is an equivalent water depth which can be approximated with the following interpolation formula (Defina, 2000a): rffiffiffiffi H Y Y 2Y =ar e ffi þ 0:27 ð7Þ ar ar ar The resulting subgrid model for ground irregularities requires the statistics of smallscale bottom topography. At present, remote sensing of topography (i.e., airborne laser altimetry, global positioning system-linked side-scan sonar and wide swath bathymetry) is proving very effective in providing high-resolution terrain data capable of parameterizing the proposed approach, even at subgrid scale (Bates and Hervouet, 1999; Bates et al., 2003, 2005). The above model proved very effective in the simulation of tide propagation in shallow lagoons, over salt marshes and tidal flats. Examples can be found in the literature (D’Alpaos et al., 1994; Defina and Zovatto, 1995; Defina, 2000a; Lanzoni and Seminara, 2002). All these examples clearly demonstrated the efficiency of the proposed equations. However a number of open issues need to be addressed: (1) the model does not account for water which may remain trapped within the REA during the drying phase. On the contrary, experimental evidence suggests that sometimes small pools or pans remain after the tidal wave recedes; (2) the model for bed shear stress (Equation (6)) neglects momentum exchange due to convective acceleration at the subgrid scale; and (3) Equation (6) for bed shear stress was derived on the assumption of an isotropic distribution of bottom irregularities. This is not always the case, often the creeks dissecting the marshes drive tidal
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Cros
s se
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Representative elementary area (REA)
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Figure 2 Representation of flow field and bottom topography within the representative elementary area (REA) when (a) the REA is completely wet and (b) the REA is partially wet. Adapted from Defina (2000a).
flow along preferential directions (Figure 3). The two latter issues are shortly discussed in the following paragraphs. The overall effects of momentum exchange due to convective acceleration at the subgrid scale can be accounted for by adjusting the friction coefficient. In fact, unresolved accelerations (i.e., subgrid accelerations) averaged over a sufficiently large area mostly produce extra dissipation, which is usually accounted for by suitably increasing the Manning friction coefficient. This is done during the model calibration step, since, at present, no relationships are available relating the friction coefficient to filtered velocity distributions. It would clearly be useful to have relationships allowing the a priori estimation of the appropriate roughness corrections. Finding suitable solutions to this problem is a quite demanding task
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Figure 3 Sample of creeks aligned in the E ^Wdirection, dissecting a salt marsh of the Venice Lagoon (site 1 in Figure 1). The gray-scale aerial photograph together with the extracted creek patterns are shown. Adapted from Marani et al. (2003b).
due to the large variety of mechanisms producing a spatially heterogeneous velocity field. Among the many, bottom topography in the presence of small water depths is possibly the easiest to handle. In this case, bottom topography generates small-scale momentum mixing thus enhancing energy dissipation. The problem can be handled in a way similar to the “mixing length” approach in turbulence and a relationship relating the ratio of the equivalent (neq) to the actual (n) Manning coefficient to water depth and bottom topography can be established (Defina, 2000b; D’Alpaos and Defina, 2007) neq 36ðY =ar Þ2 þ 53 = n 36ðY =ar Þ2 3 þ 6Cbw ðrÞ3 where Cbw(r) is an autocorrelation function of bottom elevations, given as Z zðxÞzðx þ rÞ dA z2wb Cbw ðrÞ = Aw 2wb
ð8Þ
ð9Þ
where zwb and bw are the average and the root mean square of bottom elevations within the wetted part of the REA, r is a horizontal “mixing length,” and z is local bottom elevation. Equation (8) for Cbw(r) = 0 and Cbw(r) = 1 is plotted in Figure 4. The behavior of neq/n is not symmetric about D/ar = 0. When D/ar << 0, that is, when the REA is nearly dry (Figure 2b), the flow field is characterized by a braiding pattern with the flow in each branch being independent from the others. In this case, momentum mixing is negligibly small and equivalent Manning coefficient recovers its original value (i.e., neq/n 1). When D/ar >> 1, that is, when the water depth is nearly uniform within the REA (Figure 2a), bottom irregularities have a minor impact on the velocity field and neq n. Note that when the bottom is smooth then Cbw(r) = 1 and neq = n.
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C bw(r) = 1 1.0 n/neq
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Figure 4 Computed equivalent Manning coefficient compared to theoretical prediction. Adapted from Defina (2000b).
The results of two numerical experiments (Defina, 2000b; D’Alpaos and Defina, 2007) are also plotted in Figure 4 for the sake of comparison. The numerical points lay within the region described by the two theoretical curves and the behavior of the numerical solution is in quite good agreement with the theoretical curve for Cbw(r) = 0. Importantly, the numerical solution approaches this limit curve as the Manning friction coefficient decreases. This is an expected result since increasing the Manning coefficient reduces friction and enhances large scale momentum mixing. As a consequence, the mixing length, r, increases and Cbw(r) decreases toward zero. Equation (8) must be considered as a first, promising attempt at quantifying the effects produced by small scale (i.e., subgrid) momentum mixing triggered by bottom irregularities. Indeed, further research is required to evaluate the mixing length r to be used in Equation (9). The second issue here shortly discussed focuses on the problem of modeling the creeks dissecting the marshes. The small-scale drainage network, comprising channels having a very small width cannot be resolved by a 2D model when the flow domain is comparably large as this would require a large number of very small computational elements. An example is shown in Figure 5 where a refined grid of the Venice Lagoon, comprising nearly 4 104 elements (Carniello et al., 2005), overlaps an aerial photograph of a marsh zone in the northern part of the lagoon. A tangle of small, highly meandering creeks with a width in range between 0.1 and 2 m covers most of the marsh surface. The mesh resolution required to describe all these channels is beyond present computational capabilities. In this case, and as a first approximation, the smaller channels can be treated as “topographic irregularities.” A clear dividing line between actual ground irregularities, which are expected to behave quasi-isotropically, and small-scale channels and creeks, cannot be traced, as it depends on the domain extension and on the required accuracy. Here the attention is focused on channels with a size which is small enough to prevent the use of 2D elements for their description, and large enough to dissuade
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Figure 5 Refined computational grid overlapping a marsh area in the northern Venice Lagoon near Treporti (site 2 in Figure 1). Adapted from D’Alpaos and Defina (2007).
one from including them as bottom irregularities. These narrow channels are usually very numerous and their importance turns out to be comparable to that of large channels (Defina, 2004). The problem of accounting for this channel network can be tackled by observing that the flow within relatively deep and narrow channels flanked by shallow intertidal areas exhibits a distinctive onedimensional (1D) character, thus suggesting the use of 1D elements to include them in the model. The problem of coupling 1D and 2D elements has sometimes been found in the literature, for example SOBEK, developed by WL|Delft Hydraulics (http:// www.sobek.nl) and LISFLOOD-FP (Bates and De Roo, 2000; Horritt and Bates, 2001a,b; Bates et al., 2005). To keep a high accuracy and reduce the computational effort, a particular way of coupling a 2D model to describe the shallow water hydrodynamics with a 1D model to simulate the flow in the channels has been proposed by D’Alpaos and Defina (1993, 1995, 2007). In the model,
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channels are superimposed on the 2D domain. The effects due to the momentum exchange between the channel and the two-dimensional flow are neglected. In the model, the 2D and 1D flow equations are solved by a finite element scheme. The domain is divided into triangular and linear elements with each channel lying along the common side of two adjacent triangular elements. In this way each channel can be added to or removed from the domain without any change in the main 2D discretization. The number of nodes in the computational grid remains unchanged and the computational effort is only slightly increased due to inclusion of 1D elements. In the example shown in Figure 5, Canale S. Felice, which is a very large channel, is described using 2D triangular elements, the very small creeks dissecting the tidal marsh are included in the model as bottom irregularities, while the large creek through the salt marsh, departing from Canale S. Felice, is described with 1D elements (dotted segments) aligned along the edges of the 2D triangular elements. The model has been tested against the numerical solution computed with a 2D model for a number of test cases (D’Alpaos and Defina, 1993, 1995, 2007; D’Alpaos et al., 1995) and proved very effective.
3. W IND AND W IND W AVES Surface gravity waves are one of the most important phenomena in shallow, coastal lagoons and estuaries. Estimation of wave characteristics in estuaries, tidal basins and coastal areas is essential to analyze sediment transport and local shoreline erosion processes (Anderson, 1972; Ward et al., 1984; Shoelhamer, 1995; Mo¨ller et al., 1999; Umgiesser et al., 2004; Carniello et al., 2005). Swell waves approaching the coast from the open sea are relevant to study the shoreline morphodynamic in coastal areas. On the contrary, the locally generated wave field is of importance for lagoonal morphodynamics: on the one hand, wind wave-induced bottom shear stress is the decisive process mobilizing tidal flat sediments (Carniello et al., 2005) and influencing their equilibrium configuration (Fagherazzi et al., 2006; see also Section 5); on the other hand salt marsh, because of their elevation and the presence of halophytic vegetation, greatly affect wind wave field by reducing wave energy (see Section 4). Two alternative methods are available to model wind wave generation and propagation, that is, a phase-resolving approach, based on mass and momentum balance equations (for a review see Dingemans, 1997); or a phase-averaged approach that solves the energy or wave action balance equation (e.g., Booij et al., 1999). Phase-resolving models reproduce the sea surface in space and time and account for effects such as refraction and diffraction. Bottom friction and depth-induced wave breaking can be included in the model but wind wave generation is usually absent or poorly reproduced. Phase-resolving models are thus unsuitable in lagoons where storm conditions and local wave generation are key processes. Furthermore, space and time resolutions required by phase-resolving models are of the order of
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a fraction of the wavelength and wave period, respectively, thus restricting their use to small domains and short duration events. For large-scale applications phase-averaged models are by far more suitable. Since the pioneering work of Gelci et al. (1956), many models that use the phase-averaged approach have been developed. Among them: the GLERL model developed by Donelan (1977) and revised by Schwab et al. (1984); the HISWA model (HIndcast Shallow water WAves model) (see Holthuijsen et al., 1989), and its successor the SWAN model (Simulating WAve Nearshore) (Booij et al., 1999; Ris et al., 1999); the WAVAD model (Resio, 1987; Resio and Pierre, 1989), and the ACES model (Automated Coastal Engineering System) (Leenknecht et al., 1992). Lin et al. (1998) tested all the models mentioned above against a wind and wave data set collected in the northern Chesapeake Bay, USA, during September 1992, when the tropical storm Danielle passed over the area. They found that no single model seems to be good at predicting all aspects of the surface wave field in that specific and morphologically irregular domain, but the GLERL and SWAN models were the most promising. Moreover, in shallow basins, the instantaneous local water depth is crucial to correctly predict the wave field, since water depths strongly affects wave propagation. Wave prediction can therefore be accomplished only by coupling a wave model with a hydrodynamic model. Umgiesser et al. (2004) moved a preliminary step toward this direction. They combined a 2D finite elements model with the finite difference SWAN model run in stationary mode. For consistency, all the results produced by the hydrodynamic model were interpolated to the grid of the wave model, thus introducing significant numerical approximations. Since shallow tidal basins have a very irregular morphology with large and sudden changes in bottom elevation, islands, and salt marshes which are periodically flooded and exposed, a specific framework must be adopted to model wind wave propagation in these environments. A simplified, computationally efficient model has been recently developed (Carniello et al., 2005). The wind wave module solves for the conservation of the wave action (Hasselmann et al., 1973), defined as the ratio of the wave energy density E to the wave frequency , using a first-order finite volume explicit scheme. The wave model is coupled with a finite element hydrodynamic model (D’Alpaos and Defina, 1995, 2007) sharing the same computational grid. At each time step, the hydrodynamic model yields nodal water levels which are used by the wind wave model to assess wave group celerity and wave energy dissipations. The wind wave model propagates a monochromatic wave, neglects nonlinear wave–wave and wave–current interactions and assumes that the direction of wave propagation instantaneously adjusts to the wind direction (Carniello et al., 2005). The wave action conservation equation can thus be written as @E þ H ðcg EÞ = Sw Sbf Swc Sbrk @t
ð10Þ
The first term of Equation (10) is the local rate of change of wave energy density in time, the second term represents the energy convection, cg = (cgx, cgy) being the
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wave group velocity. The source terms on the right-hand side of Equation (10) account for the wind energy input (Sw), the energy dissipation by bottom friction (Sbf) and by whitecapping (Swc), and the energy dissipation by depth-induced breaking (Sbf). The significant wave height, H, is then computed using the linear theory as sffiffiffiffiffiffiffiffi E H= 8 g
ð11Þ
Examples showing the impact of the wind action on the hydrodynamics and the generated wave field inside the Venice Lagoon under different wind and tidal conditions are briefly discussed. All simulations presented here are performed using a refined mesh reproducing the present topography of the Venice Lagoon (Carniello et al., 2005). Numerical results are compared with field measurements available at two different stations inside the lagoon: station 1BF is on a shoal in the northern part of the lagoon and station 2BF is in a deeper area in the southern part of the lagoon (Fondo dei Sette Morti) (Figure 1). Two stormy events (i.e., 3 April 2003 – see Figure 6a, and 16–17 February 2003) characterized by the Bora wind blowing from the northeast with a speed in the range between 12 and 16 m/s are simulated. The Bora wind affects near-coast sea levels resulting in lower elevations at the Lido inlet than at the Chioggia inlet (Figure 6b). Importantly, wind setup strongly affects the hydrodynamics within the lagoon. This is shown in Figure 6e,f where the measured water levels at 1BF and 2BF (see Figure 1) are compared with the water levels computed with the hydrodynamic model when wind shear stresses are included in the model or neglected. Neglecting wind shear stress produces water levels that do not fit the measured ones. The impact of wind action on the hydrodynamics is even more evident when the flow rate through the three inlets is considered. When including wind shear stress, the computed flow rate through the Lido inlet increases during flood and a decreases during ebb (Figure 6c). The opposite occurs at the Chioggia Inlet (Figure 6d), while no substantial changes occur at the Malamocco inlet. Overall, wind setup during Bora stormy conditions produces a residual current flowing from the Lido inlet toward the Chioggia inlet. Figure 7 compares the computed significant wave height with the measured one at stations 1BF and 2BF. The agreement for both storm events is quite good. Plotted wave heights follow a sinusoidal-like variation having the same phase as the tidal oscillation, confirming the influence of water level on wave height and the strong feedback between wind waves and hydrodynamics. The result is even clearer in the lower panel of Figure 7 where an example of the computed wave field at low tide and high tide during the stormy event of 16–17 February 2003 is given. The complementary effect of tidal currents and wind waves on bottom shear stresses is crucial to predict the morphodynamic evolution of tidal flats in shallow
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Figure 6 Stormy event of 3 April 2003.Wind velocity and direction (a); tidal level at the three inlets (b); computed flow rate through the Lido inlet (c) and the Chioggia inlet (d) when wind action is included in the model or neglected; and comparison between computed and measured water levels at stations 1BF (e) and 2BF (f ). Adapted from Carniello et al. (2005).
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Figure 7 Comparison of measured (solid and dashed lines) and computed (circles) significant wave height at 1BF and 2BF stations (Figure 1) during 3 April 2003 (a) and 16^17 February 2003 (b). Lower panel: computed wave field during stormy event of 16^17 February 2003 at low tide (A) and high tide (B). Adapted from Carniello et al. (2005).
tidal basins (see Section 5). Bottom shear stress due to waves ( b,wave) is computed by the wind wave-tidal model as: 0:52 1 pH um T 2 b;wave = fw um with um = ð12Þ and fw =1:39 2 T sinhðkY Þ 2pðD50 =12Þ
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where um is the maximum horizontal orbital velocity at the bottom according to the linear theory, fw is the wave friction factor as given by Soulsby (1997), T is the wave period, k is the wave number, and D50 is the median grain diameter. Since maximum shear stress max, rather than average stress m, is responsible for the bottom sediments mobilization, all the results presented and discussed herein after refer to the maximum total bottom shear stress, which is evaluated with the empirical formulation suggested by Soulsby (1997): 1=2 max = ð m þ b;wave cos Þ2 þ ð b;wave sin Þ2 " m = b
b;wave 1 þ 1:2 b þ b;wave
ð13Þ
3:2 # ð14Þ
where b is given by Equation (6), and is the angle between the current and the wave directions. Figure 8 shows the time evolution of the bottom shear stress at three different sites within the lagoon during the stormy event of 16–17 February 2003. The sites are chosen in a deep channel close to the Lido inlet (site 1H in Figure 1), on a tidal flat close to the Murano island (site 3H in Figure 1), and on a tidal flat next to the Casse di Colmata (site 2H in Figure 1). Each plot compares model results obtained with three different simulations, that is, (1) the hydrodynamics is forced by the recorded tidal levels at the three inlets, wind shear stress and wind waves are neglected; (2) the wind shear stress are included whereas wind waves are not; and (3) both wind shear stress and wind waves are included in the model. The results show that in deep channels wind waves slightly affect bottom shear stresses (Figure 8a), while no influence of wind stresses at the surface can be observed. On the contrary, bottom shear stresses on tidal flats are strongly enhanced when wind waves are included in the model (Figure 8b,c). In this case, wind shear stress gives a minor contribution, feebly enhancing the bottom shear stresses produced by tidal currents. In shallow areas the bottom shear stresses exceed the critical value ( cr ffi 0.7 Pa, Amos et al., 2004) for sediment erosion only in the presence of waves. On the contrary, the bottom shear stresses are always smaller than the critical value when wind waves are not included in the model. This result is further supported by Figure 8d,e mapping the regions where the bottom shear stress exceeds cr. No resuspension is possible on tidal flats and salt marshes if wind waves are not considered. Figure 8d,e also confirms that wind wave resuspension is complementary to tidal current resuspension since waves are able to produce high bed shear stresses in shallower areas whereas bed shear stresses due to tidal currents are high only for the deep channels where the tidal flow concentrates.
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Figure 8 Storm event of 16^17 February 2003: comparison of the shear stress at the bottom produced by the combined effect of wind waves and tidal currents (- - -), by tidal currents when the wind shear stress at the free surface is included (. . .. . .), or neglected (ççç). The comparison refers, respectively, to the Lido inlet (site 1H in Figure 1) (a), a tidal flat close to Murano Island (site 3H in Figure 1) (b), and a tidal flat close to “Casse di Colmata” (site 2H in Figure 1) (c). Spatial distribution of the area experiencing a bottom shear stress greater than 0.7 Pa inside the Venice Lagoon. The simultaneous effect of tidal currents and wind waves (d) is compared to the effect of tidal currents alone (e). The dotted line (wind ^ no waves) in panel as is hidden by the solid line (wind ^ waves). Adapted from Carniello et al. (2005).
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4. SALT MARSH V EGETATION Tidal marshes are colonized by halophytic vegetation, that is, macrophytes adapted to complete their life cycle in salty environments. Vegetation has a strong impact on the hydrodynamics over salt marshes as it affects both tidal current (Burke and Stolzenbach, 1983; Kadlec, 1990; Leonard and Luther, 1995; Shi et al., 1995; Dunn et al., 1996; Nepf and Vivoni, 1999, 2000; Neumeier and Amos, 2006; Neumeier, 2007) and wind waves (Wayne, 1976; Knutson et al., 1982; Pethick, 1992; Koch and Gust; 1999; Mo¨ller et al., 1999; Mo¨ller and Spencer, 2002; Swales et al., 2004; Mo¨ller, 2006). Moreover, vegetation reduces bed shear stress, hence erosion, and strongly affects transport and diffusion processes (Lopez and Garcia, 1998; Nepf, 1999; Nepf and Koch, 1999; Leonard and Reed, 2002; Bouma et al., 2007). Resistance to flow produced by vegetation can be included in the hydrodynamic model as an additional, equivalent shear stress v [see Equation (3)]. To compute the equivalent shear stress v any model (Shimizu and Tsujimoto, 1994; Klopstra et al., 1997; Lopez and Garcia, 2001; Righetti and Armanini, 2002; Defina and Bixio, 2005) able to predict the velocity profile in a uniform flow in the presence of vegetation can be used. Here we focus on the case of rigid vegetation and consider a uniform flow in the x-direction. In this case the velocity profile ux(z), z being the vertical direction, can be written as pffiffiffiffiffiffiffi ð15Þ ux ðzÞ = f ðCD ; m; Az ; hp ; Y Þ S0x where hp is plant height, CD the drag coefficient, Az the frontal area of vegetation per unit depth, m the number of stems per unit area, S0x the bottom slope, and f a function of water depth and vegetation characteristics. The flow rate per unit width is then given as pffiffiffiffiffiffiffi qx = S0x
Zh
pffiffiffiffiffiffiffi f ðCD ; m; Az ; hp ; Y Þ dz = S0x FðCD ; m; Az ; hp ; Y Þ
ð16Þ
zz
Recalling that x = gYS0x, extension to 2D flow gives gY jqj q v = F2
ð17Þ
Once the velocity profile ux(z) is computed for a given slope S0x, function F can be easily computed from Equation (15). Zh FðCD ; m; Az ; hp ; Y Þ = zb
ux ðzÞ pffiffiffiffiffiffiffi dz S0x
ð18Þ
Figure 9 shows the behavior of the function F for three vegetation species which colonize the salt marshes of the Venice Lagoon. In this case the velocity profiles
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Figure 9 Behaviors of CDAz for Salicornia veneta (left panel) and function F as given by Equation (18) for three different species: Spartina (m = 350), Salicornia veneta (m = 45), and Limonium (m = 100).
under uniform flow condition have been computed using the model proposed by Defina and Bixio (2005). While the above results illustrate the effects of vegetation in producing additional flow resistance, the link between vegetation parameters and wave transformation remains qualitative in the absence of well experimented quantitative relationships between vegetation structure and wave energy. Wave–vegetation interaction has been investigated to predict wave attenuation produced by vegetation. Standard approaches to model wave attenuation by vegetation are based on the time-averaged conservation equation of wave energy and assume linear wave theory or linearized momentum equations to describe the local flow field (Dalrymple et al., 1984; Kobayashi et al., 1993). For the case of small amplitude monochromatic waves impacting an array of rigid vertical cylinders of diameter Az, Kobayashi et al. (1993) proposed the following dissipation term to be added to the source term of Equation (10). pffiffiffiffiffiffiffiffiffiffiffi 3 4 2g3 = k sinh3 ðkhp Þ þ 3 sinh ðkhp Þ 3=2 Sveg = ð19Þ E CD mAz 3k cosh 3 ðkY Þ ! 3p The above equation can be extended to describe conditions of emergent vegetation by substituting hp with Y. Wave number k in Equation (19) depends not only on water depth and wave period but on vegetation characteristics as well. However, in the limit of small wave energy damping, the standard dispersion equation based on linear wave theory can be used to compute k (Kobayashi et al., 1993). The above model for wave dissipation due to vegetation suffers a number of shortcomings: (1) in the model, the flow depth is split into a lower layer containing the vegetation and a surface layer; velocity profiles are described separately for the vegetation layer and the surface layer, reflecting the different physical phenomena acting in the two layers; however, they do not match at the interface; (2) in the model turbulent shear stresses are neglected; (3) the model is not sufficiently tested (model results were compared only with experiments conducted by Asano et al.
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(1988) who used a very flexible artificial vegetation); and (4) Equation (16) does not account for the variation in both CD and Az along z. It is thus clear that the above model is deserving of further research. Once the resistance to flow and the wave energy dissipation due to vegetation have been modeled, the problem of assessing the spatial distribution of vegetation must be addressed. Halophytic vegetation over salt marshes is not randomly distributed nor spatially uncorrelated but is, on the contrary, organized in characteristic patches (Pignatti, 1966; Chapman, 1976; Silvestri et al., 2000, 2005; Marani et al., 2004). Therefore wide areas, which may extend over a few computational elements, are colonized by the same species, that is, a set of parameters describing a single vegetation species can be associated to each computational element. Quantitative remote sensing, integrated with field observations, proved very effective for mapping the different species (Marani et al., 2003a, 2006; Silvestri and Marani, 2004; Belluco et al., 2006) thus allowing for a very accurate parameterization of vegetation in the numerical model.
5. SALT M ARSHES AND T IDAL FLATS M ORPHODYNAMICS Different approaches are usually adopted to model short- and long-term morphologic evolution of coastal and lagoonal environments. Long-term models were first introduced to investigate salt marsh formation and evolution. In the pioneering point model suggested by Krone (1987), changes in marsh elevation are calculated as a function of sediment concentration, settling velocity of the suspended sediment flocs, and hydroperiod. When the marsh platform becomes emergent the inundation period decreases, so that less sediment has time to deposit leading to a reduction of marsh accretion. The model was then improved by considering sediment supply and sea-level rise (Allen, 1990; French, 1993), sediment composition (Allen, 1995), differences in sedimentation rates between creek levees and marsh platform (Temmerman et al., 2004a), and variations in sediment concentration as a function of tidal inundation (Temmerman et al., 2004b). In recent years a major development has been the inclusion in the marsh model of the vegetation effects on sediment dynamics, accumulation rates, and organogenic production by linking all these processes to the biomass of halophyte vegetation that colonizes the marsh surface (Morris et al., 2002; Mudd et al., 2004; D’Alpaos et al., 2006). Besides the vertical accretion of the marsh platform, the latest long-term analysis of salt marsh morphologic evolution consider the formation and the planimetric development of the tidal creek network. A model describing this complex morphodynamic process taking also into account the importance of vegetation distribution, and the consequence of marine transgression and regression, is presented and discussed in D’Alpaos et al. (2009). However, these models disregard the incipient formation of salt marshes. As a consequence, they can only be applied to locations in which the salt marsh is already present, but are ineffective in determining under what conditions the salt
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marsh has evolved from tidal flats. On the contrary, the evolution of tidal flats is less considered in the literature. An original contribution to tidal flat morphodynamics is given by Fagherazzi et al. (2006) and is based on the observation that the bathymetric data points out an abrupt transition between salt marshes and tidal flats with very few areas lying at intermediate elevations. To describe this evidence, Fagherazzi et al. (2006) developed a conceptual model which indicates that this bimodal distribution of elevations strictly relates to wind wave shear stresses. It has been demonstrated (Carniello et al., 2005), in fact, that the role of sediment resuspension by wind waves is decisive in shallow tidal basins, whereas tidal fluxes alone are unable to produce the bottom shear stresses necessary to mobilize tidal flat sediments. The conceptual model follows from the wave model described in Section 3. It mainly assumes (1) that wind waves are the main source of bottom shear stress (i.e., the model does not apply to tidal channels where bottom shear stress is mainly due to tidal current), (2) that in shallow basins waves quickly adapt to external forcing, and (3) the fetch required to attain fully developed condition is short. Therefore, as a first approximation, the conservation Equation (10) can be reduced to the local equilibrium between the source terms describing the unlimited fetch fully developed local wave field. The conceptual model is based on the stability curve obtained by plotting the wind wave-induced bottom shear stress as a function of water depth (Figure 10a). The model assumes that the rate of sediment erosion, ES, is proportional to the difference between bottom shear stress ( b) and the critical shear stress for sediment erosion ( cr). Therefore, the curve of Figure 10a is a proxy for bed erosion rate. The model further assumes some prescribed average annual sedimentation rate, DS. Dynamic equilibrium, ES = DS is achieved when b = eq (points U and S in Figure 10a). When b < eq then deposition exceeds erosion and the bottom evolves toward higher elevations. On the contrary, when b > eq, then erosion exceeds deposition and the bottom sinks toward lower elevations. Therefore, any point S (a)
(b) Unstable
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U
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0.0
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–2.0 –3.0 Z b (m a.m.s.l.)
Figure 10 (a) Bed shear stress distribution as a function of bottom elevation; (b) frequency area distributions as a function of bottom elevation for the Southern Venice lagoon (1901 and 2000 bathymetries).
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along the right branch of the curve is a stable point while any point U on the left branch of the curve is an unstable point. The conceptual model demonstrates that a stable morphodynamic equilibrium is possible only for salt marshes (i.e., Zb >Zc1) and tidal flats (i.e., Zc2 < Zb < Zmax). Recently we tested the conceptual model through comparison with numerical results obtained with the 2D wind waves – tidal model described in Section 3 (Defina et al., 2007). We performed two numerical simulations with two different computational grids describing the present and the 1901 bathymetries of the Venice Lagoon. The analysis focuses on the central–southern part of the lagoon where the condition of fully developed wave field establishes over most of the domain. The computed bottom shear stress plotted versus bottom elevation shows a remarkable concentration of points (corresponding to the 79.5% of the total area analyzed) around the theoretical curve (see Figure 3 in Defina et al., 2007). Whereas the few points which do not cluster along the curve pertain to tidal channels or to fetch limited areas, that is, regions that do not meet the main model hypotheses. We further showed that all points falling along the stable branch of the curve are indeed tidal flats. The few points on the unstable branch of the curve (corresponding to less than 10% of the total area) are located on tidal flats close to salt marsh edges where the lagoon morphology is likely far from equilibrium since salt marshes are progressively reducing their extension. After removing the few points that do not meet the model assumptions, the bottom elevation density functions of the two bathymetries of the Venice Lagoon are evaluated (Figure 10b). The curves show a minimum corresponding to elevations in the unstable range thus confirming that just a small fraction of the basin is characterized by these intermediate elevations, in agreement with the conceptual model. More consistent with the approach assumed in the present chapter is the shortterm morphological evolution of tidal basins. Such analysis requires to take into account all the processes acting at the daily timescale while processes such as organic soil production, soil compaction, eustatism, and sea-level rise, extremely important in long-term evolution, are usually neglected. Short-term bottom evolution of shallow tidal basins can be studied by coupling an hydrodynamic model which includes wave dynamics with a sediment transport model which includes one or a set of equations for bed evolution. In the following is a short description of the sediment transport model which is being implemented to study the short- and mid-term morphodynamic evolution of the Venice Lagoon. The bed composition of the Venice Lagoon is characterized by cohesive clayey silt with the exception of the bigger channels branching from the three inlets (see Figure 1) which are sandy and noncohesive (Amos et al., 2004). In the model we use two sedimentological classes: fine sand as a proxy of pure noncohesive sediments and mud (grain size less than 0.063 mm, i.e., silt and clay) as a proxy of pure cohesive sediments. We further assume an average grain size ds50 = 150 mm for the sand class and the grain size dm50 = 20 mm for the mud class. The model considers a 10% of mud content by dry weight to discriminate between noncohesive and cohesive behavior (van Ledden, 2003; van Ledden et al., 2004). The actual bed composition is locally computed by the model as a mix of the two sedimentological classes.
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The sediment transport model neglects the horizontal diffusion which is small compared to advection (Pritchard and Hogg, 2003) and solves the advection equation for different sediment classes: @Ci Y þ HqCi = Esand þ Emud Dsand Dmud @t
ð20Þ
where C (m3/m3) is the depth-averaged sediment concentration, Esand and Emud (m/s) are the entrainment of sand and mud computed according to the equations suggested by van Rijn (1993), van Ledden (2003), and van Ledden et al. (2004), which account for the different possible behaviors (i.e., noncohesive or cohesive) of the sand–mud mixture, Dsand (m/s) is the deposition rate for noncohesive sediments which is proportional to the local sand concentration and still water settling velocity, and Dmud (m/s) is the deposition rate for cohesive mud evaluated according to the Krone’s formula. Equation (20) is solved for each of the two sedimentological classes with a firstorder finite volume explicit scheme to obtain the time and spatial evolution of suspended sand and mud concentrations. Importantly, all the models share the same computational grid. A specific bed evolution module, based on the mixing layer concept (Hirano, 1971, 1972), has been developed to predict the time variation of bed elevation and bed composition as a consequence of sand/mud deposition and erosion. Bed elevation is governed by the following equation: ð1 nÞ
@zb = Dsand þ Dmud Esand Emud @t
ð21Þ
where zb is the bed elevation and n is the bed porosity.
Concentration (m3/m3 )
10–5
10–6 1BF measured 1BF computed 1BF computed (no waves)
10–7 10
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6
3/04/2003
12
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–4
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0
17/02/2003
3
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9 Time (h) 12
–4
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10–6 10–7 10–8
1BF measured 1BF computed 1BF computed (no waves)
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24
Concentration (m3/m3 )
Concentration (m3/m3 )
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6
3/04/2003
12
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Time (h)
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10–5 2BF measured 2BF computed 2BF computed (no waves)
10–6 10–7 10–8
0
17/02/2003
3
6
9 Time (h) 12
Figure 11 Comparison of measured and suspended sediment concentration at station 1BF and 2BF during the stormy events of 3 April 2003 (left) and 16^17 February 2003 (right).
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At each time step the bed evolution model calculates the net variation of bed elevation distinguishing between sand and mud contribution. Based on the deposition and erosion rates of sand and mud, the composition of the active (mixing) layer is then updated. Results obtained in some preliminary runs are shown in Figure 11. The simulated events are the same discussed in Section 3 (i.e., 3 April 2003, and 16–17 February 2003). Figure 11 compares measured mud concentration with concentration computed when the hydrodynamic model includes or not the wind wave module.
6. C ONCLUSIONS This review has attempted to make an examination of many aspects that must be considered when modeling the hydrodynamics and the morphodynamics of salt marshes and tidal flats in shallow tidal lagoons. While it is clear that considerable progress has been made in the development and application of shallow water models to simulate flow, waves, and sediment transport in estuaries and coastal lagoons it is also clear that, in this branch of computational fluid dynamics, many outstanding problems of physical process representation still remain for the future. Among the many, a weak area is that of short-term morphodynamics modeling. Algorithms for calculating erosion, transport, and deposition of multiple sediment classes and evolution of sediment stratigraphy caused by wind waves and currents in tidal environments need improvements. Feedback between hydrodynamics, biology, and morphology represents a further crucial aspect to be dealt with when modeling the tidal flow over salt marshes and tidal flats. This is especially true with respect to the study of morphodynamic equilibrium. To this end, algorithms to describe each specific biological process, at tidal timescale, deserve to be developed whereas long-term hydrobiological models still need to be improved. In this work it is shown that adequate solutions to many modeling problems can be accomplished through the construction of suitable subgrid models, that is, by setting up a phenomenological representation of the overall processes which provides a statistically equivalent description of the actual physics. This approach also simplifies the coupling of different models conceived to describe specific physical processes coming from different disciplines (e.g., hydraulics, hydrology, morphology, biology, ecology). Finally, setup and validation procedures based on spatially distributed field data (e.g., wind and wave fields, spatial distribution of bottom sediments and sediment concentration, vegetation seasonal patterns), is a key task to future model design in environmental science. To this end remote sensing, which have become increasingly popular over the past few years owing to large advances in the technology sector, is expected to be a very helpful tool.
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ACKNOWLEDGMENTS This research has been funded by Comune di Venezia “Modificazioni morfologiche della laguna, perdita e reintroduzione dei sedimenti.” The authors wish to thank, among others, Bruno Matticchio and Sergio Fagherazzi for their contribution in conceiving and developing part of the presented models.
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Morris, J.T., Sundareshwar, P.V., Nietch, C.T., Kjerfve, B., Cahoon, D.R., 2002. Responses of coastal wetlands to rising sea level. Ecology 83, 2869–2877. Mudd, S.M., Fagherazzi, S., Morris, J.T., Furbish, D.J., 2004. Flow, sedimentation, and biomass production on a vegetated salt marsh in South Carolina: toward a predictive model of marsh morphologic and ecologicevolution. In: Fagherazzi, S., Marani, M., Blum, L.K. (Eds.), The Ecogeomorphology of Tidal Marshes. Coastal and Estuarine Studies, vol. 59.American Geophysical Union, Washington DC, pp. 165–188. Oey, L.-Y., 2005. A wetting and drying scheme for POM. Ocean Modell. 9, 133–150. Nepf, H.M., 1999. Drag, turbulence and diffusion in flow through emergent vegetation. Water Resour. Res. 35 (2), 479–489. Nepf, H.M., Koch, E.W., 1999. Vertical secondary flows in submersed plant-like arrays. Limnol. Oceanogr. 44 (4), 1072–1080. Nepf, H.M., Vivoni, E.R., 1999. Turbulence structure in depth-limited vegetated flows: transition between emergent and submerged regimes. In: Proceedings of the XXVIII IAHR Congress, Graz, Austria, pp. 1–8. ISBN 3-901351-34-5. Nepf, H.M., Vivoni, E.R., 2000. Flow structure in depth-limited, vegetated flow. J. Geophys. Res. 105 (C12), 28547–28557. Neumeier, U., 2007. Velocity and turbulence variations at the edge of salt marshes. Cont. Shelf Res. 27, 1046–1059. doi:10.1016/j.csr.2005.07.009. Neumeier, U., Amos, C.L., 2006. The influence of vegetation on turbulence and flow velocities in European salt marshes. Sedimentology 53, 259–277. Perillo, G.M.E., Minkoff, D.R., Piccolo, M.C., 2005. Novel mechanism of stream formation in coastal wetlands by crab-fish-groundwater interaction. Geo-Mar. Lett. 25, 214220. doi:10.1007/ s00367-005-0209-2. Pethick, J.S., 1981. Long-term accretion rates on tidal salt marshes. J. Sediment. Petrol. 51 (2), 571–577. Pethick, J.S., 1992. Salt marsh geomorphology. In: Allen, J.R.L., Pye, K. (Eds.), Salt Marshes, Morphodynamics, Conservation and Engineering Significance.Cambridge University Press, Cambridge, pp. 41–62. Pignatti, S., 1966. La vegetazione alofila della laguna veneta, Istituto Veneto di Scienze, Lettere ed Arti, Memorie, vol. XXXIII – Fascicolo I, Venezia, 174pp. (in Italian). Pritchard, D., Hogg, A.J., 2003. Cross-shore sediment transport and the equilibrium morphology of mudflats under tidal currents. J. Geophys. Res. 108 (C10), 3313. doi: 10.1029/2002JC001570. Resio, D.T., 1987. Shallow-water waves. I: Theory. J. Waterway Port Coast. Eng. 113, 264–281. Resio, D.T., Pierre, W., 1989. Implication of an f-4 equilibrium range for wind generated waves. J. Phys. Oceanogr. 19, 193–204. Righetti, M., Armanini, A., 2002. Flow resistance in open channel flows with sparsely distributed bushes. J. Hydrol. 269, 55–64. Rinaldo, A., Fagherazzi, S., Lanzoni, S., Marani, M., Dietrich, W.E., 1999a. Tidal networks, 2, Watershed delineation and comparative network morphology. Water Resour. Res. 35 (12), 3905–3917. Rinaldo, A., Fagherazzi, S., Lanzoni, S., Marani, M., Dietrich, W.E., 1999b. Tidal networks, 3, Landscape-forming discharges and studies in empirical geomorphic relationships. Water Resour. Res. 35 (12), 3919–3929. Ris, R.C., Holthuijsen, L.H., Booij, N., 1999. A third-generation wave model for coastal regions. 2. Verification. J. Geophys. Res. 104(C4), 7667–7681. Schwab, D.J., Benett, J.R., Liu, P.C., Donelan, M.A., 1984. Application of a simple numerical wave prediction model to Lake Erie. J. Geophys. Res. 89, 3586–3592. Shi, Z., Pethick, J., Pye, K., 1995. Flow structure in and above the various heights of a salt marsh canopy: a laboratory flume study. J. Coast. Res. 11, 1204–1209. Shimizu, Y., Tsujimoto, T., 1994. Numerical analysis of turbulent open-channel flow over a vegetation layer using a –" turbulence model. J. Hydrosci. Hydraulic Eng. 11 (2), 57–67. Shoelhamer, D.H., (1995).Sediment resuspension mechanisms in Old Tampa Bay, Florida. Estuar. Coast. Shelf Sci. 40, 603–620. Silvestri, S., Defina, A., Marani, M., 2005. Tidal regime, salinity and salt-marsh plant zonation, Estuar. Coast. Shelf Sci. 62, 119–130.
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Silvestri, S., Marani, M., 2004. Salt marsh vegetation and morphology, modelling and remote sensing observations. In: Fagherazzi, S., Marani, M., Blum, L. (Eds.), Ecogeomorphology of Tidal Marshes. Coastal and Estuarine Monograph Series, vol. 59.American Geophysical Union, Washington, DC, 266pp. Silvestri, S., Marani, M., Rinaldo, A., Marani, A., 2000. Vegetazione alofila e morfologia lagunare. Atti dell’Istituto Veneto di Scienze, Lettere ed Arti, Tomo CLVIII (1999–2000), Classe di scienze fisiche, matematiche e naturali, pp. 333–359 (in Italian). Soulsby, R.L., 1997. Dynamics of Marine Sands. A Manual for Practical Applications.Thomas Telford, The Netherlands, 248pp. Swales, A., MacDonald, I.T., Green, M.O., 2004. Influence of wave and sediment dynamics on cordgrass (Spartina anglica) growth and sediment accumulation on an exposed intertidal flat. Estuaries 27 (2), 225–243. Temmerman, S., Govers, G., Meire, P., Wartel, S., 2004b. Simulating the long-term development of levee-basin topography on tidal marshes. Geomorphology 63 (1–2), 39–55. Temmerman, S., Govers, G., Wartel, S., Meire, P., 2004a. Modelling estuarine variations in tidal marsh sedimentation: response to changing sea level and suspended sediment concentrations, Mar. Geol. 212 (1–4), 1–19. Umgiesser, G., Sclavo, M., Carniel, S., Bergamasco, A., 2004. Exploring the bottom stress variability in the Venice Lagoon. J. Mar. Sys. 51, 161–178. Uncles, R.J., 2002. Estuarine physical processes research: some recent studies and progress. Estuar. Coast. Shelf Sci. 55, 829–856. doi:10.1006/ecss.2002.1032. van Ledden, M., 2003. Sand-mud segregation in estuaries and tidal basins. PhD Thesis, T.U. Delft, Department of Civil Engineering and Geosciences, Report 03-2, ISSN 0169-6548. van Ledden, M., Wang, Z.B., Winterwerp, H., De Vriend, H., 2004. Sand-mud morphodynamics in a short tidal basin. Oceans Dyn. 54, 385–391. doi 10.1007/s10236-003-0050-y. van Rijn, L.C., 1993. Principles of Sediment Transport in Rivers, Estuaries and Coastal Seas. Aqua Publications, Amsterdam, the Netherlands. Ward, L.G., Kemp, W.M., Boynton, W.R., 1984. The influence of waves and seagrass communities on suspended particulates in an estuarine embayment. Mar. Geol. 59, 85–103. Wayne, C.J., 1976. The effects of sea and marsh grass on wave energy. Coast. Res. Notes 14, 6–8.
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P A R T
I I I
TIDAL FLATS
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C H A P T E R
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G EOMORPHOLOGY AND S EDIMENTOLOGY OF T IDAL F LATS Shu Gao
Contents 1. Introduction 2. Basic Conditions for the Formation of Tidal Flats 3. Zonation in Sedimentation and Flat Surface Morphology 3.1. Vertical sediment sequences 3.2. Sediment and morphology on intertidal mud flats 3.3. Sediment and morphology on mixed sand–mud flats 3.4. Sediment and morphology on sand flats 4. Factors and Processes 4.1. Influences of quantity and composition of sediment supply 4.2. Sedimentation during tidal cycles 4.3. Long-term accretion–erosion cycles 4.4. Tidal creek systems 5. Summary Acknowledgments References
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1. INTRODUCTION Tidal flats are distributed widely along the world coastlines, representing an important part of coastal wetlands. A tidal flat can be divided into three parts, according to their relation to the characteristic tidal water levels (Amos, 1995): (1) supratidal zone, which is located above the high water on springs and is inundated only under extreme conditions (e.g., storm surge events); (2) intertidal zone, located between the high water and low water on springs and is inundated periodically during spring–neap tidal cycles, and (3) subtidal zone, which is below the low water on springs and is rarely exposed in air. In literature the studies on tidal flats have been concentrated mainly on the intertidal part; some authors use the term “tidal flat” to represent the intertidal zone (which is adopted in the present study), whilst others prefer the term “intertidal flat.” This chapter will concentrate mainly on the physical aspects of the intertidal flat; salt marshes will Coastal Wetlands: An Integrated Ecosystem Approach
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be described only briefly when necessary, because they are treated in detail in a separate chapter. The well studied tidal flats include those along the Dutch, German, and Danish coasts (Reineck, 1972; Reineck and Singh, 1980; Pejrup, 1988), in the Wash embayment of England (Evans and Collins, 1975; Collins et al., 1981), over the bayhead areas of the Bay of Fundy, eastern Canada (Amos and Mosher, 1985) and along the Jiangsu coastlines in eastern China (Wang, 1983; Ren, 1986) (Figure 1). Generally, these tidal flat systems are characterized by accumulation of fine-grained sediments and gentle bed slopes. Tidal currents are strong on the tidal flat, resulting in high mobility of bed materials. Over the upper parts of the intertidal zone, salt marshes may be present, with water and nutrients being supplied by the tides (Zhang et al., 2004). In tropical areas, mangroves may develop on mudflats (Wells and Coleman, 1981). Progress has been made in the understanding of the characteristics, processes, and evolution of tidal flats, which is important for the purpose of coastal wetland protection and restoration. In early times, the unique morphological features, especially zonation in geomorphology, attracted the researchers. From high water to low water marks, there are systematic changes in sediment grain size, bedforms, sedimentary structures, and biological activities, which have been studied since the 1930s (e.g. Haentzschel, 1939; Linke, 1939). Then, from the 1950s, sediment dynamic and morphodynamic studies have been carried out, in an attempt to understand the mechanisms of sediment transport and accumulation (Postma, 1954; van Straaten and Kuenen, 1957, 1958). Extensive in situ measurements
Figure 1 Coastal sections (dashed lines) and locations (triangles) along the world coastlines where extensive or detailed studies on tidal flat sedimentology and geomorphology have been undertaken (Allen, 1965; Reineck and Singh, 1980; Klein, 1985; Ren, 1986; Isla et al., 1991; Daborn et al., 1993; Perillo et al., 1996; Netto and Lana, 1997; Perillo and Piccolo, 1999; Kjerfve et al., 2002; Lim and Park, 2003; Deloffre et al., 2005; Falca‹o et al., 2006; Quaresma et al., 2007; Sakamaki and Nishimura, 2007; Anthony et al., 2008; Proske et al., 2008; Talke and Stacey, 2008).
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have been undertaken, on the flat surface and in tidal creeks, to obtain information on tidal current, suspended sediment concentration, and the benthic boundary layer (Evans and Collins, 1975; Letzsch and Frey, 1980; Collins et al., 1981, 1998; Stumpf, 1983; Bartholdy and Madsen, 1985; Pejrup, 1988; Wells et al., 1990; Alexander et al., 1991; Gouleau et al., 2000; Andersen and Pejrup, 2001; Davidson-Arnott et al., 2002). On such a basis, quantitative models for tidal flat sedimentation and morphodynamics have been proposed (Allen, 1989, 2000, 2003; French, 1993; Allen and Duffy, 1998; Roberts et al., 2000; Pritchard et al., 2002; Malvarez et al., 2004; Temmerman et al., 2004). At the same time, studies on sedimentary sequences and associated physically and biologically induced sedimentary structures were documented in detail, to obtain information on the environmental conditions under which the deposits were formed and on the environmental changes. Recently, research has been focused on the formation of tidal flat sediment systems and the information on climate, environmental, and ecosystem changes contained in the sedimentary record (Cundy and Croudace, 1995; Dellwig et al., 2000; Gerdes et al., 2003; Gao, 2007a). Furthermore, the evolution of tidal flats in response to global climate change and intensified anthropogenic activities has become an important research area (Vos and van Kesteren, 2000). The purpose of the present contribution is to provide a general description about the sediment distribution patterns and morphological features of tidal flats, together with an overview of the conditions and physical processes for the formation and evolution of tidal flats.
2. BASIC C ONDITIONS FOR THE FORMATION OF TIDAL FLATS Tidal flats are formed in areas where there is an important supply of finegrained sediment (i.e., clays, silts, and fine to very fine sands), and that tides and tidal currents dominate over other hydrodynamic forces (Klein, 1985). The first condition is satisfied for most coastal environments: fine-grained sediment is delivered by rivers and discharged into estuaries and adjacent coastal areas, erosion on the seabed and cliff recession provide additional sources of sedimentary materials, and organisms living in coastal waters and salt marshes produce shell debris and particulate organic matter. The second condition determines whether or not fine-grained sediment will be deposited on the tidal flat. Several factors influence this condition. Firstly, the tidal action should be significant. The average tidal range (R) has been used in the classification of the coast, that is, the coast can be microtidal (R < 2 m), mesotidal (R = 2–4 m), or macrotidal (R > 4 m) (Davies, 1964). On a microtidal coast, tidal currents are relatively weak unless the slope of the seabed is very small. In mesotidal and macrotidal environments, tidal currents tend to be relatively large compared with microtidal coasts, which favors the formation of tidal flats. It should be noted that tidal flats can be formed in microtidal areas of sheltered coastal embayments or semienclosed seas; this is because in such environments the wave action is of only a secondary importance compared with tidal currents.
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Secondly, the dominance of tidal action should be understood in a relative sense: tidal flats cannot be formed where wave action dominates, even if the tidal range is large. On open coasts, waves may represent a dominant force, which break in the surf zone, causing transport of fine-grained sediment toward offshore areas (King, 1972). In this case, if the supply of fine-grained materials is not abundant, then sandy or gravelly beaches will be formed, rather than tidal flats. However, if the rate of fine-grained sediment supply is high, then the accumulation of finegrained materials reduces the bed slope in the intertidal area. Eventually, wave breaking can rarely occur on the bed, that is, the wave energy is dissipated over the wide intertidal area due to bed friction, without causing breaking. The reduction of the bed slope will lead to enhanced tidal currents. Thus, in response to fine-grained sediment accumulation, wave action is weakened and the tidal action is enhanced. However, this observation does not imply that waves are unimportant on tidal flats; waves are important in the transport of sediment and the shaping of the tidal flat morphology. In situ measurements have shown that combined wave–tide action (without wave breaking) can cause intense sediment movement on the flat (Fan et al., 2006; Wang et al., 2006). Finally, while tidal action is a dominant agent, storm events can significantly modify the tidal flat environment. For example, during a typhoon event, storm surges become temporally the dominant forcing for sediment erosion, transport, and accumulation on a tidal flat (Ren et al., 1985; Ren, 1986; Andersen and Pejrup, 2001). The arguments outlined above imply that tidal flats may develop in sheltered tidal estuaries and coastal embayments where there is continuous supply of finedgrained sediment (although the rate of supply may be small), or on open coasts where tidal range is sufficiently large and sediment supply is abundant. A classical example of the former is the tidal flats in the Dutch Wadden Sea (van Straaten and Kuenen, 1957, 1958). Here, the tidal flat areas are sheltered by a series of barrier islands, the exchange of water between the Wadden Sea and the open North Sea being via tidal inlets cutting through the barrier islands. The sediment source is provided mainly by the North Sea (Postma, 1961; Pejrup et al., 1997), while river input represents a secondary source. For the open coast tidal flats, a typical example is the tidal flats on the Jiangsu coast, eastern China (Ren, 1986). During the Holocene period, because of the abundant sediment supply from two large rivers (i.e., the Changjiang and the Yellow rivers), extensive tidal flats have been formed.
3. ZONATION IN S EDIMENTATION AND FLAT SURFACE M ORPHOLOGY 3.1. Vertical sediment sequences Sediment cores taken from the upper part of the intertidal flat tend to show a “fining upward” sequence (Klein, 1985). Although the sediment deposited is ultimately determined by the source characteristics, in most cases the sediment
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source for tidal flats contain materials from sand- to clay-sized materials. Sandy materials tend to accumulate in the lower part of the flat, whilst muddy materials deposit over the upper part of the flat. The top part consists of clayey or muddy materials, often with high organic carbon content. Below this layer is a mud layer corresponding to the elevation near the high water, with very thin laminae as a major type of sedimentary structure. Then there is a mixed sand–mud layer, interbedded with sandy and muddy materials; such sedimentary structures are known as “tidal bedding” (Reineck and Singh, 1980). The lowest part of the core is a sand layer, corresponding to the lower part of the intertidal zone and the subtidal zone. Such a vertical sequence reflects the spatial distributions of sediments on tidal flats. A general pattern is that along a transect from the supratidal zone to subtidal zone there are salt marshes, mud flats, mixed sand–mud flats, and sand flats (Figure 2a). On the salt marshes, the bed material is the finest, with organic matter being derived from marsh plants and organisms. The mud flat is covered with clay and fine silts; this area is often located between the high water levels on springs and on neaps. In some regions, salt marsh vegetation may extend from the supratidal zone into the mud flat. Mixed sand–mud flats are close to the mean sea level; here, sands are deposited during spring tides and muds are deposited on neaps (for details, see Section 4.2). Over the lower parts of tidal flats, well sorted sands are present, with various types of bedforms (e.g., dunes and ripples). On the tidal flats tidal creeks may be formed, consisting of small creeks and large tidal channels. Sedimentation in tidal creeks may be different from that on the adjacent flat surface, (a) A
B
C
D HWS
A B C D
Salt marsh Mud flats Mixed flats Sand flats
LWS
(b) A
B
C
D
E
F HWS HWN
A B C D E F
Salt marsh Higher mud flats Inner sand flats Arenicola sand flats Lower mud flats Lower sand flats
LWN LWS
Figure 2 General patterns of (a) zonation of tidal flats (Reineck and Singh, 1980) and (b) the geomorphic and sedimentary features of the Wash, England (Evans, 1965). HWS, high water on springs; HWN, high water on neaps; LWN, low water on neaps; LWS, low water on springs.
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but the general “fining-upward” pattern is also maintained for the creek systems. In some tidal flat environments, the zonation can have slightly different patterns, as indicated by Figure 2b. The reason for this is that, in addition to bed elevation, other factors such as bed slope and tidal water level also influence the transport and accumulation of sediments (see below). The zonation of tidal flats implies that the finest materials tend to accumulate as the upper part of the sedimentary sequence. This raises a question of the thickness of the mud layer on the upper part of the tidal flat. Data sets from the different parts of the world (Table 1) show that the thickness varies considerably: in some places the entire profile is covered with muddy sediments, whilst on the other extreme mud is absent. Apparently, the correlation between the thickness and tidal range is poor (cf. data listed in Table 1). In Section 4, an explanation about the factors that control the thickness of the mud layer will be given on the basis of sediment dynamic analysis, for the tidal flats that are associated with a supply of different types of sediments ranging from clays to fine sands.
3.2. Sediment and morphology on intertidal mud flats Generally, mud flats are located on the upper part of the intertidal zone, with very gentle bed slopes (Figure 3). Here, the sediment is the finest for the entire tidal flat, and deposition takes places because of the settling of fine-grained materials from the water column. Therefore, although there are tidal cycle changes in grain size, mud (i.e., a mixture of clayey and silty sediments) is the major component of mudflat sediments. The deposition of coarser materials, which occasionally occur in the mudflat sediment layers, is due to extreme events (e.g., storm surges). The sedimentary record consists of laminated mud, with alternating silty and clayey layers of less than 1 mm in thickness. Because resuspension of the bed material is relatively weak here, the continuity of the mud deposit is high compared with the other parts of the tidal flat. However, in some places intense bioturbation often causes destruction of the original sedimentary structure. The landward part of the mudflat may be covered by pioneer plants extending from the salt marshes in the supratidal zone, or even covered with salt marshes. On the Jiangsu coast, eastern China, for example, extensive Spartina marshes are formed over the landward part of the mud flats, with an upper limit close to mean high water on springs and a lower limit slightly lower than high water on neaps (Zhang et al., 2004). Under tidal action alone, accretion becomes progressively slower when the bed elevation is enhanced due to sediment accumulation (Pethick, 1981); thus, the high water level on springs represents the upper limit for the accretion (Amos and Mosher, 1985; Amos, 1995). However, the bed accretion continues beyond the high water level, to form the supratidal zone. There are several processes for this phenomenon. First, during storm events the water level can become much higher than during normal tidal cycles, especially when the surge coincides with an astronomical spring tide. During such events, a large amount of sediment may be transported to the upper parts of the flat and deposits
Location
Tide range (m)
Wadden Sea (The Netherlands) The Wash (England) Colorado River Estuary (USA) Wadden Sea (Germany) Fraser River Estuary, Boundary Bay (Canada) Central Jiangsu coast (China) Salmon River Estuary, Bay of Fundy (Canada) Haenam Bay (South Korea) Newtownards (Northern Ireland) Baeksu Tidal Flat (South Korea)
1.3–2.8 5.0 4–5 2.7 11.9 3.0 3.0 3.9
RS
RN
6.5 6–8 2.6–4.1 4.1
3.5
15.2 4.0 3.5
1.8–3.1 1.5 8.7 1.8
Hm (m)
W (km)
Reference
2.0 1.5–2.0 >8 2.5 <0.5 3 4 >4 0 <0.5
7–10 1.0–6.5
van Straaten (1961) Evans (1965) Ginsberg (1975) Ginsberg (1975) Ginsberg (1975) Ren (1986) Dalrymple et al. (1990) Lim and Park (2003) Malvarez et al. (2004) Yang et al. (2005)
5 7–10 5 2–2.5 0.5–1.2 4–6
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Table 1 Thickness of the sediment layer above the mixed sand–mud flat ( = average tide range, RS = spring tide range, RN = neap tide range, Hm = thickness of mud layer, W = width of intertidal zone)
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Bed features of the bare mud flat, Jiangsu coast, eastern China.
there. Thick storm deposits have been identified in the upper tidal flat sequences (Ren et al., 1985). Second, animals living in the flat environment modify the bed morphology through bioturbation, generating unique sedimentary structures (Reineck and Singh, 1980). Biological activities produce uneven bed morphology (e.g., mud mounds formed by crabs and mud skippers): the elevated features may become above the high water, and the depressions will receive an increased amount of sediment during high water. Finally, the deposition of aerosol forms an additional sediment source (Li et al., 1997), which is not constrained by the high water mark. Tidal creeks are present on the mudflat. On the bare mud flat, the tidal creeks have a relatively small depth to width ratio, with lateral migrations. At the boundary between the mudflat and salt marshes, cliffs or scarps tend to occur, with a height of less than 1 m (Reineck and Singh, 1980). In some places much higher scarps can be formed, for example, those found in the Severn Estuary, UK, which has been interpreted as periodic erosion and accretion cycles in response to changes in hydrodynamic and sediment supply conditions (Allen, 1989). However, disagreement exists with regard to the significance of the low cliffs that are more widely distributed. Some researchers believe that the cliffs represent an indication of coastal erosion, but others argue that they result from localized scour. Observations show that because the accretion on the marsh that is more rapid than the adjacent bare flat (the flat with plants traps more sediment than the bare flat, and the organic matter further adds to the sedimentary materials in the marsh), the bed elevation gradually becomes different (Amos, 1995). Then, the edge of the marsh becomes progressively steeper, causing concentration of wave energy (the small waves would otherwise not break at this location).
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In coastal embayments, such low cliffs are often associated with low hydrodynamic forcing and low sediment supply. For example, in Christchurch Harbour, southern England, where the deposition rate has been higher on the marsh than on the bare flat. Subsequently, the small waves formed in the estuarine waters (with a fetch of only 1 km) started to erode part of the materials at the marsh edge, forming low cliffs of 0.4–0.6 m in height (Gao and Collins, 1997). Numerical model output (Gao and Collins, 1997) indicated that the action of small waves, which break at the marsh–bare flat boundary because of the enhanced bed slope at this location, is responsible for the localized scour and the cliffs can be stable or retreat slowly. In the German and Danish Wadden Sea, marsh cliff erosion was measured by Pejrup et al. (1997) and similar results were obtained. Elsewhere, measurements in Rehoboth Bay, USA, revealed that the rate of cliff recession is also small, ranging between 0.14 and 0.43 m/year (Schwimmer, 2001). Hence, it is the evolution of the tidal flat itself that creates the condition for the formation of the cliff.
3.3. Sediment and morphology on mixed sand–mud flats Mixed sand–mud flats are characterized by alternating deposition of muddy materials on neaps and sandy material on springs. The spatial distribution of this morphological unit on the tidal flat depends upon a number of factors, including tidal regime, sediment supply, and suspended sediment concentration of seawater. On the Jiangsu coast, where the tidal currents during the flood and ebb maximum periods exceed the threshold for bedload transport over the entire spring–neap tidal cycle, the mixed flat is located between high water on neaps and mean sea level (Zhu and Xu, 1982). In the sediment sequences, typical “tidal bedding” (i.e., interlayered relatively coarse- and fine-grained sediments), is found on the mixed sand–mud flats. These beddings contain information on the sedimentary processes during flood–ebb cycles and spring–neap cycles (Dalrymple et al., 1990). A well-preserved sediment record at this location will show that the thickness of the sand layer decreases with decreasing tidal range from the springs toward the neaps. Likewise, the mud layer increases its thickness toward the neap tides. However, the preservation potential for the sand–mud flat is usually not high, and most of the sedimentary record formed in tidal cycles is destroyed subsequently by reworking or resuspension of the bed materials (Fan et al., 2002; Deloffre et al., 2005, 2007; Gao, 2007a). On the bed, scour features are present, resulting from sediment reworking (Figure 4a,b). Generally, on the lower part of bare mud flats and mixed mud–sand flats, a part of the muddy material accumulated on neaps can survive until the next neap tidal phase. This is the reason for the formation of the scour features. It should be noted that such scour does not indicate long-term net erosion of the bed; on the contrary, the accretion rate on the mixed sand–mud flat is actually higher than the other parts of the tidal flat system (Gao and Zhu, 1988). The scour represents only short-term effects within a long-term accretion trend.
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(a)
(b)
2003 6 29
(d)
(c)
2003
7
9
2003 6 29
(e)
Figure 4 Geomorphic and sedimentary features of the middle to lower parts of the intertidal zone on the Jiangsu coast, eastern China: (a) scour features on mud flats; (b) scour features on mixed mud ^ sand flats; (c) a tidal creek on mixed sand ^ mud flats; (d) sand flat bed, with current ripples; and (e) sand flat near the low water mark.
Tidal creeks are highly dynamic on mixed sand–mud flats. For instance, on the central coast of Jiangsu Province, China, the strong currents in combination with the silty sediments on such flats result in rapid lateral migration of the tidal channels. Channel migration generates new tidal creeks, which can extend toward the upper part of the intertidal zone rapidly, in the form of headward erosion. During storms, the tidal creeks are even more active; significant deepening of the channel bottom, development of many new channels, and rapid migration may be observed (Ren, 1986).
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3.4. Sediment and morphology on sand flats In a typical tidal flat system with abundant sandy sediment supply and strong tidal currents, sand flats occupy the lower part of the intertidal zone. The sands are normally well sorted. Tidal creeks on the sand flat are relatively wide, with a small depth to width ratio (Figure 4c). The migration of the tidal channels is active, and is influenced by bedload transport and accumulation. Lateral migration can be caused by bedload movement in the longshore direction (i.e., transport parallel to the shoreline). Cross bedding is a common type of sedimentary structure in tidal creek sequences. During the late stages of the ebb, before the bed is exposed to air, the tidal currents are too strong for the fine-grained, suspended sediment to settle onto the bed, and the material settled during the flood slack tends to be suspended during the ebb. Numerical calculations show that net accumulation of mud on the sand flat is possible only when the suspended sediment concentration is extremely high (Amos, 1995), in which case sand flats with pure sand deposition cannot be formed. Where the bed slope is sufficiently large (1.0 10–3 or greater) to allow rapid draining of water mass, the geometry of bedforms (dunes and ripples) is preserved and exposed during low tide (Klein, 1985; Dalrymple et al., 1990). The dunes have a wavelength (or spacing) of 0.6–6 m, whilst the ripples are small bedforms (wavelength <0.6 m). Compared with the dunes, ripples are more extensively distributed. Field observations show that during the high water slack ripples are formed by combined wave–current action; during the ebb the upper part of the sand flat will be exposed rapidly and the morphology of wave ripples may be preserved, but the lower sand flat can be modified by ebb currents to form ebb-oriented, flowgenerated ripples (Amos and Collins, 1978). If the bed has an extremely gentle slope, then these bedforms may be modified by water flow generated by seepage from the substrate; because the flow depth is very small (i.e., <10–1 m) and the Froude number is large (i.e., close to 1), plane bed without bedforms can be formed in a short period of time (i.e., 1–2 h). On the Jiangsu coast, eastern China, the central parts of the sand flat (with a mean grain size of 0.06–0.1 mm) have a bed slope of around 0.0005, and the slope increases to more than 0.001 toward either the low water mark or the boundary between the sand flat and the mixed sand–mud flat (Zhu and Xu, 1982). Therefore, during low water on springs, wave ripples are present near the boundary between the sand flat and the mixed sand–mud flat, plane beds are present over the central part of the sand flat, and current ripples are found over the lower sand flat (Figure 4d,e).
4. FACTORS AND PROCESSES 4.1. Influences of quantity and composition of sediment supply The evolution of present-day tidal flats began when sea level reached its highest elevation during the Holocene period, from a base that was left from the last glacial
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period. Thus, the accommodation space for the tidal flat deposit is confined by the high water mark, the original topography/bathymetry and the tidal flat profile. While the high water level and the topographic baseline are influenced by sea-level changes and crustal movements, the rate of shoreline advancement is determined by the amount of sediment supply, and the shape of the profile is influenced by the grain size composition of the materials supplied. As shown in a simplified shore-normal transection for the geometry of a tidal flat sediment system (Figure 5), the increase of the tidal flat sediment body through time is associated with upward accretion and the advancement of the tidal flat profile toward the sea. The vertical accretion rate of a tidal flat can be related to the portion of sediment supply that contributes to the tidal flat formation, by: D=
DS sinð Þ tan DL sin sin
ð1Þ
where D is the vertical accretion rate averaged over the entire profile, DS is the proportion of sediment supply that is deposited on the flat per unit time over unit length of the shoreline, is the slope angle of the original topography, is the average bed slope angle of the flat profile, and DL is the distance of shoreline advancement during a unit period of time. In addition to the accretion rate, the thickness of the mud layer of the upper part of the sequences is also an important parameter for the tidal flat. Here, the mud layer represents the deposits of the mud flat and the mixed sand–mud flat (see Sections 3.2 and 3.3). Assuming that most of the muddy materials are deposited on the upper part of the tidal flat, as is the case for the Jiangsu coast, this thickness can be expressed approximately as (cf. the geometric relationship shown in Figure 5): Hm »
DS1 ðL þ DLÞtan DS
ð2Þ
ΔL
L
α
β
ΔS1
HWS
S LWS Original Coastal profile
ΔS
Figure 5 Schematic diagram showing a shore-normal cross section of tidal flat sedimentation (S = volume of sediment accumulated over unit length of the shoreline, L = the width of coastal plain formed by tidal flat sedimentation, DS = the amount of sediment supply per unit time over unit length of the shoreline, DL = the distance of shoreline advancement during a unit time, = the slope angle of the original topography, and = the average bed slope angle of the flat profile).
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where Hm is the thickness of the mud layer, DS1 is the mud fractions within DS, and L is the total width of coastal plain formed by tidal flat sedimentation. Equation (2) implies that the thickness is related to the composition of sediment supply and the stage of the tidal flat evolution (since it is a function of DS1 and L), but it is not directly related to tide range. Observations of the world tidal flats have provided support to this inference (cf. Table 1): the thickness of the mud layer does not increase with the increasing tide range. In an extreme case (Malvarez et al., 2004), the profile is covered entirely with mud, because in this system the supply of sandy sediments is very small. On the Jiangsu coast, the sediment input from the Yellow River has relatively stable composition; as a result, there is a trend of increase in the thickness of the mud layer during the tidal flat development (Gao, 2007a). Such an observation is also consistent with the prediction by Equation (2). The thickness of the mud layer determines the boundary on the tidal flat where sandy and muddy deposits are divided. At this boundary, the bed slope can be expressed as a function of the threshold for the initial motion of the sands, as demonstrated below. At any site over the intertidal flat, the instantaneous tidal current velocity is a vector, which consists of an onshore–offshore component, u, and a longshore component, v (Anderson, 1973; Perillo et al., 1993; Wang et al., 1999). The onshore–offshore component is controlled by the water level changes and the bed morphology, on the basis of the principle of mass conservation (Zhu and Gao, 1985; Wang et al., 1999): u=
1 dh tan dt
ð3Þ
where tan is the average bed slope over the inundated section and h is the tidal water level. At the sand–mud boundary, maximum current speeds during the flood should not exceed the threshold for initial bedload motion; otherwise, sandy materials will be transported across the boundary further toward the upper tidal flat. At the same time, it should not be smaller than the threshold; otherwise, sandy materials cannot be transported to this location. Hence, there is only one possibility: the sand–mud boundary is located where the maximum shore-normal current speed is equal to the threshold: umax = ucr sin =
1 dhb tan b dt
ð4Þ
where is the angle between the current direction and the longshore direction, the subscript b denotes the sand–mud boundary, and dhb/dt denotes the rate of water level change when the flow reaches the boundary. Thus, the slope at the sand–mud boundary can be defined by tan b =
1 dhb ucr sin dt
ð5Þ
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The threshold current speed for initial sediment motion is a function of near-bed shear stress. Equation (5) implies that a steep bed slope at the sand–mud boundary is associated with a large value of dhb/dt, or a low elevation for the sand–mud boundary (since the rate of water level change cannot be large near the high water mark). Because the thickness of the mud layer is the vertical difference between the elevation at the boundary and the high water mark, the steep slope also means a lager thickness of the mud deposit in the tidal flat sequence. This observation may explain the phenomenon that on a tidal flat with a thick mud layer maximum slope gradients occur on the mixed sand–mud flat (Zhu and Xu, 1982). The elevation associated with the critical current speed differs between the spring and neap conditions. As a result, there will be two critical elevation values, one for neap tides and the other for spring tides. Between the two critical elevations, mud is deposited on neaps and sand is deposited on springs. Therefore, the thickness of the mud layer (consisting of mud and mixed mud–sand deposits) is related to the lower critical elevation for the neap tides, and the mixed sand–mud flat itself is a result of spring–neap tidal cycles.
4.2. Sedimentation during tidal cycles On a tidal flat, why are the sediments transported toward the land? Two mechanisms have been identified, one for suspended load and the other for bedload. The movement of the suspended load is related to a physical mechanism known as “settling and scour lag effects” (Postma, 1954; van Straaten and Kuenen, 1957, 1958). The basic condition for these effects is the particular patterns of current speed variations over the tidal flat. During a tidal cycle, the rate of water level changes and, according to Equation (3), the current speed also changes. Minimum rates of water level change occur during high and low water periods, and maximum rates appear at the middle of flood or ebb phases. Consequently, only the middle and lower parts of the intertidal zone will experience large tidal currents, and the upper part is associated with weak currents. In such an environment, it will be difficult for the suspended material originated from subtidal areas to permanently stay on the middle and lower parts of the tidal flat, unless the concentration is sufficiently high (Amos, 1995). The “settling and scour lag effects” explain why a suspended particle is carried by currents toward the upper tidal flat, as shown in Figure 6. On the tidal flat, P represents the location above which the tidal current speed never exceeds the critical value for resuspension, and below which the speed is below the critical value only during the slack periods. Let us suppose that a water particle and a suspended sediment are located at the same site, A, at the beginning of a flood tide (Figure 6a). During the flood, when the current speed decreases to the threshold at location B (at this time the water level reaches the elevation of site P), the sediment particle starts to settle, but it cannot reach the bed at location B (Figure 6b). It may reach the bed at a site, C, which is further toward the high water mark (Figure 6c). Such an effect is called “settling lag.” During the ebb, when the water particle reaches the location C, its speed is still below the threshold for resuspension, because at this time the water is above site P (Figure 6d). Resuspension for the sediment in consideration does not occur until the water particle arrives at B
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(a)
Intertidal zone
HW P C
B
LW A
(b) HW P C
B
LW A
(c) HW P C
B
LW A
(d) HW P C
B
LW A
(e) HW P C
B
LW A
(f) HW P C Water particle Sediment particle
B
LW A
Figure 6 Diagram showing the settling and scour effects, by comparing the tidal cycle movement of a water parcel and a sediment particle at the stages of (a) low water (representing the beginning of the tidal cycle); (b) water level reaching the point P when the current speed is reduced to the critical value for bed erosion; (c) high water; (d) the water parcel reaching the site C where the sediment particle is settled to the bed; (e) water level reaching the point P; and (f ) low water (the end of the tidal cycle).
(Figure 6e). At the slack water of the ebb, the positions for the water parcel and sediment particle becomes different (Figure 6f). The process described in Figure 6d–f is known as “scour lag.” Thus, during each tidal cycle, the net transport is
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directed toward the high water mark. Although some biological and biogeochemical processes also influence the transport of suspended materials (Neumeier and Amos, 2006), the lag effects represent a basic physical mechanism. For the bedload, the time–velocity asymmetry patterns caused by the deformation of tidal waves favor landward transport. Over the tidal flat, shallow tides (known as “over tides”) are generated by seabed friction. As a result, the flood duration becomes shorter and the ebb duration becomes longer on the flat surface (Zhang, 1992). At the same time, the peak current speed during the flood becomes larger than during the ebb. In addition, the transport rate for bedload has a nonlinear relationship with the current speed (Hardisty, 1983; Wang and Gao, 2001). Thus, it has been shown mathematically that the transport capacity during the flood is greater than during the ebb (Zhu and Gao, 1985). This indicates that, if sufficient bedload is available, then net transport will be directed to landward during a tidal cycle. However, such a pattern may not be observed if the sites for measurements are close to tidal creek systems. The landward transport of bedload is important for the maintenance of accretion of tidal flats. Otherwise, the accumulation of fine-grained materials over the mudflat will lead to a narrower intertidal zone and reduce the average bed slope. This will, in turn, reduce the tidal current speed according to Equation (3). If the tidal dominance disappears, then development of tidal flats cannot continue. Thus, it is crucial for sands to be deposited over the lower parts of the tidal flat; only in this way can the small bed slope of tidal flats be maintained.
4.3. Long-term accretion–erosion cycles Sediment supply is a necessary condition for the growth of tidal flats. When the supply is cut off, or it becomes too small, erosion will occur. The coastline near the old Yellow River Delta, in northern Jiangsu Province, China, is a typical example which indicates the effect of sediment cut off. Before 1855, the sediment discharge of the Yellow River formed a large delta, and tidal flats were well developed. Then, in 1855 the Yellow River shifted its course to discharge into the Bohai Sea in northern China. Since then the shoreline in northern Jiangsu has been retreating. The original tidal flat has been modified in terms of sediment distribution and profile morphology. The width of the intertidal zone has been reduced to less than 2 km, sandy materials and mud pebbles are found near the high water mark. The overall profile shape is approaching to a wave-dominated beach profile (Gao and Zhu, 1988). Such responses can be explained by sediment dynamics. Because the supply is reduced, the landward transport capacity cannot be satisfied. However, at this time, the transport capacity during the ebb remains (i.e., the tidal flat surface is now transformed from a sediment sink to a sediment source). The removal of sediment from the lower part of the flat results in reduction in bed slope and in the strength of tidal currents on the flat. Eventually wave action becomes a dominant factor. Wave breaking takes place on the shore face, and the fine-grained material is transported toward deeper waters, just as observed on a sandy beach. During shoreline recession, if the sediment strata contain shells and other coarse-grained materials, then this debris may form cheniers (Augustinus, 1989; Wang and Ke, 1989). Thus, the
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presence of cheniers in tidal flats is indicative of coastal erosion periods. Often such cycles are related to sediment supply changes. Even if the sediment supply is maintained, there may still be a limit of growth for the tidal flat sedimentary system. The growth of a river delta may represent an analogue: in response to reduction in the Sediment Retention Index (SRI, the ratio of the sediment permanently retained in the system to the total sediment input provided by fluvial, marine, atmospheric, and sources) when the delta progradates toward deeper water areas, the rate of delta growth will decrease (Gao, 2007b). It is likely that similar processes are associated with the tidal flats, that is, they have a limited space for its development and after their growth over the Holocene period they may be already approaching the growth limit.
4.4. Tidal creek systems Tidal creeks have a different function compared with the flat surface. During the flood there exist several velocity maxima in a creek, in response to rapid enlargement of the inundated area on the flat adjacent to the creek, and during the ebb extra water enters the creek from the surface or through seepage from the bed (Bayliss-Smith et al., 1979; Wang et al., 1999). As a result, water balance in the creek can be asymmetric: the water discharge during the ebb is larger than during the flood. Observations demonstrate that net sediment transport in creeks tends to be seaward, especially during spring tides (Yang et al., 2003). This pattern may be further enhanced when storm occurs. Therefore, it can be inferred that tidal creeks reduce the overall accretion rate of the tidal flat. Each tidal creek system occupies a “drainage basin” on the tidal flat (Zhang, 1992). In macrotidal environments, tidal creeks may migrate laterally intensively, forming various sizes of point bars. On the sand flat, the creek channel migration is affected by bedload transport in the longshore direction, as observed on the Jiangsu coast (Wang et al., 2006). However, the migration of the creek is confined with the drainage basin.
5. SUMMARY The basic sedimentological and geomorphological characteristics of tidal flats may be summarized as follows. 1. Tidal flats are formed under the condition that tides dominate over other hydrodynamic processes. They have a significant pattern of zonation, with a general pattern that from the supratidal zone to subtidal zone salt marshes, mud flats, mixed sand–mud flats and sand flats are distributed. Such a pattern may be modified by other factors such as tidal creek formation and sediment supply. In a sediment core from an upper part of the tidal flat, the zonation is reflected by a “fining-upward” sequence. 2. Settling and sour effects are responsible for the transport and accumulation of muddy sediments over the upper parts of tidal flats, whilst the deformation of tidal waves causes landward transport of sandy materials, with the upper limit
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of sand accumulation being controlled by the tidal current speed during the flood. In addition to these physical mechanisms, biological effects on the suspended sediment transport are also important. 3. Some characteristics and/or parameters of tidal flats (e.g., the bed slope gradient, sedimentary structures, salt marsh cliffs, scour over the mixed sand–mud flat, the thickness of the mud layer, the tidal creeks, and the sediment retention index) contain important information on the system behavior and evolution of tidal flats. Such information may be obtained by means of sediment dynamic analysis.
ACKNOWLEDGMENTS This study has been supported by the SCOR-LOICZ-IAPSO Working Group 122. Financial support is also provided by the Natural Science Foundation of China (Grant Number 40476041), the Ministry of Science and Technology of China (Grant number 2006CB708410), and the Science-Technology Administration of Jiangsu Province (Grant number BK2005211). Mr. Niu Zhan-sheng is thanked for his help with the preparation of some of the figures, and Dr. Ya Ping Wang is thanked for providing Figure 4c–e. Dr. Gerardo M.E. Perillo provided the definition of sediment retention index. The author wishes to thank the reviewers (Professor Carl Amos, Dr. Morton Pejrup, and Dr. Gerardo M.E. Perillo) for their constructive comments on the original manuscript.
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I NTERTIDAL F LATS : E COSYSTEM F UNCTIONING OF S OFT S EDIMENT S YSTEMS David M. Paterson, Rebecca J. Aspden, and Kevin S. Black
Contents 1. Introduction 2. The Depositional Habitat 2.1. The physical background to life in depositional habitats 2.2. The functional difference between mud and sand systems 3. The Functional Role of Biota 3.1. Patterns of life 3.2. Effects of sediment disturbance 3.3. Biodiversity impacts 3.4. Distribution in space and time 3.5. Trophic structure 3.6. New functional groups? 4. Future Shock: Climate Change and Ecosystem Function Acknowledgments References
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1. INTRODUCTION The terminology of ecology has been steadily evolving over the years since Eugenius Warming arguably first established the discipline in 1895 (Goodland, 1975). Early ecological work was largely descriptive of species and habitats, often concerned with changes in plant communities over time. This work formed the classical study of the successional progression of assemblages toward a putative vegetative climax. Therefore, ecological concepts and terms suitable for the scales and temporal variation of terrestrial habitats were initially developed. The abiotic component of the habitat was considered to provide the overarching framework within which species might compete and be successful, leading to Clement’s original monoclimax theory of successional change (Townsend et al., 2008). The physical environment was therefore the stage upon which the biotic actors played their roles. This paradigm is now largely rejected as too simplistic since it is now widely recognized that the organisms inhabiting an ecosystem have a range of Coastal Wetlands: An Integrated Ecosystem Approach
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effects on the physical structure and dynamics of the system and contribute as architects of their own habitat (Hansell, 2005). This is particularly true of aquatic depositional systems where organisms burrow, restructure, and process the material of their surroundings in a process known as bioturbation (Reise, 2002). However, ecological paradigms are still generally being developed for terrestrial habitats, and there is a clear lag in the application of these theories to aquatic systems (Raffaelli et al., 2005). However, there is a new urgency in addressing coastal ecology (Duarte et al., 2008), which is being driven by the challenge of managing marine systems under scenarios of global climate change. One of the strongest intellectual drivers is to understand the processes that occur in an ecosystem that are beneficial, or even essential, to humans. These beneficial processes or functions have been described as ecosystem services (Chapin et al., 1997), and understanding them requires knowledge of how the biota provide to these critical services. The biotic component of the system is often reported as some measure of the variety of species that contribute to the process, under the general term of “biodiversity” (Magurran, 2004). Thus the question becomes “How does biodiversity affect ecosystem function?” The answer to this question would provide a clear scientific and political message concerning the value, economy, and health of marine ecosystems, and while coastal systems have latterly received less attention (Duarte et al., 2008), depositional environments have an advantage for addressing the above question that is currently being exploited by marine scientists. Sediment systems can be re-created in laboratory mesocosms (Benton et al., 2007; Emmerson et al., 2001), manipulated in the field, and their properties and processes relatively easily characterized. The time scale of change and even the size of the dominant organisms make this observation and manipulation more logistically amenable than for terrestrial systems. The biodiversity-ecosystem function debate is therefore a highly active area of research and is shedding further light on the dynamics and functional role of depositional systems (Solan et al., 2006). Intertidal depositional systems are often considered as physically challenging and stressful environments (Kaiser et al., 2005). Surging waves and tides impose a considerable physical challenge to the native organisms. The habitat is also highly variable in terms of physicochemical conditions such as temperature and salinity. It is this constant variation combined with the physical forces that organisms inhabiting depositional environments must tolerate to be successful. Fewer species occur in this harsh environment than found in more amenable systems (Kaiser et al., 2005), such as tropical rainforest. Under conditions where resources are available and the habitat is complex, the coexistence of species is promoted by a wide variety of niche space (e.g., kelp forest, coral reefs, and tropical rainforest). In contrast, where physical factors dominate in the context of a structurally simple habitat (e.g., intertidal flats), variation in available niche space is restricted and fewer species can compete for resources (Ricklefs and Miller, 1999). The struggle for survival under these circumstances may be against the elements and a limited number of well-adapted competitors. The open structure of depositional habitats and the life style of the burrowing organisms are considered to further reduce the influence of interference competition between species (Nybakken and Bertness, 2005).
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Sandy and muddy depositional ecosystems each have characteristic attributes that influence the metabolic processes and transformations that occur within them. In trying to conceptualize these varied ecological dynamics across systems, a new ecological terminology has been evolving and is now commonly applied. The transformations of matter or energy driven by biotic assemblages that occur within natural ecosystems are termed “ecosystem functions” while those functions that are, rather subjectively, deemed to be important to humans are termed “ecosystem services” (Chapin et al., 1997). The Millenium Ecosystem Assessment (2005) defines Ecosystem Services as “benefits people obtain from ecosystems.” These benefits may be derived directly or indirectly (Costanza et al., 1997) and can cover a wide variety of services which may include regulatory or supporting services (Beaumont et al., 2008). It is also now becoming more common to refer to “ecosystem health,” a subjective term which encompasses the implicit suggestion that a “healthy” system will deliver more appropriate levels of goods and services to humanity than an unhealthy one. The significance of this terminology essentially lies in the conceptual importance of the functional capacity of the species or groups. This has given impetus to the biodiversity-ecosystem function debate that is now gathering more empirical and theoretical evidence (Loreau et al., 2002). The central question addresses the importance of biodiversity for ecosystem function, or put another way “how many species does an ecosystem need to remain healthy?” This debate is now being advanced on many levels (Raffaelli, 2006; McCann, 2007). The logistic ease of experimental manipulation has meant that field and laboratory observations of soft sediment systems have made a significant contribution (Emmerson et al., 2001; Raffaelli, 2006; Solan et al., 2006). This chapter addresses some of these issues in the context of the depositional habitats and introduces modern approaches in terms of ecological dynamics to the discussion. These authors have provided their own working understanding of these terms in a simplified form (Table 1), but there are many other more subtle interpretations possible. For example, the term “ecosystem engineer” has broadened with time to include a wider range of organism than latterly acknowledged (Boogert et al., 2006). Further, reading across the field is advisable to fully understand the correct usage and lineage of these terms (Lawton, 1994; Loreau et al., 2002; Worm and Duffy, 2003).
2. T HE D EPOSITIONAL H ABITAT Depositional habitats can only occur where the correct boundary conditions of flow energy exist (Davis and FitzGerald, 2004). Too much energy and the sediments are swept away, too little and the system may stagnate. The extent of the physical forcing controls the type of depositional habitat formed. However, within these bounds, great opportunities exist for ecosystem engineers to moderate the habitat and increase their own fitness. Some extreme forms may even create a depositional habitat where none existed before. The red alga Audouinella sp. (Dillwyn) is a perennial rhodophyte characterized by its ability to retain sediment particles, which then become an important structural component of the algal turf.
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Table 1 Terminology commonly used in the ecosystem function debate Ecosystem process Ecosystem function
Ecosystem services
Ecosystem engineer
Ecosystem engineering Niche construction
Any process of transformation that occurs in an ecosystem, theoretically whether measurable or not. This includes all metabolism, catabolism, and dynamic processes such as sediment bioturbation or active resuspension Largely used as a synonym for “ecosystem process” but often given a more “practical” role: a process that can be measured as an attribute of the system under study. Habitat or its inherent three-dimensional structure (architecture) may itself be regarded by some authors as “functional” such as in the statement “The seagrass meadows provide refuge for juvenile fish” A very anthropocentric term in common usage. Ecosystem services are ecosystem functions that have implicit value to mankind. Common examples are carbon fixation, oxygen generation, and nutrient turnover. However, physical attributes can also be recognized as services such as the greater resilience of some habitats (e.g., mangroves) when faced with extreme conditions (tsunami, etc.) (Alongi, 2008) An organism whose activity has a significant impact on its habitat. It has varied definitions under a variety of contexts and is widely used. In fact, most organisms could be argued to act as ecosystem engineers on some scale, but a common example is the lugworm, Arenicola marina. A. marina has a great impact on sediment turnover as a bioturbator. However, even diatoms or bacteria can be regarded as ecosystem engineers since, in concert, they can produce organic material which can increase the erosion resistance of sediments (Paterson and Black, 1999) The activity of an ecosystem engineer Recently, a new evolutionary paradigm has been proposed, which suggests that organisms that engineer their environment create a selective pressure on their own future generations and those of other inhabitants of that environment. Thus, there is an evolutionary pressure inherent in ecosystem engineering. “Niche construction” is the name given to this form of evolutionary ecosystem influence (Laland et al., 2004)
Definitions are given with a perspective toward depositional habitats.
The ability to trap and retain sediments is an excellent example of “ecosystem engineering” due to the mediation of the surrounding environment in a manner that changes the nature of resource availability and therefore fitness. Audouinella sp. initially attaches to rocky substrata in the high energy zone of the intertidal, and once established it retains sand grains to form a “cushion” of sediment, often several centimeters thick over the surface of the rock (Figure 1). Experimental flume studies suggested the presence of the algae reduced bed load transport and sediment was trapped within the filamentous matrix of the algae, creating a new “depositional” environment (Aspden, 2005). Some authors even suggest that sediment stabilization by biological action may mediate the existence of alternate stable states (van de Koppel et al., 2001). This form of habitat mediation, biogenic stabilization, may be seen as an adaptation to sediment redistribution where high sedimentation rates and particle movement can have detrimental effects on the diversity and overall richness of intertidal communities (D’Antonio, 1986; Airoldi and Cinelli, 1997). The relationship between
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Figure 1 Captured sediment forms an important structural constituent of the algal turf. Prior to sediment deposition, the turf is uneven (a) with many void spaces. As sediment is captured, the surface becomes more uniform (b), reducing boundary roughness and local turbulence. In cross section, the depth of the captured layer sustained by the algae is apparent (c).
sediment transport and biota is therefore varied and becoming increasingly studied (Black et al., 2002; Orvain et al., 2004; Aspden et al., 2004a) and modeled (Widdows et al., 2004; Orvain, 2005). However, sediment movement and scour may also be responsible for promoting an increase in diversity and richness due to the creation of patches and maintenance of spatial and temporal heterogeneity, within what would otherwise be a fairly homogenous environment (Littler et al., 1983; McQuaid and Dower, 1990; Ford et al., 1999).
2.1. The physical background to life in depositional habitats Intertidal flats are largely found fringing the shorelines of estuaries although systems (typically narrower in extent in the shore-normal direction) are also found on the exposed coastal shores (e.g., as in Morecambe Bay, UK). Intertidal flats are – by definition – sedimentary environments found between the mean high-water and mean low-water spring tide datums (Dyer et al., 2000). The transport of sediments within intertidal flats is a continuous process, but for our present purposes, it can be subdivided into a number of contiguous processes: these are erosion (E), transport (T), deposition (D), and consolidation (C). This continuum of processes, driven by the local energy flux, is often referred to as the erosion, transport and deposition cycle (ETDC) (Tolhurst et al., 2006). The daily progression of the tidal wave across the intertidal zone is a fundamental factor governing sediment transport and hence system geomorphology. The tidal wave within estuaries but distant from the mouth is typified by currents that accelerate to a maximum at the mid-tide stage and then decelerate toward the high-water slack tide stage (Pethick, 1984). Further, due to asymmetric distortion of the tidal wave, flood currents are typically stronger (but of shorter duration) than corresponding ebb currents. These two features indicate firstly that the elevation of the intertidal flat within the tidal frame can be related to the local current velocities with peak current velocities associated with the mid-tidal flat area. Second, there is greater potential for a flood-directed sediment transport than ebb-directed movement, a situation which gives rise to intertidal flat regions being net, medium- to long-term depositional areas (Christie and Dyer, 1998; Bassoullet et al., 2000).
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While erosion of bed sediments can occur across the entire intertidal, it is typically most pronounced in the lower intertidal. The powerful horizontal currents associated with tidal flooding, especially in meso- and macrotidal areas (e.g., the Bay of Fundy, Canada, one of the highest tidal ranges in the world) transport the sediments shoreward. Recent research has indicated that the early flood current in these environments is in the form of a very thin (5–10 cm) ‘bore’, in which frictional stresses at the bed can generate exceptionally high entrainment rates (Bassoullet et al., 2000; Uncles and Stephens, 2000). The high levels of turbulence during tidal flooding are responsible for maintaining the sediments in suspension, thereby permitting their shoreward transport. Above the mid-tide datum, the tidal currents begin to slow and water deepens, and thus deposition of sediment can occur. Deposition on tidal flats is a function of both floc settling velocity and suspended sediment concentration (van der Lee, 2000) and therefore can be complex since each of these can vary substantially on both temporal and spatial scales. Intratidal variability in sediment deposition rates in relation to daily changes in these variables has been noted (Brown, 1998). Once sediment has settled to the bed, consolidation can occur, a phenomenon which is more acute for fine-grained, cohesive sediment than sand. Consolidation involves a collapse of the surface floc aggregate network and expulsion of porewater and the formation of a slightly more compact (and therefore erosion resistant) layer. The scale of this process on estuarine mudflats is millimeters, and it has been determined through careful measurements of bulk density of exposed sediments (Taylor and Paterson, 1998). If consolidation is able to proceed while the sediments are submerged, then the erosion resistance of the surface layers may be sufficient to withstand resuspension by the ebbing tide. It is through such a mechanism that accretion of the upper intertidal can gain a foothold, and thereby give rise to larger progradational scales (Allen, 1992). If the sediments can resist resuspension during the ebb tide, then the consolidation processes can be augmented during the period of subaerial exposure. Since the ebb removes the supportive fluid medium (i.e., the seawater), the collapse of surface floc aggregate networks is facilitated. It is feasible that the organic content of the sediments and floc particles helps the composite material to resist erosion. Once exposed, subaerial processes, including insolation and wind dehydration, can also act to increase the rate of compaction of surface sediment layers. Rain occurring during intermediate and low tides can act to reduce or prevent compaction and may have a major effect of intertidal sediment behavior (Tolhurst et al., 2006). The ETDC cycle on intertidal flat areas is relatively wellunderstood in process terms; however, a plethora of site-specific interactions influence these processes and makes sedimentation in these environments especially difficult to predict from first principles (Black et al., 2002). It is clear that bed sediments across intertidal flat areas can vary substantially and that all areas are subject to the transport of sediment through the ETDC cycle. These factors combine with the inherent biology of the system, the ecosystem engineers, to dictate the nature and mobility of surface sediments. The study of the sedimentology of intertidal flats can be traced back to the classic work of Evans (1965) in the Wash, UK. Evans provided the first comprehensive description of the sedimentary composition and sediment dynamics of
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temperate intertidal flat environments. He outlined a sequence of shore-normal depositional sediment types (or “facies”) which can be found between the low- and high-water marks and which can be related to the tidally driven hydrodynamic conditions across the flat. The occurrence of accelerating tidal currents on the lower flat area usually (assuming no sediment supply limitations) winnows any fine sediments and gives rise to a well-sorted sandy sediment bed. As the tide rises (and falls) the mid-flat area is exposed to the strongest currents found within the tidal frame and this also results in a coarser sediment texture although often there is a small secondary fine fraction. Shoreward of the mid-tide datum, current velocities monotonically decrease and this permits deposition of progressively finer sediments. The upper intertidal environment is, due to distortion of the tidal wave, dominantly depositional in nature. Although Evans’ geological-physiographic description is set principally within a tidal context, broad intertidal regions are susceptible to other hydrodynamic forcing factors which can influence the sediment character (and thus also the ecosystem composition), and these include wave action (driven by wind shear at the water surface during immersion of the tidal flat), and rainfall, which can directly impact surface sediments during periods of receding and low tide. Most tidal flats, especially those in northern temperate meso to macrotidal regimes, possess a convex shore-normal profile in which the mid- to upperintertidal areas have very low gradients (e.g., 1:1,000 or less). This means that a large proportion of the intertidal area is relatively high within the tidal frame, and consequently water depths are generally shallow or very shallow during tidal submersion/emersion. The shallow water column means that surface waves, created by local wind action on the sea surface, are able to “penetrate” to the seabed, and this can give rise directly to sediment resuspension (Soulsby and Damgaard, 2005). Resuspension of tidal flat sediments by waves has not, on the whole, been studied to the same degree as tidal resuspension, and yet wave action during periods of tidal flat submersion can be important and can even reverse the flood-dominant sediment transport characteristic of broad intertidal regions (De Jonge and van Buesekom, 1995). Black et al. (1998), for example, measured the tidally forced flux of suspended sediments on the north shore mudflats in the Humber Estuary, UK, and found that periods of wind in excess of 8 ms1 would increase the concentration of sediments in suspension by a factor of approximately 3. Winds from the north produced very short period waves capable of generating bottom erosion despite the extremely limited fetch. Wind-induced water waves are effective in transferring energy to the bed since the shear stress experienced under a fluid velocity gradient (air or water) is related both to the velocity and to the viscosity of the fluid (Vogel, 1994). Fluids of higher viscosity impart greater bed shear for the same relative flow. Water is approximately 70 times more viscous than air and therefore much higher velocities are required to entrain similar sediments in air than under water (Denny, 1993). Thus, wind erosion of exposed beaches is less common than water erosion but can be influential where wind exceeds a critical velocity over non-cohesive sandy deposits. This entrainment of sand classically leads to the formation of inshore sand dune systems. Wind erosion of cohesive (muddy) sediments is less common since they rarely dry out and particle cohesion resists wind-induced surface shear. Waves can only exert an influence when
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the tidal flat is submerged and thus the likely effect of waves can be investigated by comparison of the frequency and magnitude of wind events capable of producing waves with the phasing of tidal flooding on the flat. Although the lower intertidal flat is inundated for a greater proportion of time in relation to upper areas, the noncohesive nature of bottom sediments means that wave resuspension is offset by rapid settling of particles to the bed. Although the upper mid- and high-shore regions are inundated for less time during each tide, perhaps as little as one hour per tide for the mudflat–marsh transition, sediments are finer (and therefore of lower settling velocity), and hence resuspension by wave action can have longer lasting and more acute geomorphological consequences. Examples of marsh retreat (e.g., Wentlooge Flats, Wales) and shallow cliff formation (Pye, 2000) are a result of wave processes in the high-shore region. It is likely that for temperate and more northern intertidal flats, erosion by wave action is more important than hitherto supposed. The physical impacts of rain drops are also a key process on intertidal flats, particularly during subaerial exposure. Like waves, the overall importance of rainfall is contingent upon the relative phasings of rainfall events and the tidal cycle. Since the impact of rain on exposed sediments transfers momentum to the surface sediment layers, it constitutes an erosional force, albeit along a different axis than tidal- and wave-induced currents. The principal effect of rain is to weaken the interaggregate bonds through ballistic impacts and also through a general raising of the surface layer water content (Mwamba and Torres, 2002). Each of these changes will create a surface sediment which is weaker and therefore more susceptible to erosion by tidal and wave currents. One of the authors (Black) has observed numerous instances of the reversal of the flood sediment flux dominance following intense rainfall during subaerial exposure. While little attention has been directed toward quantification of this process, Tolhurst et al. (2006) attempted to determine the impact of rain drops on the surface of intertidal mudflat sediments. He found a 24–54% reduction in the critical entrainment stress for sediments following rainfall simulation. He also noted consequent changes (increases) in bed erosion rate at some sites. Although Tolhurst’s data are useful, and conceptually the influence of rain on sediment erosion is clear, the interaction of storm events of differencing intensities on intertidal flats and the importance of the coincident tidal phase (e.g., spring or neap) remains unexplored and is worthy of future attention.
2.2. The functional difference between mud and sand systems A sandy beach and a mudflat appear to share many similarities when viewed from a distance. Both are relatively flat and largely devoid of obvious structure which lend terrestrial and subtidal systems architectural complexity. According to Ricklefs and Miller (1999), this lack of structural complexity leads to fewer available niche spaces and a consequent reduction in species diversity. The ecology of both sandy and muddy systems is largely driven by visitors that feed from the surface and organisms that inhabit the sediments. However, on closer examination, the mud and sandy systems show considerable divergence. Physical difference in the grains can be manifest in surface topography. Ripple structure caused by waves are evident on sandy sediments while muddy sediments are usually planar with the exception of
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structures such as casts and burrows created by infaunal organisms. The cohesive mud is comprised of small (<63 mm) particles which cohere together while the sand comprises larger, independent grains (usually >100 mm). A full description of the properties of cohesive and non-cohesive beds can be found in many introductory texts on sedimentology and coastal systems (Davis and Fitzgerald, 2004) and in more specialist volumes (Soulsby, 1997; Winterwerp and Kesteren, 2004). However, it is the way in which the biology has to adapt that is important in terms of ecosystem function. The comparative assessment of ecosystem function across systems is still in its infancy and often rather subjective (Figure 2) and localized. (a)
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Figure 2 Coastal muddy (a) and sandy (b) systems. Selected ecosystem functions for coastal benthic habitats are illustrated and numbered in order of subjective “importance.” In each case, the number and distance of the secondary node from the central node (ecosystem functions) indicates a subjective measure of the importance of the function in a clockwise progression. The more important functions have lower numbers and are more closely associated with the central node. Ecosystem functions that might be considered “ecosystem services” are indicated with a triangular icon. Fisheries are flagged as potentially important in terms of secondary production dependent on local exploitation.
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The schematic diagram presented (Figure 2) cannot yet be properly based on hard evidence given that many ecosystem functions are difficult to measure and extrapolate to the appropriate scale. Rather this diagram indicates a conceptual model of habitat-specific ecosystem function taking into account relevant literature and experience. This can be used as a basis for discussion and development of hypotheses for testing. For example, we have judged secondary productivity to be more significant as an ecosystem function for muddy systems than sand flats given the greater organic production and higher biomass of prey species usually supported in intertidal muddy systems (Figure 2). There will, of course, always be exceptions to these generalities. The importance of fisheries will vary depending on the resource available and local customs. However, a conceptual view is useful but the order of functional importance may change depending on the nature of the habitat and with regard to the perspective of the researcher. A fisheries biologist is likely to place more emphasis on fish than on birds and an ornithologist vice versa. One of the problems of the ecosystem function approach and the assessment of habitat is precisely its subjective nature. Functional importance and ecosystem services are to a certain extent in the eye of the beholder. 2.2.1. Sandy systems Sandy systems provide a habitat that is highly dynamic, the smaller grains have been removed by hydrodynamic sorting or do not settle because of the high energy of the system and the sand bed is in almost constant motion. The grains are packed together but are sufficiently large that there is considerable void space between them allowing specialist organisms to inhabit the interstitial spaces (meiofauna). These voids also allow advective transport through the pore spaces and this allows oxygen penetration, driven by the pore water advection (Janssen et al., 2005). The traditional view of sandy habitats has been of relatively low organic content supporting an assemblage of filter-feeding organisms that largely rely on material from the water column. At a functional level, this leaves the sand as a rather inert substratum acting as housing for invertebrates. This view often incorporates the implicit idea that productivity, metabolic rates, and nutrient turnover are low in sandy environments as compared with the high organic content and active metabolism of mudflats. This has been recently, and effectively, challenged providing a different perspective (Buhring et al., 2006). Light penetration in sandy sediment is much deeper, and microphytobenthic photosynthesis can continue during tidal submersion, which is rare for muddy systems because the overlying water is turbid and occludes too much light for this strategy to be energetically valuable. The biomass of microphytobenthos is more dispersed with depth and can photosynthesize over much longer time periods. The integrated photosynthetic response can be much greater than previously considered (Billerbeck et al., 2007), thus the functional capacity of sandy flats to deliver carbon fixation (autochthonous production) may be a significant ecosystem service. Thus, the primary productivity of sand systems may be greater than first supposed on the basis of absolute chlorophyll content. Muddy sediments support active assemblages of microphytobenthos (Paterson et al., 2003; Underwood et al., 2005) with high level of
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biomass (Chl a) concentrated at the sediment surface. This concentration at the sediment-air interface is a requirement for photosynthetic organisms since light penetration on cohesive sediments is extremely limited (Consalvey et al., 2004). In terms of metabolic rates, De Beer et al. (2005) carried out studies demonstrating that sandy sediment systems had high aerobic degradation rates despite a much lower organic content than fine cohesive sediments with similar decomposition rates. Similar conclusions have been reached on the processing of external (allochthonous) carbon. Huettel et al. (1996), propose that organic mixing into the pore waters can occur due to hydromechanical mechanisms as well as bioturbatory and filtration activities of organisms. This transport of organic material into the sediment occurs within sediments, such as sandy sediments, which have permeability above a given threshold. Rather than considering sandy systems as areas with relatively little activity, the same system can be regarded as one that relies on the episodic allochthonous input of carbon which precipitates a period of extremely high turnover (Buhring et al., 2006). After the organic supply is exhausted, the system is quiescent again until another pulse of activity is initiated. This is similar to the model proposed for the organic supply to deep oceanic benthic systems. Both models provide a view of a relatively low turnover system for most of the time but the reasoning is different. Can the organic processing capacity of the sandy system be overloaded? In certain circumstances, the build up of organic material becomes too great to process. This is the situation that has led to the development of recognizable anoxic regions in some tidal flats termed “black spots” (Rusch et al., 1998).
2.2.2. Muddy systems Muddy systems have characteristically higher organic loads than sandy deposits. The presence of organic material associated with mud is partly due to in situ productivity but also due to the steady supply of allochthonous organic material from the coastal and riverine catchments (Nybakken and Bertness, 2005). This organic supply is deposited from the water column under conditions of relatively low energy. This supply is almost never completely exhausted but the close packing of the fine particles confines the oxygenic activity to the very surface layers (often <500 mm), and this establishes a system of extreme gradients and constant organic re-cycling (Aspden et al., 2004a). Nutrients are also usually plentiful given the charged nature of the very fine particles and their ability to sequester nutrients from the water column (Nedwell et al., 1999). The primary productivity of such systems is well known (Underwood and Kromkamp, 1999; Billerbeck, 2007) and their functional role in contributing to the primary productivity of the coastal zone often cited (Macintyre et al., 1996). This productivity is often enhanced by coastal eutrophication (Nedwell et al., 2002). In functional terms, mudflats have been given more prominence than sand flats, but this is probably debatable especially given some of the recent evidence cited above concerning the rates of metabolic processes that occur in sandy systems (Figure 2).
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The schematic representation of the importance of the functional attributes of sand and mud systems is entirely subjective (Figure 2) but presented to provide a basis for further thought. The importance of the two habits is often seen as quite varied, and it is easier to assign ecosystem services to muddy systems than to sandy ones. The evidence for this is actually more limited than might be imagined, and the recent work outlined above suggests that the systems may be closer in functional outcomes than was thought. The public perceptions of the systems are also quite different but each provides ecosystem services that cannot easily be replaced. The identification (Beaumont et al., 2007) and valuation (Beaumont et al., 2008) of these services is another matter and one that is now attracting considerable attention.
3. THE FUNCTIONAL ROLE OF BIOTA Different habitats may contain similar organisms that carry out the same ecosystem services. For example, the tidal flats of the Dutch Wadden Sea are composed of areas of varying sediment size. High densities of Corophium volutator occur in silty areas, and C. arenarium dominate in sandy areas (Beukema and Flach, 1995). It seems that the nature of the substratum is influential on the distribution of organisms that have a very similar lifestyle. On a broader scale, benthic assemblages can be placed in five categories according to the size and the type of substratum they inhabit (Table 2). In terms of the biodiversity-ecosystem function debate organisms are also often placed into categories related to their functional attributes (Figure 3).
3.1. Patterns of life Abiotic attributes, such as the erosional and depositional events (ETDC cycle) described above, are critical to the occurrence of species within a habitat and may determine the type of community that occurs (Sousa, 1984; Dernie et al., 2003). The changing nature of the physical environment and the immersion/emersion cycles lead to a zonation of species in the habitat. The evolution of strategies enabling organisms to adapt to life in sediments has led to different niches being exploited using a range of individual abilities. However, the outcome (fitness) of the species represents a successful balance that has been achieved between the driving pressures of nature and the possible niche spaces available. These pressures can be expressed as physical (wind and waves), biological (competition predation and grazing), and physiological (desiccation, salinity tolerance, temperature tolerance, etc.) (Figure 4). The most obvious outcome of these pressures are the simple distributional variations of species across environmental gradients that inherently represent these drivers. However, the classic extreme zonation of rocky shores is not found so readily within depositional systems (Nybakken and Bertness, 2005). This is because the sediments themselves act as an efficient buffer helping to reduce the effects of air exposure and minimize changes in factors such as salinity (Little, 2000). Salinity within the pore water varies much less than the salinity of the surface water
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Table 2 A broad classification of benthic fauna by size and lifestyle Microbenthos
Picoheterobenthos Picophytobenthos Microphytobenthos Microheterobenthos Meiobenthos
Hyperbenthos
Macrobenthos
Epibenthos
The microbenthos inhabit interstitial spaces between sediment particles and are comprised of unicellular organisms such as bacteria, diatoms, euglenoids, and ciliates. The photosynthetic, pigment-containing forms are known as microphytobenthos. When at high biomass, these organisms can be seen by the naked eye as a green or golden-brown biofilm in the sediment surface. These biofilms are adapted to survive depositional and highly dynamic environments (Paterson et al., 1999). Not only do microphytobenthic biofilms serve as primary producers in the ecosystem, but the group also has a number of other ecosystem services (Chapin et al., 1997), including the stabilization of cohesive sediment. The microphytobenthos are important primary produces and provide a significant source of autochthonous carbon within the ecosystem. The microbenthos could soon be commonly divided into picoheterobenthos, picophytobenthos, microphytobenthos, and microheterobenthos Prokaryotes such as bacteria and viruses Photosynthetic cells <2 mm Unicellular photosynthetic organisms >2 mm Unicellular heterotrophic organisms >2 mm The meiobenthos also occur within the interstitial spaces of sediments but consist of any multicellular organism less than 1 mm in length. This group consists of a variety of invertebrates such as nematodes and copepods The hyperbenthos occur in the water column just above the sediment surface. However, they are often found within the sediments due to the close proximity to the sediment bed and some may also burrow. Hyperbenthic organisms tend to be small, only a few millimeters in length The macrobenthos are over 1 mm in length and move freely through soft sediments. This group consists of species such polychaete worms, bivalves, and amphipods. This group has a major effect on their surrounding habitat and are important ecosystem engineers The epibenthos are often predatory and consist of large, active organisms, such as crabs, lobsters, and bottom fish that exploit the habitat and as such makes up a large proportion of the benthic biomass
(Lindberg and Harriss, 1973); however, there are always exceptional circumstances. On hot days at low tides, the sediments can be exposed to high evaporation rates and therefore the salinity of pore fluid can increase dramatically. These effects are more extreme on the high shore than the lower shore due to exposure time. Such variations in water content and evaporation help to drive patterns of change, and different life phases of intertidal organisms may show different tolerances and hence zonation patterns with age and development. It is considered that in muddy intertidal sediments, the upper limit of zonation is governed by the abiotic process, such as immersion period, salinity, and desiccation, whereas the lower limits are controlled by biotic factors such as the presence of competitors or predators. In this way, soft sediments are very similar to rocky shore environments (Little and Kitching, 1996). For example in the Wadden Sea, the lower limits of the
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Hediste diversicolor (O.F. Müller, 1776) Annelida: Polychaeta Omnivore
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Detritus feeder Filter feeder Active benthic predator, occupying a gallery of branching tubes. Hydrobia ulvae (Pennant, 1777) Mollusca: Gastropoda
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Deposit feeder with bacterial “gardening” Occupies a permanent U-shaped burrow, in muddy sediment habitats such as intertidal mud flats.
Figure 3
Examples of benthic infauna, their lifestyles, and functional roles.
populations of Corophium spp. were controlled by the presence or absence of Arenicola marina. In muddy sediments, in which A. marina were absent, Corophium spp. were able to extend to a lower tidal level than those in sandy sediments, in which A. marina were present. The occurrence of communities of decapod crustaceans studied along the Italian coast of the Tyrrhenian Sea was observed to be vertically controlled by the dynamics of the water flow and the sedimentation rates within the habitat (Scipione et al., 2005). The intertidal burrowing crab Chasmagnathus granulatus is an effective bioturbator and can modify the soft sediments in which it occurs to such a degree that the areas containing populations of
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Biological stress: competition for space and resources
Physical stress Physiological stress
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Figure 4 Balance of biotic and physical driving force on an intertidal shore.
C. granulatus are more humid, softer, and homogeneous than the areas without C. granulatus. As a result of the change in sediment properties, the vertical migrations of infaunal assemblages were observed to be greater in sediments with no C. granulatus compared to those with C. granulatus (Escapa et al., 2004). Thus, the distribution of macrofauna has to be carefully assessed since the driving factors vary with species, location, and physical forcing.
3.2. Effects of sediment disturbance Coastal management often requires an assessment of the outcome of various intrusive coastal activities on benthic communities. Knowledge of the effects of disturbance events on benthic communities and coastal sediments is an important part of coastal management. The knowledge of the recovery process for the habitat and in turn the recovery of the faunal assemblage is essential, and it is therefore crucial to understand and predict the association between biodiversity and ecosystem functioning. Physical disturbance, such as trawling (Cox, 1991; Dernie et al., 2003), the effect of foraging predators (Thrush et al., 1991), and bioturbation (Grant et al., 1982), can lead to an alteration in the sediment structure causing sediment properties such as sediment stability and permeability to vary (Thrush et al., 1996). This influences the faunal assemblage structure or composition to different degrees. Otter trawling has been shown to create a disturbance zone of up to 120–250 m wide (Schwinghamer et al., 1998) suggesting that the scale of disturbance may be larger than the footprint of the actual mechanism of disturbance. This disturbance increases the amount of resuspended sediment within the water column (Pilskaln et al., 1998; Watling and Norse, 1998), thereby reducing the amount of nutrients available. Jennings et al. (2001) found that bivalve harvesting resulted in an overall decrease in infaunal and epifaunal biomass although the
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biomass of polychaetes was unaffected. This indicates that specific life strategies/ functional groups are able to survive periodic disturbance better than others, suggesting that frequent harvesting will select for specific life strategies and perhaps specific functional groups. Therefore, if periods between harvesting activities were to change, it is probable that the communities present would also change. A similar phenomenon was observed in Hamilton Harbour (Bermuda). Large cruise liners would frequently disturb the sediments in the shipping lanes, resulting in major changes to the macrobenthic and meiobenthic communities (Warwick et al., 1990). Dernie et al. (2003) carried out a study on the effects of physical disturbance on both the habitat and fauna of a sheltered sand flat to ascertain whether the recovery of the biota could be predicted from physical attributes of the system. The outcome of this study suggested that benthic community recovery in a lower disturbance event took place within 64 days, whereas recovery after higher intensity disturbance did not occur until 208 days post-disturbance. Aspden et al. (2004b) determined that clam (Tapes phillippinarum) harvesting, in the Venice Lagoon (Italy), could be associated with decreases in sediment stability. A high frequency of harvesting prevented sediment consolidation, the normal succession of biological communities, and the development of stable microphytobenthic assemblages. The microphytobenthic assemblage composition did not significantly differ before and after harvesting at intensely harvested sites; however, community composition was affected at a site that had been subjected to negligible harvesting intensity prior to the study. Side scan sonar was used to visualise the physical disturbance to the bed before and after harvesting (Figure 5).
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Figure 5 Side scan sonar images of the same area of sea bed taken prior to (a) and post (b) the impacts of a harvesting event, aligned to a north bearing. The images of the sediment disturbance (b) provide an excellent visual representation of the damage caused by a harvesting event when compared to the homogeneous nature of the sediment prior to the harvesting event. A furrow can be seen clearly after the harvesting activity had taken place (indicated by the arrow, images Courtesy of Dr. R. Bates, University of St. Andrews).
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3.3. Biodiversity impacts Although physical attributes may have a strong influence in dictating the assemblages present, the biotic components of the systems must not be overlooked as they also play an integral role in habitat functioning. Faunal assemblages within the sediment carry out essential processes such as bioturbation, biodeposition, pelletization, biogenic stabilization, biogeochemical regulation, oxygenation of sediments, the distribution of solutes, and the transport of reactants and metabolites across the sediment water interface (Grant et al., 1982; Aller, 1994; Norling et al., 2007). The species-specific traits and the biodiversity of functional processes of benthic macrofauna can affect key functions of a community, habitat, and ecosystem (Solan et al., 2006). Ieno et al. (2006) further noted that changes in the composition of coastal benthic macrofauna changed the biogeochemistry of the system. An increase in species diversity was positively related to nutrient generation although this relationship was believed to be species specific. Bioturbatory activity was believed to be the key to the relationship, suggesting that species identity and density must be considered during studies of this nature (Biles et al., 2003). Changes in biogeochemistry can be subtle but there are many more obvious examples of the effects of organismal activity on habitat structure and function. Lanice sp. reefs may act as refuge for organisms such as bivalves, gastropods, polychaetes, amphipods, and isopods, and therefore these areas tend to be higher in individual numbers and diversity than areas of sediment with no reef structure present (Aspden, personal observation). Effects may be subtle interactions with other system drivers. The modification of bed roughness at the sediment surface due to biotic processes, such as the production of mounds or presence of tubes, can cause the flow to vary to such a degree that erosion thresholds change (Ziebis et al., 1996). Flume experiments, carried out by Eckman et al. (1981), studying the effects of Owenia fusiformis (a tube building polychaete) at varying densities, suggested that at higher densities the erosion threshold was lower than that at lower densities. Carey (1983) carried out similar experiments using the polychaete Lanice conchilega. The tubes of the polychaete caused a reduction in flow of up to 20%. Resuspension of sediments in front of the tubes occurred due to the alteration of the flow; however, in the wake of the tubes the flow was reduced, thereby allowing the deposition of sediments previously held in suspension. The presence of tubes has been suggested to encourage bacterial and microbial colonization within the surface sediments (Eckman, 1985), thereby stabilizing the sediments further due to microbial stabilization. These examples indicate organisms capable of mediating their habitats and surviving local conditions. However, particular organisms may depend on disturbance events to complete their life cycles, such as the sabellid worms which only spawn during periods of high storm activity (Barry, 1989). This dependence on a disturbance event is a response to the probability of high adult mortality due to the storm, therefore reducing competition for space and resources and providing optimum conditions for recruitment. To predict the impact of change in species diversity is strongly related not to the identity of the species per se but rather to the attributes of the species. Thus, the regulation of ecosystem functions and corresponding structural variables are
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influenced by species-specific traits rather than by species richness (Giller et al., 2004; Hooper et al., 2005). Heemsbergen et al. (2004) observed that it was diversity rather than richness that regulated soil respiration and loss of leaf litter mass in terrestrial systems. As a result of these earlier studies, it is becoming commonplace to consider the diversity of functional groups as opposed to species diversity or species richness when measuring ecosystem processes within coastal systems. Norling et al. (2007) defined functional biodiversity as the functional group richness of benthic fauna and classified dominant benthic invertebrates from the Baltic Sea and the Skagerrak as functional groups in accordance with their feeding behavior and sediment reworking activities. Great care must be taken when assigning organisms to a functional group, as many organisms can be defined as belonging to more than one group, showing functional plasticity (Boogert et al., 2006). It may be more appropriate to place organisms into a functional group based on an amalgamation of traits rather than one characteristic (Hooper et al., 2002); however, the number of traits that should be included in the determination of which group the species is assigned is open to debate. It is proposed that the number of functional attributes needed to be recorded (per species) will reach a maximum after which any extra functional attributes amalgamated will not alter the group into which the organism is placed. The dominant functions carried out by the organisms present may depend on the community assemblage and how the assemblage functions as a whole or/and the stage in an individuals life cycle.
3.4. Distribution in space and time In order to understand the relationship between species diversity and ecosystem processes, it is important to identify the functionally important species within the habitat being studied. It must also be taken into consideration that these relationships may vary with scale (i.e., local to global), and as the habitats/ecosystems get larger the effects of diversity change will change accordingly. Ninety-seven percent of the available living volume on the planet can be defined as marine environment (Raffaelli, 2006), providing a potentially huge volume to study. The examination of such large sample areas generally means that replication can rarely be adequate. In today’s studies, many of the functional measurements, such as primary productivity and nutrient flux, are measured at a very small scale. However, when contemplating these measurements in terms of entire habitats or ecosystem processes, it is not as simple as scaling up the results, as an increase in spatial variability also increases the variation of biodiversity. Approaches to this problem include large-scale mapping and remote sensing, but some experts believe that laboratory mesocosm studies may provide an empirical approach that can be applied and interpreted to help address scale issues (Benton et al., 2007). Also while many habitats may appear homogenous over a large scale (tens of kilometers), in reality they are patchy when examined at a smaller scale (Winberg et al., 2007). For example, the edge of seagrass meadows has higher abundances of organisms than those found within the bed. This is believed to be due to the settlement of drifting larvae, once they encounter the edge of the seagrass. This in turn attracts a higher abundance of predators, as was observed by Hovel and Lipcius (2002). Froja´n et al. (2006) demonstrated that the biodiversity of
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the dominant component of one benthic assemblage (polychaete worms) was highly site specific in terms of the patchiness, as was the spatial variability of seagrass beds within a mudflat. The role of system and species heterogeneity in ecosystem function is now being considered not only for the variation between regions or patches but also including aspects of the interconnectivity of different patches (Dyson et al., 2007). This work is relatively novel for intertidal systems, and the experimental designs tend to be complex but progress is being made. Results of functional studies may also differ with season. The growth of algal mats, kelps, seagrasses all induce different levels of spatial heterogeneity on a seasonal basis. Therefore, temporal variability must also be considered. The greatest limitations to such studies tend to be financial. Often this results in a lack of longterm data series and/or seasonal data. The collection of increased spatial and temporal data sets is required if ecosystem functioning is to be examined and understood fully at ecosystem scales (Raffaelli, 2006).
3.5. Trophic structure A food web can be examined from two view points: vertical and horizontal (Duffy and Stachowicz, 2006). The vertical aspect considers the length of the food chain, which if the trophic interactions change can have significant affects on ecosystem properties (Borer et al., 2005). The horizontal aspect concerns heterogeneity of biodiversity within the trophic levels, and biodiversity changes can affect ecosystem processes (Loreau et al., 2002). Highly diverse food webs tend to be less susceptible to changes in biodiversity due to the buffer effect of having more than one species filling a functional group. In comparison, low diversity, simple linear food chains may show dramatic changes to ecosystem processes following the extinction event of one species or trophic level. Therefore, an increased diversity within a functional group increases the value of the functions carried out due to the stability and the resistance the functional group has to diversity changes. For these reasons, studies on ecosystem function and how they are affected by biodiversity changes should include more research on trophic interactions within the system and how these are affected by changes in biodiversity (McCann, 2007). Changes in the dynamics of trophic interactions can have unforeseen consequences, and the loss of some species is actually more likely than others. For example, the probability of the extinction of species is related to body size with larger organisms more likely to become extinct than smaller ones (Solan et al., 2004). Since larger species are often predators at high trophic levels, this may have considerable influence on the trophic dynamics of the system. Changes that occur on the loss of a species are mediated by changes in food chain length (vertical) and changes in trophic level heterogeneity (horizontal) (Duffy, 2006). The most vulnerable systems are probably the most simple in food web terms. A habitat dominated by one or two species will have a simple trophic structure. A change in the limited biodiversity in a system like this can have major affects on functional processes. The loss of a single large dominant species is therefore more likely to remove a functional group (e.g., predator) with major cascading effects on the processes that occur, such as bioturbation, nutrient cycling, and decomposition
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(Loreau et al., 2002). The importance of the species pool has been supported by empirical studies indicating that a large pool of species allows a system to resist increasing external stress (resistance) that cannot be tolerated in depauperate systems. It is not known whether this is due to the presence of a few key species that dominate the ecosystem processes or due to the presence of many species with similar functional traits. This has particular importance for systems such as estuarine flats which are considered to be species poor. A further biological perturbation to intertidal systems is the introduction of invasive species (Williams and Grosholz, 2008). The debate about the relative harms, or sometimes benefits, of invasive species is still ongoing. Global climate changes will promote the expansion and reduction of species ranges and recognition of this and the mathematical modeling of effects is now being attempted for known invaders (Dunstan and Bax, 2007). Species have also often been introduced by man for some specific purpose such as the biocontrol of pests (the Rhodolia beetle) or for leisure activity (e.g., game fish). The aquatic system is no different and invasive species may have a large impact on coastal zone systems. The classic example is probably that of the green alga Caulerpa taxifolia in the Mediterranean. Caulerpa taxifolia is a popular aquarium alga and was released into the Mediterranean from the Muse´e Oceanographique de Monaco in 1984. Since this time, it has spread and replaced the natural seagrass meadows with its own simple monophyletic stands. This leads to a reduction in species richness and functionality of the systems and is considered by some to be a major ecological threat (Longepierre et al., 2005). The scale of the invasive problem is perhaps brought home by the work of Galil (2007) who lists more than 500 alien species from the Mediterranean alone. This might be considered as a considerable threat by preservationists wishing to retain the pristine and native charter of a system, but on a functional level, the debate is not so clear cut. If invasive species exploit unoccupied niche space, then can they not be seen as beneficial? This has certainly been argued (Oliverio and Taviani, 2003) and on a functional level it is not species identity that is important but the delivery of ecosystem services and the resistance and resilience of the system. For example, the invasive polychaete Marenzelleria cf. viridis (found in the Firth of Forth in 1982 and in the Firth of Tay in 1984) (Atkins et al., 1987) replaced the native species Hediste diversicolor. Although this invasive species may have changed the food web dynamics, there may have been little variation in trophic structure as both species are directly competitive and occupy the same functional group. Also, if invasive species increase biodiversity, then they may also stabilize the system in its response to stress. Whatever we believe about invasive species, they are almost impossible to stop, and with climate change, the term invasive may lose its meaning, overwhelmed by migration, where species move in response to changing climatic conditions. It is inevitable that we must become used to changes in species populations in coastal habitats.
3.6. New functional groups? Superficially, organisms inhabiting sandy and muddy intertidal systems have similar problems but the physicochemical habitat is quite different. Sands have much
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higher porosity and more advective exchange. Light penetration is substantial (10 cm) and the environment is oxygenic throughout the surface sediments. Mud has very tightly packed fine grains, low porosity, and very shallow oxic and photic zones. The body size of the organisms is very important in understanding the environmental pressures that they experience and this is not always intuitive. The smallest organisms (viruses, bacteria, and picoeukaryotes) require a great deal of further study. The picophytoplankton have only recently been considered as important in oceanic systems (Grob et al., 2007) and it seems that there is an equivalent group of small eukaryotic phototrophs in depositional systems which has hardly been identified or studied. The importance of the smaller organism, because of their abundance and activity, can be ecologically significant. French workers (Javanaud, 2006) have recently made the exciting preliminary finding that a functional group equivalent to the picophytoplankton (the picophytobenthos) make up to 18% of the bulk microphytobenthos community in muddy intertidal sediments. Many of these cells are not cyanobacterial as might be expected but are eukaryotic comprising an almost unstudied functional group of microbial benthic eukaryotic cells. The functional ecology of this group is almost entirely unknown and the importance of such a group in sandy sediments has not been investigated. This recent work indicates that we have not yet fully characterized all of the functional groups in depositional systems and that a great deal of work, particularly in the field of microbiology, is still to be done before we achieve a comprehensive knowledge.
4. FUTURE SHOCK: C LIMATE C HANGE AND ECOSYSTEM FUNCTION The first fossil evidence of life on the planet Earth dates about 3.8 billion years ago (Krumbein et al., 2003). The first recognizable functional ecosystem was produced by the accumulation of bacterial assemblages into the microbial mats that dominated the shallows of the Precambrian sea. These microbial assemblages trapped and bound sediments and, in the process, created the stromatolitic fossils. There are still a few existent stromatolitic systems where the nature and ecology of these mats can be studied (Dupraz and Visscher, 2005). Since their original formation, ecosystems have been evolving and changing as their component species developed new functional capabilities. Long-term change is arguably the normal status of ecosystems, but the functionality of most modern systems seems extremely robust. However, it is probably true that we are entering uncharted territory. Anthropogenic global change is a phenomenon we have not experienced before, and as the climate alters in response to the unnaturally high levels of CO2, the consequences of that change are difficult to predict. Species may move to invade new systems, reduce in distribution or become extinct, while new forms may be slowly evolving but the real crisis, from an anthropocentric perspective, will be when ecosystems no longer supply the goods and services we require. Coastal, intertidal, and wetland systems may be the first to degrade in this way. Wetland and shallow coastal areas may become drowned and what will replace them? There is
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little space available in the developed countries of Europe for new wetlands. The principles of managed retreat (Garbutt and Boorman, 2009), allowing the sea to encroach on the land and to promote the development of new coastal and wetland habitats, may be even more important when the functional role of the coastal margin is taken into account.
ACKNOWLEDGMENTS This chapter was written with support from the University of St. Andrews and Partrac Ltd. The authors acknowledge the support by the MarBEF Network of Excellence “Marine Biodiversity and Ecosystem Functioning” which is funded by the Sustainable Development, Global Change and Ecosystems Programme of the European Community’s Sixth Framework Programme (contract no. GOCE-CT2003-505446). This publication is contribution number MPS-08032 of MarBEF.
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B IOGEOCHEMICAL D YNAMICS OF C OASTAL T IDAL F LATS Samantha B. Joye, Dirk de Beer and Perran L.M. Cook
Contents 1. Introduction 2. Transport Processes on Intertidal Flats 3. Microbial Processes 3.1. Organic matter sources 4. Nitrogen Cycle 4.1. Nitrogen fixation 4.2. Nitrification and nitrate reduction 4.3. Exchange of dissolved nitrogen between the sediment and the water column 4.4. Benthic microalgal N assimilation 5. Phosphorus Cycle 6. Silicon Cycle 7. Concluding Remarks Acknowledgments References
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1. INTRODUCTION Coastal sediments are well recognized for their importance in the biogeochemical cycling of critical bioelements, such as carbon, nitrogen, and phosphorus (Alongi, 1997; Epstein, 1997; Yamamoto, 1997; Jassby et al., 2002). The coastal ocean is a shallow region (<200 m) that accounts for about 7% of the total ocean area (26 106 km2). These regions are noted by dynamic interactions between the land, ocean, and atmosphere and account for about 15% of global ocean productivity, 90% of sediment-based mineralization, and 80% of organic matter burial (Gattuso and Smith, 2007). Though the areas occupied by intertidal flats comprise a small fraction, roughly 7%, of the total shelf area (Stutz and Pilkey, 2002), they are among the most productive components of shelf ecosystems. Tidal flats are important sites of recycling for both terrestrially and marine-derived organic matter and nutrients and also support substantial rates of primary productivity. The best-studied tidal flats are in Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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Europe and the United States. Large areas of intertidal flats on the coasts of Africa (the Niger and Congo deltas), South America (the Amazon delta), and South Asia (the Ganges delta) are much less explored (Alongi, 1997). The largest intertidal flat area, the Wadden Sea, is to a large extent man-made. The original ecosystems of the Wadden Sea were salt marshes, mudflats, and seagrass beds that developed after the last ice age. The salt marshes disappeared due to peat harvesting and the flood plain areas decreased, primarily due to land reclamation. The sediments of the Wadden Sea are changing via a gradual displacement of silt with sands, especially toward the barrier islands and the large tidal channels where intertidal sand flats now dominate (Lotze, 2005). Similar coastal areas in the United States are still salt marshes. The expected increase of the sea level will affect the intertidal areas worldwide. Where dikes protect coastal regions, the total area occupied by tidal flats will shrink substantially. Intertidal flats are shallow coastal areas located between the spring high- and low-tide marks that lack rooted vegetation. Intertidal flats are highly heterogeneous and differ in sediment composition and microbial processes depending on their location. Because they are found in sheltered bays, estuaries, or are protected from exposure to oceans by barrier islands, they can range from mudbanks to coarse sand flats and receive varying degrees of terrestrial inputs from rivers and coastal seas. Intertidal flats can consist of lithogenous or biogenous sediments or mixtures of both sediment types. Lithogenous sediments originate from eroded rock and are rich in silicates while biogenous sediments are mainly carbonates, usually deriving from shells or corals. Biogenous sediments are more abundant in the tropics, especially in coral seas. This chapter will focus on the temperate tidal flats dominated by lithogenous material. The grain size distribution of lithogenous sediments ranges from fine clay (<2 mm) to coarse sand (>250 mm) and is determined by the hydrodynamic regime (Dyer, 1979). Clay-rich and silty sediments exhibit porosities ranging from 50% to over 80% (Taylor et al., 1966). Sands, on the other hand, exhibit porosities ranging from 37 to 50%, with fine sands having a slightly higher porosity than medium-grained sands (Taylor et al., 1966). Typically, an inverse relationship is observed between grain size and nutritional content (often reported as % organic carbon) (Longbottom, 1970; Dale, 1974; Watling, 1991; Alongi and Christoffersen, 1992; Vigano et al., 2003). Mudbanks and silty tidal flats are usually rich in organics (typically 0.5–3%), nutrients, and trace metals that easily bind to the large surface area of the clay particles while sandy sediments contain little organic matter (around 0.1%) and contain less metals and nutrients. However, some studies have failed to show a strong relationship between grain size and biological parameters such as bacterial abundance or chlorophyll a (Dale, 1974; Cammen, 1982; DeFlaun and Mayer, 1983; Albrechtsen and Winding, 2004) while others have illustrated good correlations between grain size, organic content, and microbial activity (sulfate reduction; Hargrave, 1972; Ziebis et al., 1996; Panutrakul et al., 2001). When storms hit tidal flats, large amounts of sediments can be suspended and reposited (Graber et al., 1989; Christiansen et al., 2006), thus mixing the top centimeters to decimeters of sediments irregularly. Bioturbation provides a constant reworking of the top 10 cm (Boudreau, 1998). This, together with their daily cycles
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of exposure and inundation makes tidal flat sediments highly dynamic. The inundation/exposure ratio strongly varies with distance from the low water line and is further dependent on the tidal range (Green et al., 1997). Moreover, there is temporal variability, with biogeochemical processes in tidal flats highly dependent on season and weather (Guarini et al., 1997; Le Hir et al., 2000; Widdows et al., 2004). Due to the heterogeneity within and between flats, it is difficult to generalize the biogeochemistry of intertidal flats. In fact, their temporal and spatial heterogeneity and their extreme dynamics are typical characteristics that distinguish these habitats from permanently inundated coastal sediments (Kornman et al., 1998; Le Hir et al., 2000; Herman et al., 2001; Widdows et al., 2004). The relative inaccessibility of intertidal flats has, to a large extent, limited their study. Being intermittently inundated and exposed to strong currents during rising and falling tide, a solid ship that can fall dry is needed to access the flats during complete tidal cycles. Explorations on foot during low tide must be planned carefully and can be difficult in soft muds and driftsands. Storms are interesting events, as this is when most sediment reposition occurs, but during such occasions the flats must be abandoned. As a consequence of the spatial and temporal heterogeneities and the difficulty of their in situ study, the phenomena reported here may be conceptually valid but may not accurately reflect the in situ situation. In this chapter, we first present an overview of the transport processes that dictate the physical nature of intertidal flats. Next, we discuss a variety of microbial processes that occur on tidal flats. We give particular attention to the microbially mediated cycling of nitrogen, phosphorus, and silicon on tidal flats.
2. T RANSPORT PROCESSES ON INTERTIDAL FLATS Biogeochemical processes in tidal flat sediments are strongly regulated by transport. Solute and particulate exchange processes between seawater and sediments and within the porewater are regulated by diffusion and advection. The dominant transport mechanism for solute exchange is advection, which is driven by water currents and waves and by bioventilation and bioturbation. The solid matrix is reworked by hydrodynamics and bioturbation; therefore the top 10 cm of sediment is often rather homogeneous. Diffusion can also be important in some cases. Whether diffusion or advection is dominant is determined by the sediment permeability for water flow, as illustrated on the extreme cases, muddy flats, and sand banks. Muddy fine-grained sediments are impermeable to water flow, so solute transport is diffusion controlled. In such sediments, oxygen penetrates a few millimeters (Llobet-Brossa et al., 2002), almost independent of currents and tides. Diffusion is described by Fick’s law of diffusion. The flux rates ( J ) are determined by the steepness of the solute concentration gradients (dC/dx) and the effective diffusion coefficient (Deff) for the solute (Berner, 1980; Boudreau, 2000): J = Deff
dC dx
ð1Þ
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The effective diffusion coefficient in sediments is the diffusion coefficient of the solute in water (Dw) corrected for the tortuosity ( ) and porosity ( ) of the sediments (Berner, 1980; Boudreau, 2000) Deff = Dw 2
ð2Þ
An empirical approximation for fine clays with high porosity (>0.7) is commonly used (Ullman and Aller, 1982): Deff = Dw 2
ð3Þ
Organic matter that is deposited on top of fine-grained sediment is transported below the surface through bioturbation and bioventilation. In porous sands, water–sediment exchange and transport inside sediments is governed by advection, porewater flow (v) driven by currents and waves (Berner, 1980; Boudreau, 2000): dC J = Deff þ vC ð4Þ dx As at even very low flow velocities, diffusion can be ignored so that the second term is sufficient to describe the flux. While advective transport seems simple in theory, the local solute concentration and an estimate for the porewater velocity must be known. In practice, these terms are very difficult to quantify. Since porewater flow is driven by currents and waves, all factors that influence the local hydrodynamics change the advectional flow regime. Such factors include measuring equipment, vehicles and ships, and the observers themselves. Thus, minimally invasive techniques need to be used for in situ measurements in permeable sediments. Advective exchange processes are strongly enhanced by the so-called ripple flow. The principle was discovered on gravel beds in rivers (Thibodeaux and Boyle, 1987) but appears to be quite important in sandy sediments (Huettel and Gust, 1992; Forster et al., 1996; Røy et al., 2002; Huettel et al., 1996; Ziebis et al., 1996). This concept can be understood from the small pressure gradients that are induced by a current traveling over a ripple. Due to the Bernoulli effect, the pressure near the top of the ripple, where flow velocity is enhanced, is lower than that in the valley between the ripples. This pressure difference at the surface drives porewater flow through the sediments: porewater enters the sediments in the valleys and leaves the sediments near the top of the ripples (Figure 1). Ripple-induced flow can be significant and leads to a distinct oxygen distribution, where oxygen penetrates more deeply below the valleys and anoxic zones reach almost the sediment surface near the ripple peak. Such flow strongly enhances the exchange of solutes, such as oxygen, nutrients, and metals (Mn2þ and Fe2þ) (Ziebis et al., 1996; Berninger and Huettel, 1997; Huettel et al., 1998; Huettel and Rusch, 2000; Rusch and Huettel, 2000; Røy et al., 2002; Rasheed et al., 2003; Precht and Huettel, 2004). As currents and waves fluctuate rapidly, oxygen penetration through porous sediments is highly dynamic (de Beer et al., 2004; Precht and Huettel, 2004; Franke et al., 2006; Werner et al., 2006). Ripple-induced flow
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Pressure at the sediment surface
0
Time = 0 min
Scale (mm)
3.2 6.4 9.6 12.8 16 19.2
250 225 200 175 150 125 100 75 50 25 0
Oxygen (μmol/L)
Direction of flow
Mobile sediment ripples
Figure 1 The left panel shows the flow direction (upper arrow), the pressure gradient, and the resulting patterns of water flow around ripples on a tidal flat. The right panel shows oxygen concentration distributions through a ripple, illustrating the change in redox regimes associated with the ripple-induced flow regimes (adapted from Precht et al., 2004).
governs the upper 5–10 cm of the sediments, and generates a constantly changing mosaic of inflow and outflow areas and thus a constantly changing pattern of oxic and anoxic volumes in the upper sediments (Figure 1). When considering the water budget of a tidal flat, one might conclude that the volume of water entering the flats equals that leaving the flats. Normally, this mass conservation rule holds for small areas, on the order of square decimeters; however, in larger flats with porous sediments, a much larger and deeper circulation pattern exists. During low tide, the flats remain covered with a thin layer of water and remain waterlogged. Consequently, a pressure head exists that drives a flow from the center of the flats toward the low water line, where water seeps out during low tide. The porewater near the seeps are highly enriched in degradation products such as nutrients, dissolved inorganic carbon, sulfide, and methane (Billerbeck et al., 2006a). At high tide, when the flats are completely covered, the pressure differences are dissipated and this flow stops. Thus an intermittent, unidirectional internal circulation pattern is common on tidal flats. Based on this concept, porewater flow patterns were quantitatively modeled in a permeable flat in the German Wadden sea (the Jansand) using ComSol (Røy et al., 2008). The calculated flow velocities were on the order of centimeters to millimeters per day, depending on the length of the pathway. The residence time of the porewater in the flats was found to vary from months to centuries and the average residence time of the porewater leaving the flat near the low-water line was estimated to be several decades. This was confirmed by 14C-dating of methane and DIC in the seeping fluid, which still carried the so-called bomb signal, a peak in radioactivity caused by the hydrogen bomb testing in the early 1960s. The level of 14 C was equal to the atmospheric levels of about 1970 (Røy et al., 2007). Another factor complicating the water budget on tidal flats is groundwater flow. Shallow groundwater has long been recognized as a dynamic and potentially important source of nutrients and dissolved organic carbon to coastal waters (Runnels, 1969; Johannes and Hearn, 1985; Li et al., 1999; Bhadha et al., 2007). Despite significant methodological advances in recent years (summarized in Burnett et al., 2003, 2006), quantifying groundwater inputs to coastal waters remains a
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High tide
Low tide
Figure 2
Flow through porous sediments at high and low tide.
challenge and, as a result, groundwater inputs are often the least constrained component of coastal nutrient and carbon budgets (Moore, 1996; Burnett et al., 2001, 2002, 2003, 2006; Moore et al., 2002). Groundwater inputs to coastal waters are believed to equal about 6–10% of surface water inputs globally (Burnett et al., 2003). However, the nutrient load associated with groundwater may rival river inputs because nutrient concentrations in groundwater often exceed those in surface waters (Krest et al., 2000; Moore et al., 2002). Recent studies (Sassa and Watabe, 2007) have illustrated the dynamic nature of groundwater flow on tidal flats but additional work is required to develop a comprehensive understanding of the role of groundwater flow in tidal flat dynamics and nutrient budgets. In summary, two primary transport phenomena control exchange between porous flats and seawater (Billerbeck et al., 2006b) (Figure 2). Surficial “skin” exchange occurs during high tide and is driven by currents and waves, which is further enhanced by ripple flow. During low tide, deep body circulation occurs, leading to seepage of highly reduced porewater at the low water line. The deep body circulation might lead to a significant release of methane and nutrients (Middelburg et al., 2002; Røy et al., 2007). The consequences of these fluxes for global elemental budgets needs further study.
3. MICROBIAL P ROCESSES The role of microbes in element cycling and biogeochemistry of tidal flats is easily underestimated. The original concept that life on tidal flats is dominated by herbivorous macrobenthos has been replaced by concepts where microbial primary producers and heterotrophic microorganisms regulate energy flow. For example, all benthic filter- and deposit feeders (bivalves and polychetes) were estimated to consume less than 25% of the yearly planktonic primary production; the remaining was either washed out to sea or microbially degraded (Warwick et al., 1979). Similarly, in tidal flats of the Wadden Sea, 75% of the annual organic input (the sum of primary production and deposition of detritus) was primarily degraded by microorganisms, with only 25% consumed by macrofauna (Kuipers et al., 1981). As in other sediments, the biogeochemical processes in tidal flat sediments are more or less stratified. Primary production is limited to the photic zone of several
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millimeters (de Beer et al., 2004; Werner et al., 2006) during low tide or an even thinner zone during inundation (Billerbeck et al., 2007). Chemoautotrophic processes occur in the oxic zone, particularly the oxidation of ammonium, reduced iron (Fe2þ) and sulfur, but these CO2 fixation rates are much lower than photosynthesis. A first step in the degradation of organic matter is hydrolysis of biopolymers, a complex process in which the cytophaga-flavobacterium (CFB) consortium is essential (Kim et al., 2004; Moss et al., 2006; Musat et al., 2006). The main reductive processes are stratified in the classical sequence of electron acceptors: O2, NO3–, Mn(IV)/Fe(III), SO42–, and CO2. Degradation of organic matter is intense in the oxic zone, denitrification and sulfate reduction rates are highest near the oxic–anoxic transition, and these processes do not appear to be entirely inhibited within the oxygen-rich zone (Dilling and Cypionka, 1990; Marschall et al., 1993; Wilms et al., 2006). In marine sediments, approximately half of the organic matter is degraded by sulfate reduction (Sørensen et al., 1981; Jørgensen, 1982; Fossing and Jørgensen, 1989, 1990). Methanogenesis was thought to be of minor importance in the majority of marine sediments, due to competition with sulfate reduction (Reeburgh, 1983). In particular, on well-flushed tidal flats, frequent replenishment of the sulfate pool was thought to limit the importance of methanogenesis. This concept might need revision however, as significant methane effluxes from estuaries occur (Middelburg et al., 1996, 2002) and a clear sulfate–methane transitionassociated distribution of sulfate reducers and methanotrophic archaea (Wilms et al., 2007) suggests that methane production is important in tidal flat sediments. Complicating the distribution of microbial activity further is the complex 3D geometry of oxic and anoxic zones that occurs around animal burrows (Aller, 2001; Hongguang et al., 2006).
3.1. Organic matter sources The unique position of tidal flats at the interface between the land and the sea means that organic matter originating from a wide range of terrestrial and marine sources, including organic matter exported from adjacent communities, including mangroves and salt marshes (Meziane et al., 1997; Meziane and Tsuchiya, 2000), terrestrially eroded material such as peats (Volkman et al., 2000), and material from anthropogenic sources (Meziane and Tsuchiya, 2002) is deposited within these environments. Detritus from marine sources, such as seagrasses, as well as phytodetritus from pelagic algal blooms also contributes to the organic matter found on tidal flats (Volkman et al., 1980; Rohjans et al., 1998). Autochthonous production of organic matter by benthic microalgae, macroalgae, and bacteria also contributes significantly to the organic matter pool in most tidal flat sediments (Volkman et al., 1980; Meziane et al., 1997; Meziane and Tsuchiya, 2000, 2002). It is important to note that while the organic content of the sediment can sometimes be dominated by terrestrial sources, it is unlikely that this refractory organic matter is mineralized extensively. Rather, a majority of sediment respiration on tidal flats appears to be driven by autochthonous production by benthic microalgae (MacIntyre et al., 1996; Miller et al., 1996; Moens et al., 2002; Cook et al., 2004b). In sandy tidal
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flats where large volumes of water are advected through the sediment, pelagic algae may make a more significant contribution to the labile organic carbon available in the sediment (Bianchi, 1988; Bianchi and Rice, 1988; Bianchi et al., 1988; Pilditch and Miller, 2006).
4. N ITROGEN C YCLE Nitrogen is a critical nutrient element that is required by all organisms for cellular biosynthesis. Primary production and/or decomposition are often limited by the availability of fixed nitrogen (Ryther and Dunstan, 1971; Smith, 1984). In addition, oxidized nitrogen species, mainly nitrite or nitrate, can be used as a terminal electron acceptor during anaerobic respiration. The sediment nitrogen cycle has received considerable attention in recent years because of the key role of nitrogen in regulating coastal productivity (Canfield et al., 2005; Joye and Anderson, 2008). Here we focus on the literature relating specifically to tidal flats.
4.1. Nitrogen fixation Nitrogen (N2) fixation in aquatic sediments is carried out by bacteria and some members of the archaea (Canfield et al., 2005). Rates of N2 fixation by phototrophic benthic cyanobacteria are generally much higher than rates supported by heterotrophic bacteria and chemoautotrophs; however, their importance to total system N2 fixation is often constrained by their limited areal coverage (Howarth et al., 1988b). The exception to this appears to be in the tropics, where cyanobacteria-dominated microbial mats often dominate the sediments of shallow subtidal and intertidal areas and contribute high rates of N2 fixation to the ecosystem (Pinckney and Paerl, 1997; Paerl et al., 2000; Joye and Lee, 2004; Lee and Joye, 2006). Tidal flat cyanobacteria are associated with high rates of N2 fixation in temperate habitats as well (Gotto et al., 1981; Bautista and Paerl, 1985; Howarth et al., 1988b; Stal, 1995; Cook et al., 2004b). Many factors identified as being conducive to cyanobacterial growth and N2 fixation exist on tidal flats (Howarth et al., 1988a). Relatively, high light levels mean that energy is not a limiting factor for the ATP-demanding process of N2 fixation. Trace metals such as molybdenum and iron are required for synthesis of nitrogenase, the enzyme that mediates N2 fixation; these metals are more abundant in sediments than in the water column and may potentially stimulate sediment N2 fixation. High concentrations of dissolved organic carbon within the sediment may also chelate trace metals, increasing their bioavailability. In fact, trace metal additions do not stimulate N2 fixation in tidal flat sediments (Paerl et al., 1993; Lee and Joye, 2006), suggesting that benthic N2 fixing microbes are not metal limited. Most of the terrestrially derived bioavailable phosphorus entering the coastal zone is adsorbed to particle surfaces. Upon deposition of these particles on tidal flats, reductive dissolution may lead to a release of phosphorus from the sediment, enriching the benthic flux of P relative to N, thus potentially stimulating N2 fixation.
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Strong diurnal variations in N2 fixation are often observed depending on the type of cyanobacteria present. Where nonheterocystous cyanobacteria such as Oscillatoria spp. dominate tidal flat communities, maximum rates of N2 fixation may be observed at sunrise and sunset or in the dark because oxygen inhibits nitrogenase activity during the day (Stal et al., 1984; Villbrandt et al., 1991; Joye and Paerl, 1994; Cook et al., 2004a). In contrast, heterocystous cyanobacteria may exhibit maximum N2 fixation rates during the day (Carpenter et al., 1978; Currin and Paerl, 1998; Lee and Joye, 2006) although some heterocystous species may temporally decouple N2 fixation and photosynthesis (Paerl, 1990). Variations in N2 fixation rates have also been observed depending on inundation and exposure, with reduced N2 fixation rates being observed during submersion in epiphytic cyanobacterial communities (Currin and Paerl, 1998). No difference in N2 fixation rates was, however, observed between inundation and exposure on a tidal flat sediment (Cook, 2003). Inorganic N availability also plays an important role in regulating N2 fixation rates (Carpenter et al., 1978), and rapid decreases in N2 fixation rates have been observed on tidal flats in response to nitrate delivery by storm runoff events (Joye and Paerl, 1993). However, in some tropical microbial mats, only high concentrations of ammonium (>500 mM) inhibited N2 fixation (Lee and Joye, 2006). In temperate systems, strong seasonal variations in N2 fixation on tidal flats are often observed with the highest rates generally occurring during spring and summer, when light levels are highest and nitrogen is most limiting (Carpenter et al., 1978; Cook et al., 2004a; Villbrandt et al., 1991). In tropical systems, rates of N2 fixation are less variable seasonally and differences in activity are driven mainly by changes in precipitation or tidal inundation (Lee and Joye, 2006). In comparison to temporal variations in N2 fixation, much less is known about the spatial distribution of N2 fixation on tidal flats because most published studies so far have focused on light and nutritional controls on N2 fixation. Substantial gradients in algal biomass and N2 fixation occur across tidal flats and activity varies by an order of magnitude over distances of meters (Cook et al., 2004a). It is likely these gradients are mediated by light levels, exposure to wave energy, and grazing. Quantifying the exact rates of N2 fixation is complicated by the fact that the generally used technique for measuring N2 fixation, the acetylene reduction assay, is an indirect method. Thus, assumptions must be made to convert moles of acetylene reduced to N2 equivalents. For cyanobacterial mats, a factor close to the theoretical range of 3–4 can be used based on calibrations of the acetylene reduction assay with 15N2 (Howarth et al., 1988b; Cook et al., 2004a; Lee and Joye, 2006). For consistency with the existing literature, we use a ratio of three to convert acetylene reduction rates to N2 fixation rates when discussing activity below. Another complicating factor is that rates are often reported relative to chlorophyll a rather than on an areal basis (Potts, 1979) making it difficult to compare rates between studies. In general, rates of N2 fixation in cyanobacterial mat communities are among the highest measured in aquatic systems, ranging from 0.09 to 5.4 mol N/m2/ year. These rates are much greater than those measured in unvegetated (0–0.11 mol N/m2/year) or vegetated (0–3.64 mol N/m2/year) subtidal sediments (Howarth
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et al., 1988b; Joye and Anderson, 2008). Therefore, one would expect tidal flats colonized by cyanobacteria to exhibit some of the highest N2 fixation rates observed in aquatic ecosystems. Given the high spatio-temporal variation in N2 fixation rates and the limited data available, we have represented the rates measured on mudflats on an hourly basis. Rates are often essentially 0 when there is low cyanobacterial biomass (during autumn and winter). During spring and summer, when cyanobacteria are abundant, rates generally are in the range of 10–100 mmol N/m2/h with extreme highs of 350 mmol N/m2/h also noted (Carpenter et al., 1978; Gotto et al., 1981; Stal et al., 1984; Bautista and Paerl, 1985; Abd. Aziz and Nedwell, 1986; Joye and Paerl, 1993; Cook et al., 2004a).
4.2. Nitrification and nitrate reduction Nitrification refers to the sequential oxidation of ammonium to nitrite and ultimately nitrate. Nitrification is regulated by substrate availability (oxygen and ammonium) as well as by hydrogen sulfide, which inhibits ammonia oxidation, the first step of the process (Joye and Hollibaugh, 1995; Joye and Anderson, 2008). Nitrification links the reduced and oxidized portions of the nitrogen cycle and is often closely coupled to nitrate reduction processes. The most well-studied pathway of nitrate reduction is denitrification, the process whereby nitrate is utilized as a terminal electron acceptor for an oxidation reaction and is reduced to nitrous oxide or dinitrogen gas. Denitrification has received much interest over the past 30 years because it represents a significant sink for fixed nitrogen in aquatic ecosystems. Denitrification is an obligately anaerobic process and only takes place below the oxic surface layer of sediment. Conceptually, denitrification is often partitioned into denitrification fueled by nitrate transported into the sediment from the water column and that produced within the sediment by nitrification. If water column nitrate concentrations are low, denitrification will be driven mainly by nitrification within the sediment. As water column nitrate concentrations increase, denitrification fueled by the nitrate flux from the water column will increase proportionally; however, this may plateau at extremely high nitrate concentrations (Dong et al., 2000). On tidal flats, the highest rates of denitrification take place when the sediments are exposed to high nitrate concentrations, such as those on the European north Atlantic coast where nitrate concentrations can be up to 1 mM (Nedwell et al., 1999; Dong et al., 2000). Because water column nitrate concentrations are generally highest during winter, the maximum rates of denitrification are often measured during the winter as well (Ogilvie et al., 1997; Cabrita and Brotas, 2000). During exposure, the nitrate source from the water column is eliminated, leaving only sediment nitrification to fuel denitrification. To date, only one study has measured denitrification during exposure of intertidal flats, finding that denitrification rates were similar during exposure and inundation at night (Ottosen et al., 2001). Coupled nitrification–denitrification was much lower during the day, possibly because increased salinity and competition for NH4þ with benthic microalgae inhibited nitrification. These authors also noted that denitrification rates were three times higher in a nearby subtidal location, suggesting that intertidal conditions
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may have a generally inhibiting influence on nitrification and denitrification. A similar observation was made by Cook et al. (2004a) who found primarily coupled nitrification–denitrification rates were consistently highest on the lowest parts of tidal flats. This observation is not universal, however, and we note that Cabrita and Brotas (2000) observed no effect of elevation on denitrification rates. On a more localized scale, there have been observations of differences in sedimentary nitrification (and presumably denitrification rates) between ridge and runnel structures on tidal flats, with nitrification rates being sixfold higher in runnel structures (Laima et al., 1999). Another important factor influencing nitrogen cycling on tidal flats are the ubiquitous benthic microalgae that coat the flats (Figure 3). In addition to directly assimilating inorganic nutrients, the activity of benthic microalgae influences nitrification and denitrification through two primary mechanisms: alteration of redox regimes and direct assimilation of inorganic nitrogen (de Beer, 2000, 2001). Their production of oxygen during photosynthesis increases the depth of the oxic zone in the sediment, which may enhance nitrification and hence coupled nitrification– denitrification (Risgaard-Petersen et al., 1994; An and Joye, 2001). Denitrification driven by nitrate supplied from the water column will be reduced, however, as the distance that nitrate has to travel through the sediment oxic zone to the zone of denitrification is increased (Risgaard-Petersen et al., 1994). N2 Water NH4+
NO3–
NH4+
NO3–
NO3–
NH4+ Org N
Org N Sediment Inundation
Exposure
Figure 3 A conceptual diagram of aspects of the N cycle pertinent to tidal flats. Benthic microalgae (gray circles) play a key role in the N cycle on tidal flats assimilating N from both the sediment and the water column, the flows of N through benthic microalgae are likely to dominate the N cycle. Organic N is mineralized in the sediment releasing NH4þ, a large fraction of which is assimilated by benthic microalgae. Nitrifying bacteria may be out competed by benthic microalgae for NH4þ, and thus nitrification rates are likely to be low (thus limiting denitrification rates), particularly in oligotrophic and mesotrophic systems. In eutrophic systems, rates of denitrification may be high, driven to a large extent by NO3^ from the water column, and under these circumstances tidal flats mat be an important sink for fixed N. During exposure, exchange between the sediment and the water column is cut off resulting in a depletion of NO3^ in the porewaters and an enrichment in NH4þ which may be released upon inundation. Where the biomass of cyanobacteria is high, N2 fixation rates may be extremely high, and in these circumstances tidal flats may be a net source of fixed N. The thickness of the arrows gives a rough indication of the relative magnitude of the N flows.
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Benthic microalgae also influence nitrification and denitrification by competing directly with nitrifying and denitrifying bacteria for inorganic nitrogen. Mesocosm studies have clearly demonstrated a negative impact of benthic algae on nitrification and denitrification rates (Risgaard-Petersen, 2003, Risgaard-Petersen et al., 2004), and it appears that the competition for nutrients arises from both benthic microalgae and from actively growing heterotrophic bacteria that are fed by algal exudates. Overall, the impact of benthic microalgae on denitrification seems to be negative based on a meta-analysis of denitrification from 18 European estuaries: net autotrophic sediments had significantly lower rates of denitrification than did net heterotrophic sediments (Risgaard-Petersen, 2003). These findings also agree with field studies of denitrification on tidal flats which have reported negative correlations with benthic microalgae-associated parameters (such as chlorophyll a and photosynthesis) and denitrification (Cabrita and Brotas, 2000; Cook et al., 2004a). The strong influence of benthic microalgae on denitrification means that there may be large changes in coupled nitrification–denitrification rates under light versus dark conditions (Dong et al., 2000; Cook et al., 2004a). However, benthic microalgae can also assimilate inorganic nutrients during the dark periods, so competition for nutrients between nitrifying and denitrifying bacteria and benthic microalgae can still occur during dark periods (Porubsky et al., 2008a,b). During dark periods, oxygen inhibition of denitrification is eliminated such that denitrification rates may be stimulated to some extent. For the most part, however, prolonged dark incubations (>48 h) are required to remove the negative impact of benthic microalgae on sediment inorganic nutrient fluxes (assimilative consumption) and nitrification–denitrification (oxygen inhibition) on denitrification rates (Porubsky et al., 2008a). As noted earlier, fluid flow through sandy tidal flats drives significant rates of solute transport via shallow and deep advective processes. No studies have investigated the influence of these transport processes on nitrification and denitrification on tidal flats. However, based on recent data, one would expect shallow advective transport processes to substantially increase rates of denitrification over those observed in cohesive sediment, particularly if high nitrate concentrations exist in the water column (Cook et al., 2006). The phenomenon of deep porewater transport in tidal flats is analogous to nitrate-rich groundwater transport which stimulates denitrification (Hoffmann et al., 2000; Addy et al., 2002). So, we predict that deep porewater transport in tidal flats will result in high rates of denitrification, particularly when water column nitrate concentrations are high (Hoffmann et al., 2000). Most of the measurements of denitrification on tidal flats over the last 10 years have used the isotope pairing technique, and for consistency we only consider the rates measured using this method here. A more detailed summary of coastal sediment denitrification rates is available in Joye and Anderson (2008). As previously noted, water column nitrate concentration is a critical factor regulating denitrification rates. The highest denitrification rates measured on tidal flats were observed in Colne Estuary (UK), where activity in excess of 400 mmol N/m2/h was noted (Ogilvie et al., 1997). More generally denitrification rates on tidal flats in
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eutrophic estuaries fall in the range of 10–100 mmol N/m2/h (Ogilvie et al., 1997; Cabrita and Brotas, 2000; Dong et al., 2000). In more mesotrophic tidal flats denitrification rates are much lower, ranging from undetectable to generally <10 mmol N/m2/h (Trimmer et al., 2000; Cook et al., 2004a; Porubsky et al., 2008a). In the oligotrophic tidal flats of coastal Georgia (USA), in situ denitrification rates were low (1–10 mmol N/m2/h) but rates increased by more than 10-fold when the sediments were amended with nitrate to mimic a nutrient-rich runoff event (Porubsky et al., 2008a). The importance of denitrification on tidal flats in attenuating N loads to the coastal zone is quite variable. In the tidal flats of the Lower Great Ouse Estuary, between 2 and 56% (depending on season) of the nitrate load was attenuated by denitrification (Trimmer et al., 1998). This compares to an attenuation rate of 32–44% of the nitrate load estimated for the Colne Estuary (Ogilvie et al., 1997). In the Tagus Estuary, the tidal flats in the lower estuary attenuated 35% of the incoming DIN load compared to only 4% in the inner estuary (Cabrita and Brotas, 2000). Other nitrate reduction pathways include the anammox reaction and dissimilatory nitrate reduction to ammonium (DNRA). DNRA tends to occur in reducing sulfidic sediments (Brunet and Garcia-Gil, 1996; Christensen et al., 2000) as well as at temperature extremes (Kelly-Gerreyn et al., 2001). Such conditions often occur on tidal flats and substantial rates of DNRA have been inferred in these environments, with the proportion of nitrate reduced to ammonium varying between 5 and 54% (Kelly-Gerreyn et al., 2001; Porubsky et al., 2008a). The anammox reaction has never been directly measured on tidal flats; however, this process generally makes up only a small fraction of total denitrification rates in nearshore sediments (Dalsgaard and Thamdrup, 2002), and we therefore speculate that it is unlikely to be important in tidal flat sediments.
4.3. Exchange of dissolved nitrogen between the sediment and the water column The exchange of nitrogen between the sediment and the overlying water is modulated by the periodic exposure to the atmosphere. During exposure, ammonium accumulates in the porewater and may be released in a pulse into the overlying water during inundation. In permeable sediments, this process is enhanced by buoyancy-driven porewater exchange (Rocha, 1998; Falcao and Vale, 1995). At the edge of tidal sand flats, extremely high fluxes of ammonium may be observed on the low/falling tide driven by tidally induced porewater circulation phenomena described earlier (Figure 4; Billerbeck et al., 2006a). Such circulation patterns would also impact phosphorus and silicon cycling (discussed further later in the chapter). In situ flume experiments have also shown that ammonium exchange rates scale with current velocity, underscoring the importance of advective flushing in mediating fluxes upon tidal flat inundation (Asmus et al., 1998). Impacts of exposure have also been observed in cohesive tidal flat sediments, with an increased efflux of ammonium observed after inundation following a 6.5-h exposure period in the dark. Interestingly, the same exposure period in the light led to a large net uptake of ammonium upon inundation, highlighting how interactions
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N2
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Figure 4 The same general features of tidal flats apply to those formed on a permeable sandy substrate. A key difference, however, is that during inundation, the transport of dissolved N species between the sediment and the water column will be enhanced by advective porewater exchange (water flow through the sediments) (white arrows). This process may also enhance important reaction rates such as denitrification. During the falling tide and exposure, water drains out through the flat (black arrows) leading to locally extreme fluxes of solutes such as NH4þ at the edge of the flat.
between exposure and illumination can influence benthic fluxes in tidal flats (Thornton et al., 1999). In areas with high porewater nitrate concentrations, nitrate becomes depleted in the porewater through nitrate-reducing processes during exposure (Malcolm and Sivyer, 1997) and increased sediment uptake rates of nitrate have been observed after exposure (Feuillet-Girard et al., 1997).
4.4. Benthic microalgal N assimilation Nitrogen fluxes across the sediment–water interface of tidal flats often show a distinct diurnal variation, as a consequence of N assimilation by benthic microalgae at the sediment surface. During the light periods, sediment effluxes are greatly reduced or reversed and sediment uptake rates increase (Cabrita and Brotas, 2000). Measuring the nutrient assimilation rate of benthic microalgae is much more complicated than for pelagic algae, because they derive a large fraction of their nutritional requirement from nutrients produced within the sediment. Simply estimating the difference between light and dark fluxes will underestimate nutrient assimilation rates, because benthic microalgae are capable of assimilating nutrients for long periods in the dark. Estimates of N assimilation by benthic microalgae are often made based on measured rates of photosynthesis and then applying a C : N ratio measured for algal cells (often 8–10) (Dong et al., 2000). This stoichiometric approach is often used to estimate the relative amount of N assimilated by benthic microalgae so this value can be compared to the amount of N that is denitrified. Within the Colne Estuary, benthic microalgae assimilated a similar amount of N to that denitrified at a tidal flat site in the upper estuary; however, the amount of N
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assimilated was approximately two- to fourfold greater than that denitrified at a site in the lower estuary. Similarly, in the Tagus Estuary, twice as much N, on average, was assimilated by benthic microalgae than was denitrified (Cabrita and Brotas, 2000). We note, however, that the stoichiometric approach used for these calculations neglects the fact that benthic algae exude a large fraction of C fixed as dissolved organic carbon. Therefore, applying the C : N ratio measured in cells may not be appropriate, particularly in systems with low dissolved inorganic N concentrations. A more direct estimate of nutrient assimilation by benthic microalgae on a tidal flat was made by Cook et al., (2004a). They estimated total nutrient assimilation from nutrient uptake in cores, the upward flux of ammonium (based on porewater profiles), and measured rates of N2 fixation. When compared with rates of photosynthesis, the C : N uptake ratio was conservatively estimated at being between 17 and 52 during all the seasons except winter. The N cycle on tidal flats is highly dynamic with large changes in rates occurring temporally and spatially driven by an interaction between benthic microalgal production and the tidal cycles. Perhaps of greatest significance, tidal flats may be both substantial sources and sinks of fixed nitrogen. On Tomales Bay (CA, USA) intertidal flats, Joye and Paerl (1994) documented simultaneous N2 fixation and denitrification and concluded these sediments, inhabited by dense microbial mats of cyanobacteria, were a net source of N to the system over an annual cycle. However, given the general lack of simultaneous measurements of nitrogen fixation and denitrification, it is hard to comment on the role of tidal flats as net sources or sink of N in general. In highly eutrophied systems, tidal flats are likely to be an important sink for nitrate, whilst in more mesotrophic to oligotrophic systems nitrogen fixation may be an important local source of N (Joye and Paerl, 1994; Fulweiler et al., 2007). Tidal sand flats are important sites for the mineralization of particulate N trapped from the water column and subsequent release as ammonium, which is often a key nutrient limiting the productivity of phytoplankton in coastal waters (Ryther and Dunstan, 1971; Howarth et al., 1988a). The rates and controls on denitrification in tidal sand flats remain poorly characterized, and it will require combined methodological and modeling approaches to address this question. Simultaneous measurements of nitrogen fixation and denitrification are rare, and future studies should consider both these processes of appropriate spatial and temporal scales to evaluate the N balance of tidal flats.
5. P HOSPHORUS C YCLE Like nitrogen, phosphorus (P) is an essential nutrient element that is required for cellular biosynthesis and thus can limit rates of biological processes in the environment. Unlike nitrogen, sediment phosphorus is present in only one inorganic species/valence state, dissolved phosphate, which is typically present as HPO42– at circumneutral pH and seawater salinity (Ruttenberg, 2004). Phosphorus is present in sediments as a variety of organic compounds, including monophosphate esters, nucleotides and their derivatives, vitamins, phosphonates, and humic acids (Ruttenberg, 2004). Phosphorus is also present in a number of inorganic
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mineral forms. The microbially mediated component of phosphorus cycling is restricted to biological uptake and regeneration of inorganic phosphorus from organic forms because the phosphorus cycle has no dominant gaseous component. Continental weathering and subsequent sedimentation of particulate organic and inorganic phosphorus is the primary input of phosphorus to coastal sediments. Ensuing regeneration/release of dissolved inorganic phosphorus from the solid phase (both organic and inorganic forms) regulates P availability in sediments (Ruttenberg, 2004; Paytan and McLaughlin, 2007). A large portion of the solid phase P entering coastal sediments is inorganic phosphorus sorbed onto iron oxyhydroxides (Ruttenberg, 2004). Iron oxyhydroxides undergo reductive dissolution in anaerobic sediments by biotic and abiotic mechanisms thereby releasing dissolved inorganic phosphorus into the porewater where it is then available for biological consumption (Figure 5, adapted from Ruttenberg, 2004). Several excellent reviews of the phosphorus cycle are available, for example, see BenitezNelson (2000), Ruttenberg (2004), and Paytan and McLaughlin (2007). Phosphorus dynamics on tidal flats has received much less attention than N dynamics, possibly because P is not often considered the limiting nutrient in marine systems (Ryther and Dunstan, 1971). However, recent data illustrate the previously proposed idea (Smith, 1984) that widespread P limitation of biological activity can occur during certain seasons or within certain components of the ecosystem (Thingstad et al., 1998; Thingstad et al., 2005; Wu et al., 2000; Sundareshwar et al., 2003; Nicholson et al., 2006). Phosphorus availability can also regulate
Seawater
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Figure 5 Simplified schematic of reductive dissolution of Fe ^ P oxides in sediments. Reactions on the right side illustrate biological iron reduction resulting in release of reduced iron (Fe 2þ) and dissolved phosphate (PO43^). Reactions on the left show the abiotic reduction of iron oxyhydroxides coupled to hydrogen sulfide oxidation. This process also leads to release of Fe 2þ and PO43^. In oxic sediment, iron oxyhydroxides may reprecipitate.This gives rise to a complex internal cycle of dissolution and (re)precipitation that helps sequester both Fe and P in sediments as long as the sediments are overlain by an oxic layer. In anoxic sediments without benthic microalgae, both Fe2þ and PO43^ are released from the sediment.
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nitrogen fixation in some environments (Tyrrell, 1999; Ruttenberg, 2004; Baldwin and Williams, 2007). Variation in the position of the oxic–anoxic interface within sediments results from daily variations in benthic primary production rates or from hydrodynamic forcing. These changes in sediment oxygen distribution strongly regulate benthic P cycling. Sediments experiencing daily cycles of inundation and exposure, such as those of tidal flats, have a higher binding capacity for inorganic P because Fe oxides are recycled efficiently during periods of exposure (Figure 5; Baldwin, 1996; Mitchell and Baldwin, 1998; Lillebo et al., 2004; Mitchell et al., 2005). Iron oxyhydroxides effectively scavenge and retain inorganic P in sediments, which may ultimately sustain or stimulate the activity of benthic primary producers. Seasonality in inputs of particulate phosphorus versus organic phosphorus can further influence the impact of oxic–anoxic fluctuations on benthic P dynamics (Coelho et al., 2004). During periods of time with high particulate Fe–P inputs, movement of P between dissolved and inorganic pools can be largely driven by diurnal variations in oxygen availability (Figure 5). If, however, particulate Fe–P pools are limited, recycling of organic P may be a much more important mechanism for providing inorganic P to fuel biological processes. Benthic microalgal activity suppresses P release from sediments, either by direct P assimilation into biomass or by increased oxygen availability, which may stimulate the formation of Fe oxides and thus increase the inorganic P sorption capacity of the sediment (Figures 5 and 6; Joye et al., 1996, 2003; Sakamaki et al., 2006; Porubsky et al., 2008b). Typically, during submerged light incubation, P uptake from the overlying water rather than P release to the overlying water is observed Particulate P (organic and inorganic) Inundation
Exposure PO43–
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Figure 6 Phosphorus dynamics in tidal flat sediments. Benthic microalgal (gray circles) assimilation plays a significant role in limiting the P flux from the sediment and form sequestering water column P into sediments. As with N, P flows through benthic microalgae may dominate the P cycle. Organic P is mineralized releasing PO43^, which can be assimilated by benthic microalgae. Dissolved inorganic P is released from particulate inorganic P in anoxic sediments, providing another P source to biota. During exposure, exchange between the sediment and the water column is eliminated and enrichment in PO43^ may occur; this may drive P release from sediments upon inundation. The thickness of the arrows gives a rough indication of the relative magnitude of the P flows.
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(Joye et al., 1996; Magalha˜es et al., 2002; Porubsky et al., 2008b). Sediments inhabited by benthic microalgae appear to be able to take up P from water column and prevent sediment P release, even during nighttime (dark) incubation (Joye et al., 1996). Prolonged (>24 h) dark incubation in combination with anoxic conditions is required to initiate P release from sediments inhibited by benthic microalgae (Porubsky et al., 2008b). Anoxic conditions promote P release by stimulating either biological reduction of Fe oxyhydroxides or hydrogen sulfide-mediated reductive dissolution of Fe oxyhydroxides (Joye et al., 1996; Roden and Edmonds, 1997; Roden et al., 2000; Rozan et al., 2002; Benner et al., 2002). Release of inorganic P from Fe oxyhydroxides is often strongly correlated with sulfate reduction, and presumably hydrogen sulfide production, rates (Roden and Edmonds, 1997; Rozan et al., 2002). Anoxic bottom water further stimulates P release from sediments, presumably by increasing both biological reduction of Fe oxyhydroxides and hydrogen sulfide-mediated Fe oxyhydroxide reductive dissolution (Joye et al., 1996; Correll, 1998; Gunnars and Blomqvist, 2004). Additional research is needed to fully understand the mechanisms and controls of P dynamics on tidal flats.
6. S ILICON C YCLE Silicon dynamics on tidal flats is also poorly understood. Silicon (Si) is required by benthic and pelagic diatoms that take up dissolved silicate (silicic acid, Si(OH)4) and produce elaborate biogenic silica (SiO2) frustules (Davis, 1976). Silicon is supplied to coastal systems mainly via terrestrial runoff. Subsequent sedimentation of lithogenic inorganic primary and secondary Si minerals as well as biogenic Si provides Si to the benthos. Dissolved silicate is released through weathering and/or dissolution reactions of primary minerals or via dissolution (regeneration) of biogenic silica phases (e.g., opal). Biogenic silica dissolves about five times faster than lithogenic (mineral) silica (Ragueneau et al., 2006). In sediments, dissolved silicate accumulates in the pore water as accumulated diatom frustules dissolve (DeMaster, 1981) and then fluxes upward against the concentration gradient, where it may be consumed by benthic microalgae or may escape across the sediment–water interface (Figure 7). Groundwater may also supply the Si required by diatoms to some coastal systems; such inputs are important in some lakes (Hurley et al., 1985) but have not been documented so far in coastal waters. While benthic microalgae take advantage of the immediate sediment source of dissolved silicate, they also assimilate dissolved silicate from the water column (Sigman and Cahoon, 1997). Benthic microalgae often contain more silicon per chlorophyll a than their pelagic counterparts and are often highly silicified (Sigman and Cahoon, 1997; Ragueneau et al., 2006). Because diatoms cannot store silicon internally, they accumulate it immediately prior to cell division (Busby and Lewin, 1967); thus, benthic silicon fluxes may follow the cycle of diatom cell division as much or more so than that of light availability or photosynthesis (Ragueneau et al., 2006). The reason for this excess silica accumulation
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Lithogenic Si and biogenic Si Exposure
Inundation DSi
DSi BSi (LSi) (dissolution)
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Figure 7 Silica dynamics in tidal flat sediments. Benthic microalgal (gray circles) assimilation plays a significant role in regulating both the dissolved silicate (DSi) flux at the sediment ^ water interface and in sequestering Si as biogenic Si (BSi) in the sediments. As with N and P, benthic microalgal-mediated Si uptake may dominate the Si cycle. BSi dissolves releasing DSi, which can be assimilated or may flux from the sediment. Dissolved inorganic Si can also be released during the weathering of lithogenic Si but rates of such processes are poorly constrained. During exposure, exchange between the sediment and the water column is eliminated and an enrichment of DSi in the pore water may occur; this may drive DSi release from sediments upon inundation. The thickness of the arrows gives a rough indication of the relative magnitude of the Si flows.
in benthic diatoms is not known but may be related to low light availability in the benthos, it could assist benthic diatoms in resisting suspension by tidal currents or help them to return (sink) to sediments faster following a suspension event (Sigman and Cahoon, 1997). The availability of dissolved silicate can regulate primary production in pelagic waters (Officer and Ryther, 1980). Because benthic microalgae often regulate dissolved silicate fluxes in coastal systems, their activity can generate Si limitation in pelagic waters, leading to shifts in the dominant primary producer from silicified diatoms to nonsilicified phytoplankton (Sigman and Cahoon, 1997). Low water column dissolved silicate concentrations (<2 mM) drives silicate limitation of diatoms in some environments (Billen et al., 1991). Whether dissolved silicate availability limits production of benthic microalgae is unclear, but the limited data available suggest that Si limitation may be common during certain times of year (Sigman and Cahoon, 1997). Sequential nutrient limitation of benthic diatoms has been documented in some coastal waters (Figure 8; Joye et al., 2003; Porubsky et al., 2008b). In coastal Georgia (USA), intertidal flats occupied by benthic diatoms, nitrate pulses to the overlying water resulted in rapid nitrate uptake followed by silicate and/or phosphorus uptake from the overlying water. The pattern of nutrient uptake suggested that benthic diatoms are primarily nitrogen limited. However, once nitrogen demands were met, the diatoms become limited by either silicate (more often) or phosphorus (Joye et al., 2003; Porubsky et al., 2008b). The potential for nutrient limitation of benthic diatoms on tidal flats warrants further study.
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Figure 8 Sequential N, P, and Si limitation of benthic diatoms in coastal Georgia during a 3-day incubation of triplicate flow through cores. The arrows denote P followed by Si drawdown following nitrate addition. Methods described in Joye et al. (2003) and Porubsky et al. (2008a).
7. CONCLUDING REMARKS Temporal and spatial heterogeneity and strong forcing by tidal and wind dynamics generates extreme biogeochemical variability in intertidal flats and distinguishes these habitats from subtidal coastal sediments. Because of this inherent variability and a limited number of interdisciplinary studies, it is difficult to generalize regarding the biogeochemistry of intertidal flats. Nonetheless, it is clear that benthic
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intertidal flats play important roles in cycling of carbon, nitrogen, phosphorus, and silicon in coastal ecosystems. Physical (tidal and storm) forcing in tidal flat sediments modulates benthic exchange and the position of redox gradients in the sediments. Enhanced fluid exchange at low tide may stimulate N, P, and Si exchange and enhance bioactive nutrient flux to the overlying water column. Groundwater flow through tidal flat sediments is another potential means to elevate rates of fluid and materials exchange, but little data are available and the understanding groundwater dynamics in tidal flats is incomplete. More detailed studies of the interactions between biogeochemical cycles, such as documenting feedbacks between benthic primary production (e.g., C fixation and DOC release) and associated heterotrophic processes, regulation of benthic primary production by nutrient and light availability, and the balance between nitrogen fixation and denitrification, are needed in a broad array of intertidal flat habitats. Detailed studies examining the role of phosphorus and silicon availability, relative to nitrogen, in regulating biological dynamics on tidal flats is needed; at present, phosphorus and silicon cycling on tidal flats is poorly understood. Finally, more detailed studies of heterotrophic anaerobic microbial processes, such as sulfate reduction and methanogenesis, are required to complete our understanding of carbon dynamics on intertidal flats.
ACKNOWLEDGMENTS Support for preparation of this chapter was provided by the National Science Foundation’s Long Term Ecological Research Program (OCE 99-82133 to SBJ) and National Oceanic and Atmospheric Administration Sea Grant Programs in Georgia (award numbers NA06RG0029-R/WQ11 and R/WQ12A to SBJ) and South Carolina (award number NA960PO113 to SBJ).
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P A R T
I V
MARSHES AND SEAGRASSES
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C H A P T E R
1 3
P RODUCTIVITY AND B IOGEOCHEMICAL C YCLING IN S EAGRASS E COSYSTEMS Marianne Holmer
Contents 1. Introduction 1.1. Primary productivity 1.2. Fate of primary productivity – export and burial 2. Sediment Biogeochemistry – Modified by Seagrasses 2.1. Microscale effects 2.2. Nutrient cycling – importance of root uptake 3. Human Pressures and Effects on Biogeochemistry 4. Future Perspectives and Conclusions Acknowledgment References
377 378 380 383 391 392 393 395 396 396
1. INTRODUCTION Seagrass beds are widely distributed ecosystems in subtidal coastal zones around the world, except in Arctic and Antarctic, where ice cover limits their expansion (Green and Short, 2003; Larkum et al., 2006b). In the tropics high diversity is found with up to nine species present in the same meadow. Under temperate conditions fewer species or monospecific stands are more common (Green and Short, 2003; Larkum et al., 2006b) with some exceptions such as along the coasts of temperate Australia where high species diversity can be found (Carruthers et al., 2007). Compared to macroalgal communities, the diversity of seagrasses is relatively low, but among the 60 species found, there are large morphological differences extending from small species of a few centimeters in length to the meter-long leaves of Zostera marina and Enhalus acoroides (Figure 1). Seagrasses are anchored in the sediments by rhizomes and roots of similar lengths extending from few centimeters to several meters for Halophila sp. and E. acoroides, respectively (Duarte et al., 1998). Seagrass beds are some of the most productive ecosystems in the world, and they contribute significantly to carbon and nutrient cycling in the ocean (Duarte et al., 2004). In addition, they act as habitats for a large Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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Figure 1 Photo of an Enhalus acoroides meadow in Bolinao, Philippines.
number of animals and other plants, and play an important role as refuges for juvenile fish (Heck and Valentine, 2006). The rhizomes and roots of seagrasses stabilize sediments and to some extent protect the coastal zones against erosion. At the same time organic matter is buried in the meadows, which affects sediment biogeochemical conditions, including microbial activity and redox potential. Seagrass beds are threatened all over the world due to increasing human activity in coastal zones, where urbanization, tourism, and eutrophication are among the major threats (Duarte, 2002). This chapter reviews the importance of subtidal seagrass beds for coastal biogeochemistry by (1) describing their role as primary producers in the ocean, (2) describing the fate of the produced organic material within the meadows and outside, (3) elucidating the effects of seagrasses on sediment biogeochemistry from micro- to macroscale, (4) describing nutrient regeneration in the sediments, and finally (5) discussing the consequences of human pressures on seagrass–sediment interactions.
1.1. Primary productivity Many seagrass beds are very productive and belong to some of the most active marine ecosystems with similar productivity such as coral reefs (Duarte et al., 2004). In comparison with macroalgae, seagrasses take advantage of their ability to acquire nutrients from the sediments as well as the water column. Their light requirements, however, are higher because seagrass tissues support a larger fraction of respiratory organs (rhizomes and roots, Borum et al., 2006). Due to the large size differences in seagrasses, the productivity of the aboveground biomass varies as much as a factor of 500 between the least (Halophila ovalis) and the most productive species (Phyllospadix torryi) (Duarte and Chiscano, 1999). As for terrestrial ecosystems, belowground
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production is seldom measured, but is expected to reflect aboveground production. The few studies that are available show that the belowground production can account for half of the aboveground production (Duarte et al., 1998). Further studies may, however, reveal a large range in belowground production depending on the complexity of the species from large and complex to small and simple. In a terrestrial ecosystem, this could be equivalent to comparing a shrub (or even a tree for Posidonia species) versus a grass, where there are large differences in above- to belowground production. Also differences in nutrient availability in the water column or sediments may favor excessive growth of leaves or roots, respectively, as reported by Hillman et al. (1995) for Halophila ovalis in an estuary in Western Australia. Seagrasses are widely distributed from the equator to high latitudes with observations of Z. marina as far north as in Greenland (Rysgaard, pers comm.). The productivity of Z. marina has been measured as far north as at 63350 650N latitude, where it was about 62.5 mmol C/m2/day, only slightly lower compared to lower latitudes (Duarte et al., 2002). In temperate areas, seagrasses show a significant seasonal variation with maximum production and biomass during summer, and low biomass and low or insignificant net production during the winter (Olesen and Sand-Jensen, 1994; Risgaard-Petersen and Ottosen, 2000). In the subtropics and the tropics, seasonal studies show less variation although changes in light intensity and salinity may significantly affect productivity (Johnson and Johnstone, 1995). Seagrasses are generally the dominant source of the total primary productivity in seagrass beds, with epiphytes contributing up to 60% of the total (Hemminga and Duarte, 2000). There are many different types of epiphytes from small calcareous algae to large thread like macroalgae. A growing problem due to the increasing eutrophication in coastal zones is the invasion of seagrass beds by drifting macroalgae (Hauxwell et al., 2001; Kopecky and Dunton, 2006). While these growths may contribute significantly to net primary production, they negatively affect the seagrasses during decomposition due to increased community respiration that results in low dissolved oxygen concentrations in the seagrass meadows (KrauseJensen et al., 1999; Holmer and Nielsen, 2007). Whereas seagrass productivity is assessed from growth measurements, for example, by puncturing the leaves or by following changes in biomass over time, total community metabolism of seagrass beds can be estimated in situ by the use of benthic chambers. Here vertical fluxes of O2 and CO2 are measured and compared with bare sediments to assess the total community production and respiration. Such a seasonal study of Posidonia oceanica showed an autotrophic system (6 mol C/m2/ year) compared to the bare sediment where the annual net production was close to zero (Barron et al., 2006). This study also showed that CaCO3 precipitation and dissolution played a major role in the CO2 balance, emphasizing the role of carbonate sediments in community metabolism. As expected for temperate conditions, a large seasonal variation in net community production (NCP) was observed with high NCP in the warm summer months and net respiration in the winter months. It is therefore important to consider seasonal variations also in studies of biogeochemical conditions in seagrass beds. Seagrasses are generally considered to
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be carbon limited due to the low availability of CO2 at seawater pH where HCO 3 is the most available carbon source, and many seagrasses have developed mechanisms for its uptake (Larkum et al., 2006a). Due to the structural complexity of seagrasses compared to phytoplankton, seagrasses are considered to be more carbon limited, and in particular under low turbulence when the diffusive boundary layer around the leaves restricts CO2 uptake. As an example Thalassia testudinum appeared to be adapted to the calm conditions in a coral reef lagoon where low daytime CO2 concentrations that might otherwise limit productivity is compensated by possessing a carbon concentration mechanism in the leaves and an ability to take up HCO 3 . Much work is needed, however, for understanding the inorganic carbon dynamics at individual plant level and at community level in the seagrass beds (Larkum et al., 2006a). The release of oxygen from the roots may also affect the pools of inorganic carbon through enhanced aerobic respiration and reoxidation in the rhizosphere sediments (Burdige and Zimmerman, 2002). Aerobic respiration increases pools of CO2 and possible decreases in pH, which in carbonate sediments may result in significant carbonate dissolution (Hu and Burdige, 2007). Oxygen release may also enhance sulfide reoxidation and increase carbonate dissolution through reduced pH when sulfuric acid forms (Nielsen et al., unpublished). One advantage of the dissolution of carbonates is the associated release of phosphorus, which is considered a limiting nutrient for seagrasses in carbonate systems (Nielsen et al., 2006). Both diurnal and seasonal changes in water column inorganic carbon concentrations and sediment dynamics are of importance for the understanding of CO2 balance in seagrass meadows. Recent observations suggest much larger dynamics in the inorganic carbon cycling in seagrass meadows than originally expected (Barron et al., 2006). In spite of an estimated cover of only <2% of the ocean surface, seagrass beds together with other types of benthic vegetation contribute significantly to the oceanic carbon cycle because of their locally high productivity. In particular, their burial capacity was highlighted by Duarte et al. (2004), where vegetated habitats (including salt marshes and mangrove forests) account for about 50% of the total burial of carbon in the ocean. This is of major importance for the biogeochemical processes in sediments as they are generally limited by the availability of organic matter.
1.2. Fate of primary productivity – export and burial Only a limited part of the primary production from seagrasses enters directly into the grazing food chain, and with the loss of large herbivore megafauna, such as dugong and turtles, grazing has declined even further (Valentine and Duffy, 2006). An average value for grazing in seagrass beds is estimated to be 10%, but large variability exists as up to 50% of leaf production has been found to be grazed for small species such as Halophila sp. (Cebria´n and Duarte, 1998). The nongrazed biomass contributes to the support of an assemblage of species in the detritus food chain (Heck and Valentine, 2006). Seagrass detritus is relatively
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complex compared to other primary producers (macroalgae and phytoplankton) in the ocean. As such, seagrasses require the activity of specific enzymes such as cellulases and pectinases that limit their rate of decomposition, especially under low nutrient availability (Enrı´quez et al., 1993; Holmer and Olesen, 2002). Detritivores may contribute significantly to the decomposition process by fragmenting detritus, in particular in the tropics where detritivore diversity is high (Valentine and Duffy, 2006). Fragmentation may significantly enhance bacterial decomposition, which is a major importance for the turnover of organic matter in the sediments (Marba´ et al., 2006). Bacterial decomposition in the sediments is, however, also controlled by the complexity of the seagrass detritus, and is particularly slow for the structural and low-nutrient belowground parts compared to other types of marine detritus (Holmer and Olesen, 2001; Vichkovitten and Holmer, 2004). An important research question is to what extent is seagrass detritus decomposed within the meadows or is exported? Large piles of P. oceanica or Z. marina leaves on the beaches after autumn storms, for instance, indicate a significant export, and export to the oceans is considered to be important as well (Duarte et al., 2004). Of the few studies of belowground production, the high production that has been measured suggests a significant contribution to the pool of detritus in the sediments (Mateo et al., 2003). Also, organic material imported from other ecosystems due to increased sedimentation in seagrass beds and allochthonous production by epiphytes within beds both contribute to detritus in the surface sediments (Gacia et al., 2002). In comparison with bare sediments, detrital enrichment in seagrass sediments stimulates microbial activity and enhances the turnover and regeneration of nutrients (Holmer et al., 2003a; Duarte et al.; 2005, Marba` et al., 2006). The accumulation of organic matter in the sediments reflects the composition of detritus, where for instance rhizomes of P. oceanica may resist decomposition over millennia (Mateo et al., 2006). Duarte et al. (2004) have calculated that vegetated sediments contribute about half of the organic carbon burial in the coastal and global ocean at present. Vegetated coastal habitats are thus very important sinks of fixed carbon and may thus play a role in global CO2 balance and climate change. Burial of organic matter in seagrass beds depends on several factors such as the flux of organic matter to the sediment, sediment accumulation rate, sediment grain size and surface area, availability of oxygen, and the content of refractory organic components (Gacia et al., 2002; Marba` et al., 2006; Mateo et al., 2006). Current knowledge on burial is, however, limited, and most of this knowledge is from the peat-like accumulations as observed for P. oceanica where a matte of several meters thickness and thousands of years old can be found (Mateo et al., 2003). Also other species show a tendency of matte formation, and it is common to find intact fractions of rhizomes in the deeper parts of the sediments even for fast growing and less refractory species such as Cymodocea nodosa (Pe´rez et al., 2001). Both roots and rhizomes are difficult to decompose and factors such as low N and P content contribute to low degradability (Vichkovitten and Holmer, 2004; Moore and Fairweather, 2006). This is further accentuated by low concentrations of dissolved nutrients in seagrass sediments in general, leading to nutrient limited bacterial growth (Lopez et al., 1995).
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The use of bacterial markers such as 13C phospholipid analysis has led to interesting observations of the fate of organic matter in seagrass beds (Boschker et al., 1999, 2000; Holmer et al., 2003a; Bouillon et al., 2004). These techniques are useful because seagrasses have a relatively positive signal ( 13C –10‰) compared to other potential bacterial substrates. It is thus possible to identify the origin of carbon for bacteria from such sources as seagrass detritus, macroalgae, phytoplankton, or seston through the use of mixing models. The specific bacterial biomarkers used in this method are branched polar lipid-derived fatty acids (Boschker et al., 1999). This technique has been used in several seagrass beds from temperate and tropical areas, and a general picture has evolved whereby seagrass detritus plays an important role in oligotrophic conditions in the tropics (Holmer et al., 2001; Bouillon et al., 2004; Bouillon and Boschker, 2006). In contrast, nutrient-rich seagrass ecosystems are dominated by external sources such as macroalgae and phytoplankton. These substrates show higher degradability, which increases the microbial activity compared to less nutrient-enriched seagrass beds (Boschker et al., 2000; Holmer et al., 2003a). Seagrass detritus is exported from seagrass beds in dissolved and particulate forms. Seagrasses release dissolved organic carbon (DOC) directly from leaves (Ziegler and Benner, 1999) and high release rates have also been found during decomposition of detritus (Vichkovitten and Holmer, 2004). Seagrass sediments appear to be net DOC producers on an annual basis, which, if exported, may stimulate bacterial growth beyond their boundaries (Barron et al., 2004, 2006). Net DOC release is highest during the day and lowest at night. For T. testudinum, DOC release represented 10% of the net primary production (Ziegler and Benner, 1999) with a strong seasonal peak during the summer (Ziegler and Benner, 1999; Barron et al., 2006). Recent studies, however, suggest that seagrass sediments undergoing colonization are net sinks of DOC, but that they shift to net sources at maturity (Barron et al., 2004). Accumulation of seagrass detritus on the beaches is clear evidence of a large export of aboveground seagrass material (Mateo et al., 2003; Orr et al., 2005). The accumulation is in some locations so massive that it has been utilized for commercial purposes, such as construction of roofs and as fertilizers in Denmark (Sand-Jensen et al., 1994). Such accumulations are often considered as problems for sun lovers on the beaches in the popular tourist areas where the material is regularly removed. This may significantly impact biogeochemical cycles, as most of the nutrients bound in detritus is washed back to the sea, where it is again incorporated into the nutrient cycles (Mateo et al., 2003). Seagrass detritus is also important for the stability of beaches and hinders erosion during storms. Mechanical removal of seagrass detritus may thus increase erosion (Duarte, 2002). The extent of export of detritus from seagrass beds varies a lot, and estimates ranging from 1% to 80% of production have been proposed (Hemminga et al., 1991; Mateo et al., 2006). The large range is due to many different factors such as local currents, seafloor characteristics, wind exposure, and leaf morphology. A considerable fraction of detritus is considered to be exported to greater water depths where it contributes to the organic matter and nutrient pools in the sediments (Duarte et al., 2004).
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2. SEDIMENT BIOGEOCHEMISTRY – MODIFIED BY SEAGRASSES The higher pools of organic matter and the higher loading of organic matter to seagrass sediments increase the microbial activity compared to bare sediments. As a result the biogeochemical conditions in seagrass beds are often strongly modified compared to bare sediments (Figure 2, Marba` et al., 2006). At the same time the multifaceted belowground environment, with a mix of rhizomes and roots with large surface area (Duarte et al., 1998) where many different components can be released and taken up, increases the complexity of the sediments considerably. Roots show remarkably fast growth with elongation rates of several millimeters per day for Z. marina (Frederiksen and Glud, 2006), and as the exchange of compounds primarily occurs across the root tips, the movement of roots through the sediment is expected to have major influence on the local biogeochemistry (Figure 2), although the existing knowledge is limited (Frederiksen and Glud, 2006). Most of the seagrasses also show fast vegetative growth whereby the seagrasses spread into bare sediments
Import: DOM Nutrients
Export: Seagrass detritus Epiphytes Inorganic particles Carbonates
DOM, O2 Nutrients, O2
Nutrients, CO2
Concentration Low
High
0 O2 NO3
Aerobic respiration denitrification
DOM, O2 Nutrients
2
Manganese reduction
MnO2
Depth (cm)
FeOOH
Iron reduction
4 SO4
Sulfate reduction
POM +N +P
O2
6 FeOOH
DOM
–N –P
H2S
8
Direction of growth
10
Figure 2 Conceptual figure of macro- and microscale interactions between seagrasses and sediment biogeochemistry. IAN Symbol Libraries is acknowledged for the seagrass symbol.
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and modify them. As the shoot density increases to steady state or maturity with dense meadows, the biogeochemical conditions change as well (Pedersen et al., 1997; Pe´rez et al., 2001; Barron et al., 2004). Several microbial processes play important roles in marine sediments (Canfield et al., 1993). In the upper sediment layers, where oxygen penetrates, aerobic respiratory processes dominate the decomposition of organic matter. In unvegetated sediments, aerobic respiration often plays an important role, as organic matter is supplied from the water column and this fresh organic matter is mineralized rapidly in the oxic layer (Figure 2, Rasmussen and Jørgensen, 1992). The organic matter left behind in anoxic layers is often relatively refractory and results in much lower rates of decomposition (Canfield et al., 1993). In seagrass sediments, labile organic matter is supplied to the deeper layers as seagrass detritus and root exudates that potentially stimulate the microbial activity in these layers. At the same time oxygen is released by the roots, which may enhance the aerobic respiration in these layers as well (Figure 2). There is an overlap between the different respiration processes in marine sediments due to the patchy availability of electron acceptors and donors in the heterogenic sediment environment. In the oxic layers, nitrification usually takes place transforming ammonium to nitrate, which either diffuses to the water column or to the anoxic layers, where it is denitrified (Canfield et al., 1993). The denitrification process is succeeded by manganese and iron reduction, and the importance of both these processes for organic matter decomposition is controlled by the availability of oxidized pools of the components (Thamdrup, 2000). The magnitude of iron and manganese reduction is thus to a large extent determined by the reoxidation potential in the sediments, and a supply of oxygen from the roots is expected to increase the reoxidation of reduced manganese and iron and thus increase the pools of oxidized compounds available for the bacteria. Oxygen release from salt marsh plants has been found to stimulate the bacterial iron reduction significantly (Gribsholt et al., 2003), and preliminary results from seagrass (Z. marina) sediments show slightly stimulated rates near root surfaces (Holmer and Thamdrup, unpublished). In carbonate-rich sediments, which are important seagrass substrates in the subtropics and tropics, iron and manganese pools are often low due to low terrigenic influence (Berner, 1984). Here the terminal mineralization is dominated by sulfate reduction, as sulfate-reducing bacteria take advantage of the large pools of sulfate available in seawater (Holmer et al., 2001, 2003a). Sulfate reduction is thus often the dominant process in organic carbon oxidation after aerobic respiration. In organic-enriched seagrass sediments where the oxygen penetration is limited, sulfate reduction can dominate the mineralization (Holmer et al., 2003a). The belowground system of rhizomes and roots has several functions in addition to anchoring the seagrasses in the sediments. Carbohydrates reserves in rhizomes can be utilized when light availability is reduced during winter in temperate areas (Burke et al., 1996; Vichkovitten et al., 2007). Rhizomes also translocate nutrients between shoots, thus allowing seagrasses to maintain optimal growth conditions despite a heterogenous nutrient distribution in the sediments (Marba` et al., 2003).
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Rhizomes do not, however, interact with the sediment; rather nutrient exchange occurs primarily through the roots. Recent studies have shown that the exchange of compounds is further restricted to root tips as barriers against exchange develop in the older parts of the roots (Barnabas, 1994; Jensen et al., 2005). Most studies of exchange across roots have so far concentrated on the exchange of oxygen by the use of microsensors, while the knowledge of nutrient exchange processes is quite limited. Exchanges are primarily deduced from bulk analyses of sediment or observations of the distribution of bacteria in, on, and around the roots (Blaabjerg and Finster, 1998; Marba` et al., 2006). Bulk analyses of sediments show depletion of nutrient pools such as ammonium and phosphate (Risgaard-Petersen and Ottosen, 2000; Holmer et al., 2006a), and oxidation of sulfide and iron pools (Holmer et al., 2003a). Several stress studies, such as shading experiments, suggest a release of different organic compounds (sugars, ethanol, and amino acids) probably due to root anoxia and associated anaerobic respiration in roots (Smith et al., 1988; Pe´rez et al., 2007). Enhanced activity of sediment microorganisms during the day suggests a release of photosynthetic compounds to the sediment (Isaksen and Finster, 1996; Blaabjerg et al., 1998) and split-chamber experiments have shown significant release of DOC through the root system (Holmer et al., 2001). There is thus significant evidence that the release and uptake of compounds by the roots affect sediment biogeochemistry. Despite the fact that many seagrass meadows are autotrophic, the sediment compartment appears as net heterotrophic often with relatively high oxygen consumption rates and low oxygen penetration depth (Holmer et al., 2003a; Barron et al., 2004). A few early studies measuring nitrification and denitrification showed higher rates compared to unvegetated sediments (Caffrey and Kemp, 1990), whereas more recent studies report lower rates (Risgaard-Petersen et al., 1998; Ottosen et al., 1999; Welsh et al., 2000; Table 1). So far there are no reports on the importance of microbial iron reduction in seagrass sediments, but iron reduction may play a role in sediments with high pools of oxidized iron, such as terrigenic sediments, and for seagrass species with a high oxygen release capacity to speed up
Table 1 Rates of nitrification and denitrification in seagrass sediments Species
Location
Nitrification (mmol N/m2/day)
Denitrification (mmol N/m2/day)
Reference
Zostera marina
USA
410–2,481
71–209
Denmark
–
17
–
1.5
–
3–12
–
2–6
Caffrey and Kemp (1990) Risgaard-Petersen et al. (1998) Ottosen et al. (1999) de Wit et al. (2001) Welsh et al. (1996, 2000)
Zostera noltii
France
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the turnover of reduced iron into the oxidized pool and thus feed the bacteria. In contrast to the processes mentioned above, sulfate reduction has been examined in more detail during recent years, stimulated by observations of high sulfide pools in dieback areas (Carlson et al., 1994). Sulfate reduction is an important process in seagrass sediments, especially because seagrass sediments often have a high organic content that stimulates sulfate reduction rates (SSRs) in the deeper layers where other electron acceptors are exhausted (Holmer and Nielsen, 1997). There are, however, large variations within and between species (Table 2). Seasonal variations in water temperature and inputs of organic matter are important factors for the actual SRRs (Holmer and Laursen, 2002; Vichkovitten and Holmer, 2005; Holmer et al., 2003a). A study of P. oceanica at two locations, where one was characterized as pristine and the other as organicenriched (e.g., the site had restricted water circulation and a discharge of untreated sewage), SRRs were four times higher in the late spring with high water temperatures compared to winter with 8C lower temperatures (Figure 3). This temperature effect is somewhat higher than observed for bare sediments (Moeslund et al., 1994), but consistent with observations of enriched fish farm sediments, where the input of organic matter in combination with increasing temperatures resulted in very high SRR (Holmer and Kristensen, 1996). For Zostera species a positive relationship between diurnal changes in radiation and sulfate reduction has been found, suggesting a stimulation of SRR in the root zone due to release of organic matter during photosynthesis (Blaabjerg et al., 1998). At the same time correlations between shoot density and SRR (Holmer and Nielsen, 1997) and between the depth profile of root biomass and SRR (Vichkovitten and Holmer, 2004) have been found. This suggests that SSRs in Z. marina sediments are stimulated by a release of organic matter from the seagrass, although it cannot be excluded that the relationship with shoot density also is due to increased sedimentation of external sources, and the relationship with root biomass is caused by increased pools of root detritus. The role of external sources was discussed earlier in the isotopic studies. For species like P. oceanica, a negative correlation between the root biomass and SRR has, however, been found, suggesting an inhibition of the microbial rates at increasing seagrass abundance. This is probably due to the release of oxygen from the roots, which increases the oxidation of the sediments with increasing root biomass, and thus favors aerobic respiration rather than sulfate reduction (Holmer et al., 2003a). P. oceanica has an extensive air–tissue system, but oxygen release from the roots has not been demonstrated. High concentrations of reduced sulfides have been reported for many seagrass sediments (Table 2) where particulate pools of acid-volatile sulfides (AVSs) and chromium-reducible sulfur (CRS) are higher compared to those of unvegetated sediments. The dissolved sulfide pools do not, however, necessarily reflect the high SRR, as unvegetated sites may show much higher accumulations of dissolved sulfides (Thode-Andersen and Jørgensen, 1989; Moeslund et al., 1994). In unimpacted seagrass sediments, sulfide pools ranging from below detection limits (<1 mM) and up to 50–100 mM have been measured (Holmer and Nielsen, 1997; Lee and Dunton, 2000; Frederiksen et al., 2006; Calleja et al., 2007). In dieback areas of T. testudinum meadows, however, concentrations up to several millimolars
Table 2 SRRs and pools of reduced sulfides in seagrass sediments
Cymodocea nodosa Cymodocea rotundata Cymodocea serrulata Enhalus acoroides Halodule beaudetti Halophila ovalis Posidonia oceanica
Thalassia hemprichii Thalassia testudinum Zostera marina
Zostera noltii
Location Spain Thailand Thailand Thailand Thailand Australia Thailand Thailand Jamaica Thailand Thailand Spain Spain Spain Spain Italy Greece Cyprus Spain Thailand Thailand Florida Denmark Denmark Netherlands Denmark Denmark Denmark Denmark Denmark France
SRR (mmol/m2/day) 21 6.4 110 – – 90 80 – 34 80–120 – 3–12 – 10–15 8–24 19–42 6–19 17–27 – 2.0 40–130 – 25–59.1 13.2–29.6 7.4 12–70 19–41 – 16–40 38–110 13.7 29
– – 0.3 <0.5 <0.5 – 0.5 <0.5 – 0.1 <0.5 – – 0–0.16 0.04–0.12 0–0.40 0.04–0.08 0.05–0.07 0.37–2.29 – 0.1 – – – – – – 0.00–0.63 0–1.84 0.03–0.20 – –
CRS (mol/m2) a
12.80 9.50a 3.0 1.3–7.4 2.2–6.0 – 11.5 1.7–8.2 – 3.5–4.0 1.9–7.8 0.16–3.82a 0.05–1.81a 0.86–11.98 15.57–23.83 11.72–17.54 0.70–3.21 9.23–20.59 12.24–17.38 7.50a 3.2–4.2 0–40b 0.07–0.20 – – – 0.07–7.0a 4.56–8.19 0.76–40.3 0.46–0.59 – –
Reference Holmer et al. (2004) Holmer et al. (2001) Holmer et al. (2006b) Delefosse (2007) Delefosse (2007) Pollard and Moriarty (1991) Holmer et al. (2006b) Delefosse (2007) Blackburn et al. (1994) Holmer et al. (2006b) Delefosse (2007) Holmer et al. (2004) Calleja et al. (2007) Holmer (unpublished) Frederiksen et al. (2007) Frederiksen et al. (2007) Frederiksen et al. (2007) Frederiksen et al. (2007) Pe´rez et al. (2007) Holmer et al. (2001) Holmer et al. (2006b) Chambers et al. (2001) Holmer and Nielsen (1997) Boschker et al. (2000) Boschker et al. (2000) Blaabjerg et al. (1998) Holmer and Laursen (2002) Frederiksen (2005) Frederiksen et al. (2006) Vinther et al. (in press) Welsh et al., 1996 (2000) Isaksen and Finster (1996)
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Values are depth-integrated over the upper 0–10 cm. a Given as total reducible sulfur pools (TRS = AVS þ CRS). b Given in mmol/g. AVSs, acid-volatile sulfides; CRS, chromium-reducible sulfur; SSRs, sulfate reduction rates.
AVS (mol/m2)
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Species
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22
15
SRR (mmol/m2/day)
20 10 18
16
5
Water temperature (°C)
Impacted Pristine
14 0 February
April
June
October
Figure 3 Seasonal variation in depth-integrated sulfate reduction rates in two different Posidonia oceanica meadows, one located at a pristine site (Pristine) and one in a human impacted site (Impacted).The water temperature is shown in the line graph. Modified from Holmer et al. (2003a).
can be found (Carlson et al., 1994). The low pools in unimpacted sediments suggest a rapid reoxidation to oxidized forms or a precipitation in the particulate sulfide pools. It is important to emphasize, however, that even low concentrations of dissolved sulfides in the pore water imply that the sediments are anoxic. A negative correlation between dissolved sulfides and net population growth of P. oceanica was observed with threshold values as low as 10 mM causing reduced growth rates (Calleja et al., 2007). In contrast to the particulate sulfide pools, dissolved pools show large diurnal and seasonal variation in seagrass sediments (Lee and Dunton, 2000; Pedersen et al., 2004; Borum et al., 2005; Frederiksen et al., 2006). Borum et al. (2005) found low pools of dissolved sulfide in T. testudinum meadows during the day due to release of oxygen from the plants, which shows the large potential of this species for reoxidation of the sediments. Frederiksen et al. (2006) found high pools of dissolved sulfide in the summer months for Z. marina, indicating that this species is not able to oxidize the sediment despite a high capacity for oxygen release from the roots. Accumulation of sulfide was coincident with an unexpected decline in carbohydrate content of the belowround biomass. This suggests that plants were not able to maintain their carbohydrate reserves during the summer period but instead utilized the reserves probably due to root anoxia (Vichkovitten et al., 2007). Similarly P. oceanica showed a decline in nonstructural carbohydrate pools as a result of sediment organic matter enrichment followed by the development of high sedimentary sulfide pools (Pe´rez et al., 2007). The effect of root anoxia on whole plant carbon balance, however, remains to be quantified. In a number of recent studies, stable sulfur isotopes have been used to track a potential invasion of sulfides into seagrasses. The principle behind these measurements is to use the large difference in the 34S signal between seawater sulfate
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( þ20‰) and sediment sulfides (–15‰ to –30‰). Sulfate-reducing bacteria discriminate toward the lighter isotope and sediment sulfides can be up 50‰ lower than seawater sulfate. The stable isotopic signatures of sulfur are reported in standard delta notation (units per mil, ‰), calculated as follows: Rsample 34
S= 1 1; 000 Rstandard where Rsample = 34S/32S in the sample and Rstandard is the isotopic composition of the troilite standard from the Canyon Diablo meteorite that is used as a zero point for expression of sulfur isotopes. Sulfide inside plants has been observed in T. testudinum (Borum et al., 2005) and Z. marina (Pedersen et al., 2004) in both unimpacted sediments and in dieback areas. Sulfide invasion into the air–tissue system of the plants is expected to be passive. Here it is oxidized relatively fast, when it meets oxygen (Pedersen et al., 2004). While there are several possible reoxidation products such as sulfate and thiosulfate, it was elemental sulfur that accumulated in P. oceanica, Thalassia sp., and Z. marina (Holmer et al., 2005b, 2006b; Frederiksen et al., 2006, 2007; Koch et al., 2007). One objective of using the stable isotope technique is early identification of the health of seagrass beds before dieback or mortality occur. This is based on observations of negative effects of sulfide on parameters such as photosynthetic capacity, growth rates, and population dynamics (Goodman et al., 1995; Holmer and Bondgaard, 2001; Calleja et al., 2007). P. oceanica, growing near fish farms in the Mediterranean, was clearly affected by the release of waste products from the farms (Holmer et al., 2003b; Frederiksen et al., 2007). In one example, mortality of P. oceanica correlated with sedimentation of organic matter along transects radiating from the farms (DiazAlmela pers. comm.). High SSRs were measured in the sediments and dissolved sulfides were detected in the pore waters at the stations near the fish farms (Holmer and Frederiksen, 2007; Frederiksen et al., 2008). The 34S signals in the plants were clearly lower near the net cages (Figure 4) and a general pattern correlated with the rates of sedimentation from the farms could be observed across the Mediterranean (Frederiksen et al., 2007). Roots showed the lowest values, indicating that sulfide invaded through the roots and was spreading from roots to the rest of the plant. Also the rhizomes showed low values, while the leaves overall reflected the seawater sulfate signal. This indicates that sulfide remained in the belowground parts where it was oxidized or incorporated into organic sulfur compounds (Frederiksen et al., 2007). The 34S signal was negatively correlated with the content of total sulfur (TS) in the plants, where the 34S signal decreased with increasing TS. Similar observations have been made for Z. marina (Frederiksen et al., 2006) and indicate that sulfide is incorporated into the sulfur components of the plants. For Z. marina high accumulations of elemental sulfur were found in the plants (Holmer et al., 2005b; Frederiksen et al., 2006), while only low accumulations were found for P. oceanica (Frederiksen et al., 2007). The authors mention that the low elemental sulfur in this study may have been due to an analytical error or due to incorporation of sulfide into organic sulfur components.
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Plant δ 34S (‰)
Spain
Greece
Cyprus
25
25
25
20
20
20
20
15
15
15
15
10
10
10
10
5
5
5
5
Leaves Rhizomes Roots
0 1
Fsulfide(%)
Italy
25
2
0
0 1
3
2
3
0 1
2
3
1
50
50
50
50
40
40
40
40
30
30
30
30
20
20
20
20
10
10
10
10
0
0
0
1
2
3
1
2
3
0 1
3
2
2
3
1
2
3
Station
Figure 4 Changes in Posidonia oceanica 34S signal and contribution of sediment sulfides (Fsulfide) along transects from four different fish farms in the Mediterranean. Station 1 is located 5^25 m from the net cages (in Cyprus 300 m), Station 2 is located 25^50 m from the net cages (in Cyprus 400 m), and Station 3 is located 1,000^1,200 m from the net cages, where Station 3 is considered as a control site. Modified from Frederiksen et al. (2007).
It is possible to estimate the contribution of sediment sulfides to the TS content of the seagrasses by calculating Fsulfide, which is a ratio between the 34S signal in the plant tissue and the 34S signal of the potential sulfur sources: Fsulfide =
34 Stissue 34 Ssulfate
34 Ssulfide 34 Ssulfate
34Stissue is the value measured in the leaf, rhizome, or root, 34Ssulfate is the seawater value, and 34Ssulfide is the sediment sulfide value (derived from pore water sulfides or AVS) (Frederiksen et al., 2006). In the study of fish farms, Fsulfide showed a high invasion of sulfides in roots (up to 43%) and rhizomes (up to 28%) at the two stations near the farms, while sulfide invasion into the leaves only was found on one occasion (in Cyprus, data not shown). Surprisingly, the control plants also showed an invasion of sulfide with about 20%. This suggests that the control plants were influenced by sulfide, which may be due to natural causes or because the plants at the control sites also were impacted by sulfide. P. oceanica was in decline at three out of the four fish farms, and the sulfide invasion into the plants may thus be a sign of negative impact (Frederiksen et al., 2007). An experimental study of Z. marina exposed to various thicknesses of mats of filamentous algae also showed higher accumulation of sulfides in the leaves (up to 46%) compared to untreated plants (20%) (Holmer and Nielsen, 2007). The accumulations were directly correlated with the pools of sulfides in the sediments, and the negative effects on the relative growth rates of the plants were
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directly correlated with the Fsulfide. This suggests a direct relationship between sulfide invasion into the lower part of the shoot and plant performance when the plants are exposed to anoxia. Meristem anoxia in Z. marina resulted in plant mortality; sulfide toxicity was suggested as a contributing factor (Greve et al., 2003).
2.1. Microscale effects Most of the above-mentioned biogeochemical results are based on analysis of intact sediment cores and depth profiles with relatively small resolution (centimeterscale). The approach is justified because of the quite compact networks of belowground tissues are assumed to have pervasive effects on bulk sediment biogeochemistry. New studies, however, show that the most dynamic parts of seagrass belowground tissues are the root tips (Figure 5) and that roots of species like Z. marina grow very fast (8.7 + 1.7 mm/day, Frederiksen and Glud, 2006). Roots become lignified due to formation of Casparian Bands along the longitudinal interior, leaving only the roots tips without a barrier for exchange of compounds (Barnabas, 1994). As the root tip penetrates the sediment, the biogeochemical conditions have been found to be highly dynamic in the newly exploited sediment (Frederiksen and Glud, 2006). Studies with microelectrodes and optodes have shown that oxygen diffuses into the sediment in a radius of 10 mm from the root tip (Figure 5) with no sulfide present in this zone. It can be expected that DOC is released and nutrients are taken up in the same zone (Figure 2). The plant benefits from access to unexplored sediment where nutrient pools most likely are higher than in the sediment above. The disadvantage is that penetrating roots are exposed to higher sulfide pools and thereby invasion of sulfides. This can be compensated by
176.3 152.5 128.8 105 81.25 57.5
O2 concentration (µmol/l)
200
180 160 140 120 100 80
Tip
60 40 20 0
33.75 10
0
2
4 6 8 Distance (mm)
10
Figure 5 Oxygen distribution around roots from a Z. marina plant at light intensity above photosynthetic saturation (500 mmol photons/m2/s). The root distribution is shown in the black-and-white image (left), where roots appear as white lines (marked by black dots). Planar optode images of oxygen distribution are shown in the middle and peak oxygen concentrations along the root length axis starting from the root tip are shown to the right. Modified from Frederiksen and Glud (2006).
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an efficient oxidation of sulfide by oxygen released from the roots (Frederiksen and Glud, 2006). The importance of oxygen release from roots in comparison to total oxygen supply to the sediments has not been completely resolved (Duarte et al., 1998). Jensen et al. (2005) and Frederiksen and Glud (2006) estimated that the release of oxygen from the roots contribute a maximum of 12% to the diffusive oxygen uptake in the sediment based on the root surfaces. This indicates a limited effect of oxygen release from roots on benthic sediment respiration. These results support findings that oxygen-related processes play a minor role in stimulating nitrification and coupled nitrification–denitrification (Risgaard-Petersen and Ottosen, 2000). If oxygen production is short-lived for seagrasses in general, as shown by Frederiksen and Glud (2006) for Z. marina, nitrifying and denitrifying bacteria may not be able to respond by increased growth because oxygen supplies would be first consumed and sulfide produced leading to an unfavorable environments for this type of bacteria. It is thus most likely that oxygen is rapidly consumed for sulfide oxidation (in minutes) compared to bacterial growth (in hours). Despite the unfavorable conditions for nitrifiers and denitrifiers, the anaerobic bacterial activity has been found to be stimulated in the area around the root tips by the release of DOC from roots (Nielsen et al., 2001), a process that is supported by the generally high rates of sulfate reduction found in seagrass sediments (Marba´ et al., 2006). These hot spots of bacterial activity and nutrient regeneration may be crucial to assure plant survival and nutrient uptake, and the microscale approach should be explored further for understanding the availability of nutrients and nutrient uptake in seagrass sediments.
2.2. Nutrient cycling – importance of root uptake Seagrasses contribute in many ways to the nutrient cycles in coastal zones, not least due to retention of nutrients in the vegetation during the growing season. Nitrogen and phosphorus, for example, are taken up from the water column and the sediments and retained in above- and belowground biomass (Risgaard-Petersen and Ottosen, 2000; Holmer et al., 2006a). The sedimentation of organic matter from allochthonous and autochthonous sources in seagrass beds also retains nutrients when the material is mineralized and incorporated into the vegetation as discussed above. Access of nutrients from the sediments is considered as an important adaptation, which ensures a wide distribution of seagrasses in particular in oligotrophic waters (Hemminga, 1998). The existing knowledge on uptake mechanisms and nutrient availability in sediments is, however, limited. Nutrient uptake occurs through the root tips, and the nutrient availability will to a large extent be determined from the rates of mineralization in the sediments that in turn is determined from an array of factors such as organic matter pools, temperature, and physical conditions of the sediments (Canfield et al., 1993). The importance of sediments as a nutrient source is often determined through indirect studies, such as short- and long-term nutrient enrichment of sediments followed by measurement of growth and nutrient concentrations in the plant tissues (Erftemeijer et al., 1994; Ceccherelli and Cinelli, 1999). Some studies have focused on uptake kinetics by
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incubations of plants in split chambers or in sediments with flow-through of nutrient solutions (Gras et al., 2003; Nielsen et al., 2006). The uptake of nutrients is primarily determined from the nutrient availability in the water column and the sediments; sources may shift from water to sediment depending on the availability. Leaves, however, have high affinity for nitrate and roots for ammonium (Touchette and Burkholder, 2000). The affinity for phosphate uptake is lower than that reported for ammonium. Phosphate can be taken up at very low concentrations, however, which is an advantage in carbonate sediments where P is bound in the carbonates (Nielsen et al., 2006). Roots are thus able to take up nutrients over a large range of concentrations, an advantage in sediments where pore water pools may vary significantly over the season. Work by Holmer et al. (2001) in intertidal seagrass beds outside a mangrove in Thailand evaluated the capacity of the sediments in supplying nutrients to the plants. The sediment showed that high pools of nutrients and the rates of mineralization were significantly higher than those in bare sediments. The anaerobic mineralization was able to support from 6% and up to 81% of the plant nutrient uptake, most importantly for phosphorus. Because plant nutrient contents were relatively high and did not indicate nutrient limitation, remaining nutrients can be presumed to have been derived from the water column. In carbonate-rich sediments common near coral reefs, P limitation of seagrasses is generally observed (Jensen et al., 1998). This is for instance the case for T. testudinum in Florida Bay (Gras et al., 2003; Nielsen et al., 2006). While the water column has very low P concentrations, large pools of P are bound in the sediments, sufficient to support the growth of the seagrasses for decades (Jensen et al., 1998; Nielsen et al., 2006). There is, however, the problem that P is bound tightly to the calcium fraction and thus is not directly available. Pore water pools of P are often quite low, and a standing question is, if the P in the sediments is available for the plants at all? While it has been hypothesized that organic acids or bicarbonate released from plants are able to dissolve calcium-bound pools (Jensen et al., 1998; Burdige and Zimmerman, 2002; Hu and Burdige, 2007), the acidification is not likely to be sufficient to dissolve adequate amounts of P. Ongoing studies investigate if the interaction between oxygen released from the roots can oxidize sulfide to sulfuric acid, which could have the capacity to contribute to the dissolution, even in well buffered carbonate sediments. P limitation is a general phenomenon in the tropics, and while this mechanism may be widespread elsewhere, it is less importance in iron-rich sediments where iron-bound pools of P are more easily accessed by plants (Holmer et al., 2006a).
3. HUMAN PRESSURES AND E FFECTS ON BIOGEOCHEMISTRY Seagrasses are increasingly exposed to many kinds of human pressures, in part because of their coastal distribution (Duarte, 2002; Orth et al., 2006). Among these effects are runoff of nutrients from agriculture, wastewater discharge, siltation due to soil erosion, tourism development, artificially amended beaches, desalinization plants, extraction of sand, expansion of harbors, etc. Regardless of the specific type
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of impacts, a common feature for many is reduced light availability either as a result of more suspended particles or enhanced phytoplankton. The Baltic Sea and the inner Danish coastal waters have drastically reduced depth limits over the last 100 years with a dramatic reduction in the 1970s due to major increase in the use of artificial fertilizers in the agricultural sector (Nielsen et al., 2002; Frederiksen et al., 2004). Many seagrasses require about 11% of incident light in part to support the maintenance of belowground nonphotosynthetic biomass where nutrients are supplied and in particular oxygen is transported to oxidize the sediments (Duarte, 1991; Dennison et al., 1993; Ralph et al., 2007). Due to these large losses of seagrasses, shading experiments have been undertaken for a wide selection of seagrasses. Where sediment biogeochemical conditions have been examined at the same time, declining redox potentials and increasing pools of sulfides have generally been found (Enrı´quez et al., 2001; Holmer and Laursen, 2002). These findings underline the ability of the seagrasses to oxidize the sediments, and as species such as P. oceanica are very sensitive to the presence of sulfide in the pore waters, shading may increase the mortality. Other species such as Z. marina are more tolerant of reduced conditions, but sediment conditions contribute to lightrelated depth limitations that may further compound reduced sediment conditions (Greve and Krause-Jensen, 2005; Ralph et al., 2007). Eutrophication and increased loading of particulates from land can contribute to organic enrichment of sediments. While low levels of enrichment may stimulate the growth of seagrasses, high rates of loading can enhance sulfate reduction with accumulation of sulfide in pore waters. Several experiments have added organic matter to the sediment or have injected sulfide solutions (Holmer et al., 2005b; Kilminster et al., 2006; Koch et al., 2007). The different seagrasses show large variation in their response to increased sulfide concentrations. Some species are more tolerant probably due to a large capacity for oxidation of invading sulfides. This has been found in Z. marina and T. testudinum, where treatments led to high accumulations of elemental sulfur in the plants. It was not until a combination of different factors, such as organic matter addition and reduced light conditions, that reduced growth rates and increased mortality were observed for Z. marina (Holmer et al., 2005b). T. testudinum showed, in an experiment with multiple stressors using organic matter addition and increased temperature, reduced performance only at high temperatures (>33C) and high sulfide concentrations, whereas Halodule wrightii had higher thermal resistance but only in the absence of sulfide (Koch et al., 2007). In carbonate-rich sediments, an organic enrichment may show even more dramatic impacts compared to terrigenous sediments, as the buffer capacity toward sulfide is reduced due to the low iron content. In terrigenous sediments iron potentially buffers sulfide by precipitation of iron sulfides, which reduces the pools of dissolved sulfides in the pore water. In carbonate sediments the iron pools are rapidly exhausted, as the degree of pyritization already is close to maximum (Berner, 1984), and the sediments have limited capacity for binding sulfides. Sulfide may accumulate in the pore water provided that it is not reoxidized by the release of oxygen from the roots. The sediment oxygen demand thus increases significantly, which increases the risk of root anoxia, and by accumulation of
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sulfide, the possibility of sulfide invasion increase as well. A second important aspect is that the availability of iron can be expected to decrease if it is bound irreversibly to iron sulfides, which may reduce growth. Seagrasses in carbonate-rich sediments generally have low iron contents, and iron itself can be a limiting nutrient (Duarte et al., 1995). In an experiment with organic enriched carbonate sediments due to loading with sewage and increased phytoplankton sedimentation, iron addition to the sediment resulted in higher content and increased growth rates of the plants (Holmer et al., 2005a). In a 2-year experiment of iron injections, decline in seagrasses in the organic enriched sediments was reversed (Marba` et al., 2007).
4. FUTURE P ERSPECTIVES AND C ONCLUSIONS Seagrass ecosystems are facing a global crisis due to human pressures in the coastal zone (Orth et al., 2006) and this chapter examines several examples of the crucial role of the biogeochemistry in seagrass sediments for the observed seagrass declines. The most documented impact is the reduction of the sediments either as a result of less oxidation of the sediments due to reduced seagrass performance (lower photosynthetic activity, lower shoot density) or due to increased oxygen demand of the sediments from increased organic enrichment and increased microbial activity. There are, however, many known and unknown factors that contribute to seagrass decline and several show cumulative effects, such as shading and organic enrichment of sediments, or organic enrichment and iron limitation. Seagrass scenarios under these conditions can be very difficult to predict considering the limited knowledge on multiple stressors now available. In particular sediment biogeochemistry has received limited attention and is almost nonexistent for subtropical and tropical seagrasses with a few exceptions. This may be very critical for the future distribution of seagrasses as the human pressures are considered to grow more rapidly in coastal zones in the tropics. One other important aspect of seagrass decline is the consequences for coastal and oceanic ecosystems due to changes in coastal sediment biogeochemistry associated with the loss of seagrass meadows. Some information is available from the wasting disease for Z. marina on the northern hemisphere during the 1930s, where major changes in coastal erosion and loss of fisheries habitats followed the loss of seagrass meadows. These ecosystems had not fully recovered before new seagrass losses occurred due to eutrophication (Frederiksen et al., 2004). Currently, losses of other more slow-growing species, such as P. oceanica in the Mediterranean (Marba´ et al., 2005), may be even more dramatic due to the slow regrowth of this particular species. It is hard to predict whether reduced nutrient loadings will allow recolonization of Danish coastal zones by Z. marina because many concurrent factors appear to be involved. In some areas the organic enrichment of the sediments during eutrophication in the 1970s and 1980s led to repeated oxygen depletion events each year and thus prevented recolonization (Frederiksen et al., 2004). In other areas, a faster colonization by the blue mussel Mytilus edulis hampered the seagrass recolonization process (Vinther et al., 2008). In both examples sediment biogeochemistry played
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a central role due to high oxygen demand and high pools of sulfides, factors that limited the growth performance of seeds and seedlings during the recolonization. To be able to understand, manage, and conserve seagrass meadows it is thus necessary to include studies of the biogeochemistry of coastal sediments and to expand these studies to a much wider geographical extent and for more seagrass species. One important objective is to determine ecological thresholds for seagrass performance in ecosystems under human pressures so that further losses of seagrasses can be avoided.
ACKNOWLEDGMENT M.H. was supported by the Danish Natural Science Foundation grants no. 272-07-0031 and 272-05-0408.
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T IDAL S ALT M ARSHES : G EOMORPHOLOGY AND S EDIMENTOLOGY John R.L. Allen
Contents 1. 2. 3. 4.
Introduction Geographical Distribution Why Salt Marshes Exist? Geomorphology 4.1. Marsh evolution versus inheritance 4.2. Marsh edges and coastal change 4.3. Marsh terraces 4.4. Channels, creeks, and gullies 4.5. Creek networks 5. Morphodynamics 5.1. Tidal regime 5.2. Sediment sources and supply 5.3. Channelized flows 5.4. Platform flows 5.5. Accretion, compaction, and sea level change 6. Sedimentology 6.1. Grain size 6.2. Tidal bedding 6.3. Lithostratigraphic architecture 7. Concluding Discussion References
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1. INTRODUCTION Where the tide is a dominant coastal agent (Anthony and Orford, 2002), some or all of the sedimentary surface in the uppermost part of the tidal frame is underlain by muddy sediment and, under appropriate climatic conditions, colonized by salttolerant herbs, grasses, and small shrubs. These salt marshes consist typically of two morphologically and hydraulically distinct environments (Allen, 2000a). A vegetated, silt-trapping platform above the level of mean high water of neap tides is seen to be Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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Dryland
High marsh
Low marsh
Sandflat
Figure 1 Air photograph (2.0 1.8 km, north toward bottom) of a two-phase (terraced) salt marsh, Stiffkey (Norfolk), North Sea coast, eastern England. The mature high marsh (light tone) shows well-developed tidal drainage networks composed of strongly meandering elements. Drainage on the much younger low marsh (dark tone), a few decimeters lower, is less developed. A low sand chenier (lightest tone) tops the outer edge of the older marsh to the east of the main channel and in the extreme west. A broad sandflat lies to seaward.
dissected by networks of tidal waterways varying from several meters deep and tens to hundreds of meters wide to blind-ended, decimeter-scale features (Figure 1). Tidal salt marshes play vital roles. Hydraulically, they help to protect the coastal zone, damping waves, storing surge waters, and trapping fine sediment. Ecologically, salt marshes offer rich habitats for invertebrates, birds and wild herbivores, as well as discharging nutrients. Culturally, prehistoric humans exploited them for such largely seasonal activities as hunting, wild-fowling, herding and grazing, fishing, and reed-gathering. Their drainage networks have long been important means of communication by water. Over the last two millennia, many salt marshes, especially in northwest Europe, have been embanked, permanently settled, cultivated, and exploited industrially or commercially, with often damaging consequences. Elsewhere, navigation channels have been cut through marshes. Many that remain active are now sinks for anthropogenic contaminants with the potential to be released on coastal change. This chapter attempts to identify key understandings and uncertainties regarding salt marshes and to suggest future needs. It deliberately takes the view that the
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“one-time” studies which dominate the literature are insufficient to understand these critical environments. Salt marshes arose often during the Holocene, an epoch of changing and fluctuating climate and sea level. Change on decadal to millennial scales continues. As complex dynamical systems functioning in the flow of time, these marshes have reacted to change in ways that are significant for the prediction of future responses. Envirostratigraphic studies and modeling therefore have essential roles.
2. GEOGRAPHICAL D ISTRIBUTION Salt marshes occur almost worldwide (Whigham et al., 1993). Reviewing Chapman’s (1960) botanical classification, Adam (1990) recognized six floral associations: Arctic, Boreal, Temperate (European, western North American, Japan, Australia, and southern Africa), West Atlantic, Dry Coast (seasonal or permanent dryness), and Tropical. These marshes remain better known botanically and ecologically than geomorphologically, but the level of knowledge is extremely uneven. Knowledge is most complete in the case of northwest Europe and the West Atlantic (Allen, 2000a; Friedrichs and Perry, 2001), but new work continues to appear (Duffy and Devoy, 1999; van Proosdij et al., 1999; Mellalieu et al., 2000; Thibault et al., 2000; Chmura et al., 2001a,b; Davidson-Arnott et al., 2002; Healey and Hickey, 2002; Ke and Collins, 2002; Nielsen and Nielsen, 2002; Reed, 2002; van der Wal et al., 2002; Cooper and Power, 2003; Haslett et al., 2003; Temmerman et al., 2003a; Bartholdy et al., 2004; Swift et al., 2004; Ollerhead et al., 2005; Frouin et al., 2007a). There are accounts from Spain (Castillo et al., 2002; Cearreta et al., 2002), the Adriatic (Fagherazzi et al., 1999; Rinaldo et al., 1999a,b), the Persian Gulf (Baeteman et al., 2004), and China (Yang, 1999). These reports confirm the great variability of salt marshes. How this depends on sediment supply, tidal regime, climate, and evolution versus inheritance remains poorly understood.
3. W HY SALT M ARSHES E XIST? Availability of suspended fine sediment is a necessary condition for the existence of a salt marsh at a point in a tidal system. The tidal streams in estuaries and embayments are slack at approximately low water and high water (Pugh, 1987), but the mud that can then settle may be eroded-advected if the intervening ebb/ flood currents are sufficiently strong and enduring. Extending earlier ideas (Allen, 2000a; Roberts and Whitehouse, 2001), the sufficient condition therefore is that Z ðHyÞ;T Z ðHyÞ;T ðdepositionÞdydt > ðerosionÞdydt ð1Þ where H is the level of the highest astronomical tide (HAT), y height of the sedimentary surface in the tidal frame, t time, and T an average wave and tidal
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year. Permanent accretion then occurs over a certain vertical range in the uppermost part of the tidal frame, but how close it extends up to the level of HAT is conditioned by other factors (Equation (2)). Deposition and erosion are predictable (McCave and Swift, 1976; Dyer, 1986), given the concentration and settling velocity of the fine sediment in the tidal waters, and the stresses due to wave and tidal currents. This conceptual model demands empirical data (Roberts and Whitehouse, 2001) and for predictions needs to be extended geographically. A first step would be to map the occurrence of salt marshes in energy space, for example, a wave height-tidal range variogram (Anthony and Orford, 2002).
4. G EOMORPHOLOGY 4.1. Marsh evolution versus inheritance The history of contemporary marshes is often ignored, with the consequence that the extent to which their geomorphological characteristics are inherited or evolved remains unknown. Marsh history can be resolved by stratigraphical and environmental analyses, constrained by high-resolution dating and by appropriate geochemical and textural work. Any available documentary and air-photographic record should never be ignored. At least two developmental paths are suggested by envirostratigraphic work and modeling (Allen, 2000a,b, 2003; Allen and Haslett, 2002, 2006a, 2007). Because of regime change, a marsh can build up from a mudflat, sandflat, or erosional surface lying significantly below HAT, independently of the behavior at the time of relative sea level (RSL). Vertical accretion on these surfaces is initially very rapid, typically measuring centimeters annually, but falls off with growth toward HAT and loss of accommodation, equivalent to hydraulic duty (Allen, 1997). Such a marsh can inherit a drainage pattern from a prior depositional surface, as a number of early conceptual models allow (Allen, 2000a), but subsequent drainage evolution and complication is far from precluded. On the other hand, as a result of rising RSL, marsh growth can begin on an unchanneled peat wetland initially active at or above HAT, as happened many times during the Holocene. With growing hydraulic duty, a tidal drainage network, without benefit of some prior system, evolves in its entirety simultaneously with the marsh. The vertical accretion rate is at all times commensurate with the behavior of RSL and can at times be high if the latter is fluctuating.
4.2. Marsh edges and coastal change Active salt marshes display a three-member continuum of outer-edge forms (Allen, 1993, 2000a). At a gradational edge, the marsh passes smoothly down through a ragged, patchy zone of pioneer plants into a mudflat of a similar to slightly lower slope. Such edges are typical of marshes growing seaward as well as upward. High rates of vertical accretion may be obtained, but these can be misinterpreted in the
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Holocene stratigraphic record where evidence for simultaneous lateral growth has been overlooked. At the other extreme are edges defined by eroding decimeter– meter high cliffs (Ke and Collins, 2002). Blocks of marsh silt toppled or slumped at their feet testify to often runaway retreat. A low chenier of storm-deposited sand with lumps of marsh silt may line the cliff top. An intermediate, near-static case is suggested by ramped margins. The marsh is bordered by an outer zone of moderate slope marked at right angles by wave-scoured, parallel ridges and grooves with little or no vegetation. These edge forms were placed conceptually by Ollerhead et al. (2005) in a natural cycle of marsh growth retreat, a feedback model pioneered by Yapp et al. (1917). Essentially, differential vertical growth between a sediment-trapping marsh and an adjoining exposed mudflat increases wave impacts along the steepening boundary and promotes eventual cliffing and retreat. Gao and Collins (1997) and, more satisfactorily, van de Koppel et al. (2005) modeled this pattern, the latter suggesting a decadal-century timescale for the process. Other possible causes of marsh retreat – all with some empirical support – include channel wandering (Pringle, 1995), changes in wave focussing (Otto, 1998), hurricanes (van de Plassche et al., 2006), sea level rise (Schwimmer and Pizzuto, 2000; van Wijnen and Bakker, 2001; Adam, 2002), and shipping movements (Castillo et al., 2002).
4.3. Marsh terraces Many salt marshes comprise a sequence of terraces (Figure 1) that step down seaward, the clifflets between descending beneath the marsh surface to link with buried wave-cut platforms (Allen, 2000a). These sequences record cyclic marsh growth and retreat under regimes permitting net accretion. Some clearly depended on periodic channel wandering (Pringle, 1995), but others seem in terms of timescale to support the Yapp-van de Koppel model above. Up to four terraces mark Severn Estuary marshes in southwest Britain. The youngest, locally ramped or cliffed, dates from the 1940s and its predecessor from the 1880s or 1890s. The marsh before that arose some time in or after the late 17th century, and its predecessor in medieval times. Simultaneous marsh edge retreat and renewed marsh growth are not precluded by the model and indeed can be observed. The development of terraced marshes calls for more envirostratigraphical and historical work, in order to refine timescales and the sweep of the coastal movement.
4.4. Channels, creeks, and gullies The components of marsh drainage networks vary hugely in scale. Strictly, channels are the widest and deepest, with beds partly submerged at low tide. The lowestorder elements are decimeter-scale gullies, wetted only by the highest tides. Between come creeks, which offer an appropriate collective term. Many salt marsh creeks have a winding planform which can be fractal (Angeles et al., 2004) but meander scale and degree vary much within and between marshes (Allen, 2000a). Scale increases with tidal discharge but sinuosity tends to decline. Analyzing three different marshes, Marani et al. (2002) found that tidal meanders
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plotted with river and other meanders on the same meander length-channel width graph. They also concluded that the width:depth ratio was constant at approximately 6, even for the smallest creeks, but this is not a universal finding (Fagherazzi and Furbish, 2001). An understanding of how meander and channel characteristics are controlled by tidal range, sediment character and sequence, the vegetation cover (Gran and Paola, 2001), and the role of evolution versus inheritance remains limited. Progress is being made through modeling in the case of cross section and scale (D’Alpaos et al., 2006). Modeling has also been successfully used to explore channel initiation through the focussing of tidal currents by clumps of pioneer vegetation on an otherwise bare sedimentary surface, another process Yapp et al. (1917) were early to model conceptually. It is now clear in the mobile-bed case (Eaton et al., 2006; Seminara, 2006) that meandering is essentially related to a fundamental instability of the erodible boundary in the presence of the flow. This approach should be extended to tidal meanders, subject to bidirectional (but generally unequal) flows of suspended fine sediment and with cohesive, partly vegetated banks that erode very slowly but not simply (Fagherazzi et al., 2004). What each meander sweeps out and encloses as it wanders is a tidal lateral deposit (McClennan and Housley, 2006). Such deposits should be common in Holocene estuarine sequences but will be hard to detect by borehole/auger sampling
4.5. Creek networks The issue of evolution versus inheritance is perhaps most acute for marsh drainage networks (Figure 1). Are network properties independent of environment and time? There is abundant evidence from marshes prograding over mudflats or sandflats (review in Allen, 2000a) for the inheritance of at least the higher order network components (Ollerhead et al., 2005) and also for the evolution of lower order elements within such marsh as has formed (Perillo and Iribarne, 2003) along Glock’s (1931) subsequently elaborated (Schumm et al., 1987; Stark, 1991; Kramer and Marder, 1992; Hasbargen and Paola, 2000; Pelletier and Turcotte, 2000; Clevis et al., 2003) conceptual lines. On the other hand, as exposed Holocene sequences reveal (Allen, 2000b), when unchanneled peat marshes experience transgression followed by regression as they are replaced by salt marsh during a fluctuation of RSL, creek networks extend and expand and then shrink and retract (Allen, 2000a, 2003). The conceptual treatments of these processes have been largely vindicated by an important series of models distinguished by the inclusion of vegetational factors (Mudd et al., 2004; D’Alpaos et al., 2005, 2007; Marani et al., 2006). Sufficiently large channels, however, can survive all RSL changes and be persistent wetland features (Baeteman, 2005). Further modeling of the steady-state case has been undertaken by Kirwan and Murray (2007), who have also confirmed a previously demonstrated lag between sea level and marsh response in the case of fluctuating RSL (Allen, 1995; Kirwan and Murray, 2005). The extent to which tidal drainage networks resemble rivers has been further explored. Although fractal (Cleveringa and Oost, 1999), tidal-flat networks are not equivalent to fluvial ones (Hofstede, 2006). Salt marsh networks also differ in a
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number of respects from terrestrial systems (Fagherazzi et al., 1999; Rinaldo et al., 1999a,b; Marani et al., 2003; Novakowski et al., 2004). Radical differences in functional scale and of role could be responsible, for terrestrial networks are sediment sources, whereas tidal ones are net sinks.
5. MORPHODYNAMICS 5.1. Tidal regime A multiperiodic tidal regime (Pugh, 1987) affects salt marshes (semidiurnal/diurnal, spring-neap, lunar monthly, equinoctial, and lunar nodal). The pattern of high waters changes over the tidal year and the timing of these relative to daylight hours varies geographically. While the first factor affects sedimentation, the second influences biological processes. Whereas most to all tides enter the dissecting creeks (undermarsh/bankfull tides), only a modest proportion (overmarsh tides) rise high enough to drown intervening vegetated platforms (Steel and Pye, 1997), the duration of tidal wetting (hydroperiod) falling steeply as this surface approaches HAT. There can be long periods (mainly summer) in certain years when the higher marshes are never or seldom drowned (storm surges excepted). Recognition of tidal complexity is important in the design and execution of field studies.
5.2. Sediment sources and supply Both exogenous and endogenous sources furnish sediment to salt marshes. The exogenous source is suspended fine mineral matter from river catchments and from cliffs and the sea bed within and external to the tidal system and includes detrital plant debris. Land use, changing over time, partly through human interference, strongly affects the amount and grade of the catchment supply. Marsh plants are the endogenous source, affording belowground root matter and aboveground litter (reviews in Turner, 1976; Good et al., 1982; Boorman, 2000; Allen, 2000a; Friedrichs and Perry, 2001). Absolute and relative productivities – still poorly understood – vary with species but tend to grow with decreasing hydroperiod and the proportion of mineral matter and decline with distance from creeks. Rainfall (de Leeuw et al., 1990), nitrogen availability (Kiehl et al., 1997), and herbivory (Kiehl et al., 1996; Ford and Grace, 1998; van Wijnen et al., 1999) all constrain productivity. Productivity shows great interannual variation (de Leeuw et al., 1990; Boorman and Ashton, 1997), an effect detectable in subfossil salt marsh sediments (Allen and Haslett, 2006a). Deposited plant matter decays exponentially (grazing, bacterial and fungal attack, and oxidation), with the proportion of refractory compounds (e.g., lignin) increasing.
5.3. Channelized flows Flow and sediment transport in salt marsh channel networks received early attention (reviews in Allen, 2000a; Friedrichs and Perry, 2001) but has seen little recent
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work (Schostak et al., 2000; Lawrence et al., 2004). Typically asymmetrical over a tidal cycle, the flow depends for strength on whether the tide is undermarshbankfull or overmarsh. In the latter case, as the marsh platform floods and then drains, speeds in the channels can exceed 1 m/s and turbulence levels are high. A significant proportion of the tidal prism that drowns a marsh enters over the marsh edge and not through channels (Lawrence et al., 2004). The proportion appears to be inversely related to marsh age (height) but to increase with tidal height.
5.4. Platform flows Tidal flows over salt marsh platforms are an order of magnitude weaker and less turbulent than in adjacent channels and exhibit a time-dependent boundary layer varying with the seasonally changing areal density, scale, and architecture of the plant cover and with the relative position of the water surface (reviews in Allen, 2000a; Friedrichs and Perry, 2001). These streams advect suspended fine sediment across marsh platforms from the immediate sources provided by the water flowing along marsh edges and within channels. This difficult aspect of the marsh environment attracts considerable attention, demanding modeling as well as experimental and field studies (Nepf, 1999; Nepf and Koch, 1999; Christiansen et al., 2000; Nepf and Vivoni, 2000; Shi et al., 2000; King, 2001; Davidson-Arnott et al., 2002; Leonard and Reed, 2002; Zheng et al., 2003; Lawrence et al., 2004; Neumeier and Amos, 2004; Neumeier and Ciavola, 2004). The pattern of flow vectors changes quickly as tidal waters spread over and then drain from marsh platforms. Flow speed and turbulence decay rapidly with increasing distance from marsh edges and channel banks and with increasing density of vegetation. Plant architecture also influences the plant–flow interaction. All of these factors favor the rapid settling of sediment and its retention on the surface. Silt can, however, be temporarily trapped on the plants themselves. That some degree of wave activity invariably accompanies overmarsh flows further complicates matters, introducing its own patterns of velocity and turbulence. The damping of waves on salt marshes because of the frictive effect of plant–flow interactions is of great value in coastal flood protection (Mo¨ller et al., 1999, 2001).
5.5. Accretion, compaction, and sea level change Modeling provides essential insights into the morphodynamics of intertidal mudflats and salt marsh platforms (Day et al., 1999; van Wijnen and Bakker, 2001; Allen, 2003; Temmerman et al., 2003b; Kirwan and Murray, 2005, 2007; D’Alpaos et al., 2007). The change DE in the elevation of the surface (x, z) relative to the moving tidal frame during a time increment Dt ca be stated as DEðx; zÞ = ðDSmin þ Sorg Þðx; zÞðDM þ DPÞðx; zÞ
DMðx; zÞ = DMtr ðx; zÞ þ DMfl ðx; zÞ þ
DMextr 2ðx; zÞ
ð2Þ
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where DSmin and DSorg are, respectively, the added thicknesses of mineral (tidally advected) and organic (indigenous plant root matter and litter) sediment accreted on the surface; DM the total change in RSL (oceanic and crustal components); DP the lowering of the surface through sediment compaction; DMtr and DMfl, respectively, the underlying long-term and fluctuating components of RSL; and DMextr the extreme tidal range. DSmin is especially complex, increasing with the settling velocity and concentration of suspended fine sediment and with hydroperiod. Hardest to model is DSorg because of the many, poorly understood controls on plant productivity. Equation (2) emphasizes the many controls on salt marsh morphodynamics. Any factor that behaves unsteadily and nonuniformly will excite a corresponding unsteady and nonuniform response in terms of relative elevation, potentially evoking a geographically as well as temporally varying and probably lagged environmental response. Only when all factors are steady and uniform will a salt marsh platform maintain a relative elevation constant in time and space. Equation (2) also reveals when sediment supply allows salt and peat marshes to interchange. Compaction is the continuous, irreversible vertical contraction of a sediment sequence because of compression of the mineral skeleton and chemical/biochemical loss of mass (shell debris and plant matter). The lowering of the sedimentary surface noted above is one consequence. Because it is difficult to either model or establish empirically (Hutchinson, 1980; Cahoon et al., 2000, 2002a,b), compaction in salt marsh studies is often either ignored or handled in an over-simplified way (but see Massey et al., 2006). However, the rate of surface lowering in the context of Holocene silt-peat sequences (Nieuwenhuis and Schokking, 1997; Allen, 1999, 2000c; Cahoon et al., 2000; Reed, 2002; Cohen, 2003; Williams, 2003) can at times significantly exceed in magnitude (by 2–4 times) the underlying rate of RSL rise (mid–late Holocene). Compaction can therefore be an important salt marsh process but its role in particular cases calls for appropriate evidence (Long et al., 2006). In embanked or drained marshes, the lowering of water tables can lead to compaction rates an order of magnitude greater than those allowable under natural circumstances (Hutchinson, 1980). Salt marsh accretion (reviews in Allen, 2000a; Friedrichs and Perry, 2001) rightly continues to attract attention (Reed et al., 1999; Christiansen et al., 2000; Chmura et al., 20001a,b; Davidson-Arnott et al., 2002; Neubauer et al., 2002; Haslett et al., 2003; Temmerman et al., 2003a; Culberson et al., 2004; van Proosdij et al., 2006). Long-term accretion is maintaining most marshes at a comparatively constant position relative to HAT. Aside from a wave-influenced outer fringe, and also on the long term, the rate declines inland from the marsh edge/channel banks on embedded subcatchment, catchment, and whole marsh scales (Figure 2). On the short term, the rate increases with hydroperiod and with fine-sediment concentration. Organic matter grows in proportion with the decline of accreted mineral sediment (Reed et al., 1999). The concentration of intercalated peats in mid-Holocene sequences of tidal silts affords the most striking evidence that water and relative marsh level have fluctuated postglacially, causing frequent and rapid changes in the wetland environment (Allen, 2005). Much less marked fluctuations continue, but the main concern must be the recent anthropogenic acceleration of RSL rise (Gehrels et al., 2002, 2005; Church and White, 2006).
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Figure 2 Textural values on salt marsh platforms, Southern Basin, Venice Lagoon, northeastern Italy: (a) percent sand (>63 mm), (b) percent clay (<4 mm). Data of Barillari (1977^1978).
6. S EDIMENTOLOGY 6.1. Grain size Grain size on marsh platforms declines on embedded subcatchment, catchment, and whole marsh scales away from immediate sources of silt (Woolnough et al., 1995). This aspect of marsh accretion has been neglected (reviews in Allen, 2000a; Friedrichs and Perry, 2001) despite its influence on properties such as compressibility, which increases with clay content. Barillari (1977–1978) and Yang (1999) found the expected whole marsh pattern (Figure 3). Laser granulometry is an appropriate measurement technique (Allen, 2004; Blott et al., 2004), but hydrogen peroxide treatment during preparation, removing all plant matter, should be avoided (Allen and Thornley, 2003). Curve-shape analysis is helpful in assessing granulometer results (Allen and Haslett, 2006a).
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Figure 3 Sediment thickness (mm) accumulated over a 1-year period (1979^1980), New Marsh, TheWash, North Sea coast, eastern England. Data of Hartnall (1984).
6.2. Tidal bedding Where accretion is sufficiently rapid, salt marsh bedding registers multiperiodic tidal events and seasonal to longer term environmental change (e.g., sea-temperature, windiness, and river discharge). These patterns may be established by the direct (Allen and Haslett, 2002; Long et al., 2006; Frouin et al., 2007b) or indirect (Weidlich and Bernecker, 2004; Meyer et al., 2006) assessment of layer numbers and thickness, perhaps after impregnation (Boe¨s and Fagel, 2005) and by highresolution grain-size analysis (Allen, 2004; Long et al., 2006; Allen and Haslett, 2006a). However, work on thickness, amounting almost to an obsession, emphasizes tidal periodicities, but grain size best reveals nontidal effects. Many salt marsh deposits show submillimeter–millimeter scale laminae ascribable to individual tides and commonly a centimeter–decimeter scale coarse–fine banding (Figure 4), with or without lamination (Figure 5), attributable to seasonal
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Figure 4 Mid Holocene, annually banded, laminated salt marsh silts (coarser-textured laminae recessive), Caldicot Level,Welsh Severn Estuary southwest Britain. Scale, 0.2 m.
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Figure 5 Photomicrograph of lamination and banding, mid-Holocene, annually banded silts, Caldicot Level, Welsh Severn Estuary, southwest Britain. Clayey sublaminae light-colored. Original sample measures 18 31 mm.
patterns (Tessier, 1998; Allen and Haslett, 2002, 2006a; Stupples, 2002; Allen, 2004; Long et al., 2006; Frouin et al., 2007b). Pollen evidence, carbon isotope values, and C/N ratios supports the annual origin of the banding (Dark and Allen, 2005; Allen and Dark, 2008; Allen et al., 2007). In a representative sequence of banded, salt marsh silts from the Holocene of the Severn Estuary, southwest Britain, the mean grain size in the coarsest parts of the bands is up to twice that in the finest (Figure 6). Whereas the textural contrast within bands denotes seasonal changes in
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Figure 6 Textural profile for a selected sequence of annual bands, mid-Holocene salt marsh silts, Caldicot Level,Welsh Severn Estuary. Data of Allen (2004). The sediment, collected as a monolith, was subsampled in contiguous, 5-mm slices parallel to the bedding and subjected to laser grain-size analysis.The modal strength is the frequency of the modal grain size.
sea temperature and windiness, that between bands would appear to reflect interannual variations (Allen, 2004; Allen and Haslett, 2006a). The presence of lamination and/or banding in intertidal sediment points to a high rate of short–medium term accretion, which need not depend on the contemporaneous behavior of RSL. Such rates can arise in at least four stratigraphic contexts: (1) on intertidal erosion surfaces as the result of regime change (Allen and Haslett, 2002, 2006a; Allen, 2003, 2004), (2) on matured salt marsh platforms when water levels are rising most rapidly during a local regional fluctuation (Allen, 2003, 2004; Allen and Haslett, 2007), (3) when salt marshes, such as those in the Baie du Mont Saint Michel (Larsonneur, 1994; Tessier, 1998), on the French Channel coast, spread rapidly over a prior sandflat or mudflat whether or not RSL is also changing, and (4) during channel-infilling (Stupples, 2002), for example, during the regressive phase of a water-level fluctuation (Allen, 2000b, 2003; Allen and Haslett, 2006b). Only in the second context is accretion likely to be purely vertical. In the closely linked first and third, significant lateral accretion is likely as well, making the rate measured from a borehole or vertical sedimentological log merely a component of the true rate. A fifth possible context is where a shallow lying peat is collapsing and lowering a wetland, a process which, it has been pointed out (Allen, 1999, 2000c), can promote local marine transgression and significantly enhance tidal accretion rates (Baeteman et al., 2002; Baeteman 2005). Long et al. (2006) attribute the late Holocene transgression of the extensive Romney Marsh peatlands on the
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English Channel coast to this factor. Two difficulties suggest a possibly exaggerated role for peat compaction in this particular case. First, the variously laminated, postpeat sands and silts on which the claim rests occur at an inner estuary site, but without contextual evidence that the accretion was purely vertical and not on an erosion surface, accretionary terrace, or sloping point-bar (Dalrymple et al., 1991; McClennan and Housley, 2006). Second, the sharp increase in grain size low in the postpeat sequence suggests discontinuous accretion and an early change of context.
6.3. Lithostratigraphic architecture Holocene coastal envirostratigraphy is best known in northwest Europe, where it is clear that the deposits are laterally very variable (Allen, 2000a,c, 2001; Streif, 2004). Variability has several expressions and is a response to RSL fluctuations, geographical environmental gradients, coastal change, and differential compaction. The facies involved are mudflat and salt marsh silts, rooted peats deposited in highest intertidal-supratidal marshes, paleochannel fills of silt or sand, and sandy-gravelly barrier deposits. The underlying rate of sea level rise during the early Holocene was so great that any contemporaneous water level fluctuations introduced no easily recognized variations into the mudflat-salt marsh facies deposited. During the mid-Holocene, however, RSL fluctuations, partly under a regional control, created almost everywhere a sequence of alternating salt marsh silts and peats with strong teleconnections (Long et al., 2000; Allen, 2005). Unique networks of tidal channels grew and decayed during each transgressive–regressive cycle, eliminating pre-existing strata and introducing new facies (Allen, 2000b; Allen and Haslett, 2006b). High-density radiocarbon dates reveal that environmental boundaries shifted rapidly and systematically across the depositional surface (Allen, 2005). Peats thickened and fused landward but seaward passed into buried soils with root horizons. Meanwhile, coastal change, perhaps favored by accelerated sea level rise, created erosional discontinuities within the sequence, locally eliminating some silts and peats and replacing them either with coastal barrier deposits or with salt marsh silts that grew up on erosional platforms. Late Holocene salt marsh deposits lack intercalated peats but, where human interference was minimal, do reveal occasional buried soils, some with evidence for temporary human occupation. The lack of late Holocene peats has been ascribed chiefly to the decay of the mid-Holocene RSL fluctuations (Allen, 2005), but other factors have been favored (Long et al., 2000, 2006). The provision of accommodation by underlying sea level rise, allowed the accretion of a final Holocene sequence 10–15 m thick, in the course of which differential compaction proceeded unbroken. Not only was deposition on the wetland surface continuously affected, but significant lithostratigraphic stratigraphic distortion and displacement occurred in the underlying deposits (Allen, 1999, 2000a,c). Depending on circumstances, essentially isochronous lithological contacts now vary in altitude by as much as approximately 4 m on geographical scales of tens to hundreds of meters. The latest illustrations of the effect are from Romney Marsh (Long et al., 2006) on the English Channel and southwest England (Massey et al., 2006).
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7. CONCLUDING D ISCUSSION Salt marshes have rightly continued to attract attention especially because of their importance as habitats and as defensive components of the coastal zone. Although much has been learned about them, knowledge of their distribution, character, and functioning remains uneven and not well systematized. In the case of marshes active today, the emphasis has fallen on “local” rather than on comparative, “regional” studies, and on “one-time” rather than historical work, with the consequence that there is little understanding of how such major, changeable factors as climate, tidal range, and sediment supply determine their geomorphology and functioning. Although restricted to a single aspect of their morphodynamics, the work of Leonard and Reed (2002) around the North Atlantic region nonetheless demonstrates what can be achieved by a comparative approach over a wide geographical range. More attention should be given to the developmental history of contemporary marshes, in order that the effects of relative maturity and of inheritance versus evolution can be properly understood. Salt marshes are dynamical systems that exist in the flow of time and are under the influence of changeable factors. The state of a marsh at any one time affects future states and is determined by previous states. In terms of status, the marshes that are active today present a considerable range, from those initiated a millennium or two ago to those no more than a few decades old. Two kinds of study would in particular facilitate this understanding. One is the contriving of better general numerical models based on sound principles that attempt to describe the developmental history of marshes under changing conditions of sea level, tidal range, wave activity, and sediment supply. This aim will call for the creation or refinement of submodels for, in particular, sediment transport and deposition across marsh platforms, the supply of organic sediment, and sediment compaction. Great promise here is shown by the new generation of models which incorporate the role of vegetation. The other and complementary kind of study is the high-resolution environmental and sedimentological analysis of Holocene stratigraphic sequences of coastal wetland origin, especially where these sequences can be examined on natural exposures. This would provide a detailed picture of the circumstances under which salt marshes have arisen in the recent past and reveal how they have responded to conditions that have changed on subdecadal to millennial timescales.
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J.R.L. (Eds.), Coastal and Estuarine Environments: Sedimentology, Geomorphology and Geoarchaeology. Geological Society, London, Special Publications 175, pp. 59–93. Schumm, S.A., Mosley, M.P., Weaver, W.E., 1987. Experimental Fluvial Geomorphology. John Wiley, New York, Schwimmer, R.A., Pizzuto, J.E., 2000. A model for the evolution of marsh shorelines. J. Sediment. Res. 70, 1026–1035. Seminara, G., 2006. Meanders. J. Fluid Mech. 554, 271–297. Shi, Z., Hamilton, L.J., Wolanski, E., 2000. Near-bed currents and suspended sediment transport in salt-marsh canopies. J. Coast. Res. 16, 909–914. Stark, C.P., 1991. An invasion percolation model of drainage network evolution. Nature 352, 423–425. Steel, T.J., Pye, K., 1997. The development of saltmarsh tidal creek networks: evidence from the UK. Proceedings of the Canadian Coastal Conference, 1997. Canadian Society for Coastal Sciences, Guelph, pp. 267–280. Streif, H., 2004. Sedimentary record of Pleistocene and Holocene inundations along the North Sea coast of Lower Saxony, Germany. Quatern. Int. 112, 3–28. Stupples, P., 2002. Tidal cycles preserved in later Holocene tidal rhythmites: the Wainway Channel, Romney Marsh, southeast England. Mar. Geol. 182, 231–246. Swift, L.F., Devoy, R.J.N., Wheeler, A.J., Sutton, G.D., Gault, 2004. Sedimentary dynamics and coastal changes on the south coast of Ireland. J. Coast. Res. SI 39, 110–117. Temmerman, S.S., Govers, E., Wartel, S., Meire, P., 2003a. Spatial and temporal factors controlling short-term sedimentation in a salt and freshwater tidal marsh, Scheldt Estuary, Belgium, SW Netherlands. Earth Surf. Proc. Land. 28, 739–755. Temmerman, S.S., Govers, G., Meire, P., Wartel, S., 2003b. Modelling long-term tidal marsh growth under changing tidal conditions and suspended sediment concentration, Scheldt Estuary, Belgium. Mar. Geol. 193, 151–169. Tessier, B., 1998. Tidal cycles: annual versus semi-lunar records. In: Alexander, C.R., Davis, R.S., Henry, V.J. (Eds.), Tidalites: Process and Product. Tulsa, Society of Economic Paleontologists and Mineralogists Special Publication 61, 69–74. Thibault, J.J., Swabson, L.A., Chmura, G.L., 2000. Coastal peatlands and salt marshes of New Brunswick, Canada. New Brunswick Department of Natural Resources and Energy, Minerals and Energy Division, Open File 2000-12, Fredericton, 38pp. Turner, R.E., 1976. Geographic variations in salt marsh macrophyte production: a review. Contrib. Mar. Sci. 20, 47–68. van de Koppel, J., van der Wal, D., Bakker, J.P., Herman, P.M.J., 2005. Self-organization and vegetation collapse in salt marsh ecosystems. Am. Nat. 165, E1–E12. van de Plassche, O., Erkens, G., van Vliet, F., Brandsma, J., van der Borg, K., de Jong, A.F.M., 2006. Salt-marsh erosion associated with hurricane landfall in southern New England in the fifteenth and seventeenth centuries. Geology 34, 829–832. van der Wal, D., Pye, K., Neal, A., 2002. Long-term morphological change in the Ribble Estuary, northwest England. Mar. Geol. 189, 249–266. van Proosdij, D., Ollerhead, J., Davidson-Arnott, R.G.D., 2006. Seasonal and annual variations in the volumetric sediment balance of a macro-tidal salt marsh. Mar. Geol. 225, 103–127. van Proosdij, D., Ollerhead, J., Davidson-Arnott, R.G.D., Schostak, L.E., 1999. Allen Creek Marsh, Bay of Fundy: a macro-tidal coastal salt marsh. Can. Geogr. 43, 316–322. van Wijnen, H.J., Bakker, J.P., 2001. Long-term surface elevation change in salt marshes: a prediction of marsh response to future sea-level rise. Estuar. Coast. Shelf Sci. 52, 381–390. van Wijnen, H.J., van der Wal, R., Bakker, J.P., 1999. The impact of herbivores on nitrogen mineralization rate: consequence for salt-marsh succession. Oecologia 118, 225–231. Weidlich, O., Bernecker, M., 2004. Quantification of depositional changes and palaeoseismic events using outcrop data. Sediment. Geol. 166, 11–20. Whigham, D., Dykyjova´, D., Hejny, S., 1993. Wetlands of the World, vol. I. Kluwer, Dordrecht. Williams, H., 2003. Modelling shallow autocompaction in coastal marshes using caesium-137 fallout: preliminary results from the Trinity River Estuary, Texas. J. Coast. Res. 19, 180–188. Woolnough, S.J., Allen, J.R.L., Wood, M.A., 1995. An exploratory numerical model of sediment deposition over tidal salt marshes. Estuar. Coast. Shelf Sci. 41, 515–543.
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C H A P T E R
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E COSYSTEM S TRUCTURE OF T IDAL S ALINE M ARSHES Jenneke M. Visser and Donald M. Baltz
Contents 1. Introduction 2. Saline Marsh Communities 2.1. Emergent vegetation 2.2. Benthic algae 2.3. Nekton 2.4. Reptiles 2.5. Birds 2.6. Mammals 3. Interaction Among Communities 3.1. Effects of animals on emergent vegetation distribution 3.2. Emergent vegetation as animal habitat 3.3. Nursery function 3.4. Saline marsh food webs Acknowledgments References
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1. INTRODUCTION Tidal saline marshes are found in middle and high latitudes along shores throughout the world (Chapman, 1977). The largest concentrations of tidal marshes are found along the South Atlantic and Gulf coasts of North America followed by China (Greenberg et al., 2006). The physical features of tides, sediments, freshwater inputs, and shoreline geomorphology determine the development and extent of tidal saline wetlands (Mitsch and Gosselink, 2000). Tidal saline marshes are generally found in sedimentary environments and can be broadly classified into those that originate on reworked marine sediments and those that are formed at the margins of river deltas on riverine sediments. Tidal regimes vary from microtidal (<2 m) to macrotidal (>6 m) and can also be diurnal, semidiurnal, or mixed. Although not included in this chapter, we recognize that humans have had substantial impacts on coastal systems, including salt marshes and associated estuaries Coastal Wetlands: An Integrated Ecosystem Approach
2009 Published by Elsevier B.V.
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(Baltz and Ya´n˜ez-Arancibia, in press). Over several centuries, expanding human populations in the coastal zone and expanding fishing pressure on finite resources have led to altered estuarine and coastal ecosystems through a series of anthropogenic effects and can be characterized as the Shifting Baseline Syndrome (Pauly, 1995). Fishing has generally been first and foremost, coming before pollution, habitat destruction, introductions of exotic species, and climatic change in the timing and degree of impact (Jackson et al., 2001). Historic alterations have significantly modified salt marsh ecosystems to the point of dramatic functional impairment or outright marsh destruction (Bertness et al., 2002). To fully appreciate the current condition of saline marshes and foresee the direction of future changes, we need a better understanding of the history of interactions between nature and society in our coastal systems (Kates et al., 2001).
2. SALINE MARSH COMMUNITIES 2.1. Emergent vegetation Tidal saline marshes are defined as natural or seminatural halophytic emergent vegetation on alluvial sediments (Beeftink, 1977). The emergent vegetation of tidal saline marshes changes in species composition depending on geographical location (Adam, 1990, Table 1). Within geographical locations species composition varies along inundation gradients. A low (submergence) and a high (emergence) marsh zone are generally recognized. The marsh below mean high water is regularly flooded and has reduced soils for most of the time except along creek banks with good drainage (Armstrong et al., 1985). This zone is generally species poor and is dominated by species that are both salt and flood tolerant. The high marsh is found above mean high tide and is less frequently flooded, and has oxidized soils that may briefly become reduced during flood events (Armstrong et al., 1985). In some areas, an intermediate middle marsh zone may be distinguished by different plant species composition. Salinity often decreases inland, but salinity inversions are relatively common, particularly when evapotranspiration exceeds precipitation (Mahall and Park, 1976; Callaway et al., 1990). In areas that are irregularly flooded, zonation may be absent or zonation may be related to distance from tidal creeks (Zedler et al., 1999; Costa et al., 2003). Flooding effects on biogeochemical cycling are described in detail in Chapter 16). The distribution of plant species in tidal saline marshes is determined by the physical and biological environment. Physical and geochemical factors that affect the distribution of tidal saline marsh species include flooding, salinity, and the ratio of sodium to potassium as well as the ratio of calcium to magnesium (Clarke and Hannon, 1970; Olff et al., 1988; Partridge and Wilson, 1989; Cantero et al., 1998; Alvarez Rogel et al., 2000; Huckle et al., 2000). Over a stress gradient, the distribution of a species is determined by physical constraints at higher stress levels and competition with other species at lower stress levels (Pennings and Callaway, 1992). Plant establishment where physical stress is high may decrease stress levels by increasing elevation, oxygenating the rhizosphere, or reducing soil salinity thereby
Dominant emergent plant species of tidal saline marshes
Family
Species
Marsh zone
Geographic extent
Source
Amaranthaceae Amaranthaceae Chenopodiaceae Chenopodiaceae
Low Low Middle Middle
Western North America South Pacific Europe Western North America
Pennings and Callaway (1992) Thannheiser and Holland (1994) Castellanos et al. (1994) Pennings and Callaway (1992)
Chenopodiaceae Cyperaceae Goodeniaceae Juncaceae Juncaceae Juncaginaceae
Salicornia virginica Sarcocornia quinqueflora Arthrocnemum perenne Arthrocnemum subterminale Suaeda maritima Carex glareosa Selleria radicans Juncus kraussii Juncus roemerianus Triglochin maritima
Low High Middle High Middle to high Middle to low
Japan Arctic South Pacific South Pacific Eastern North America Western Europe, Arctic, Asia
Plumbaginaceae Poaceae Poaceae
Limonium vulgare Distichlis scoparia Distichlis spicata
Middle to low High middle
Poaceae Poaceae Poaceae Poaceae Poaceae Poaceae Poaceae Poaceae Poaceae Poaceae Poaceae Primulaceae
Festuca rubra Pucinellia maritima Pucinellia phryganodes Spartina alterniflora Spartina anglica Spartina densiflora Spartina foliosa Spartina maritima Spartina patens Sporobolus virginicus Zoysia sinica Samolus repens
High Middle to low Low Low Low Low Low Low High High High Low
Europe and North Africa South America Western North America, South America Western Europe Western Europe Arctic Eastern North America Western Europe South America Western North America Europe Eastern North America Tropics Japan South Pacific
Adam (1990) Adam (1990) Thannheiser and Holland (1994) Adam (1990) Mitsch and Gosselink (2000) Bakker (1985), Davy and Bishop (1991) Boorman (1967), Bakker (1985) Cantero et al., (1998) Adam (1990), Cantero et al. (1998)
Ecosystem Structure of Tidal Saline Marshes
Table 1
Gray and Mogg (2001) Gray and Mogg (2001) Beaulieu and Allard (2003) Mitsch and Gosselink (2000) Gray and Mogg (2001) Cantero et al. (1998) Adam (1990) Castellanos et al. (1994) Mitsch and Gosselink (2000) Adam (1990) Adam (1990) Thannheiser and Holland (1994) 427
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facilitating the establishment of higher marsh species (Bertness and Shumway, 1993; Castellanos et al., 1994; Figueroa et al., 2003). However the relative effects of facilitation and competition are highly dependent on the stress tolerance of the local species (Pennings et al., 2003). Increasing nitrogen availability allows species to expand along the competitive edge (van Wijnen and Bakker, 1999; Bertness and Pennings, 2000). Sometimes parasitic plants can alter competitive outcomes. For example, in tidal saline marshes of the western United States, the parasitic salt marsh dodder (Cuscuta salina) facilitates two relatively uncommon plant species, sea lavender (Limonium californicum) and sea heath (Frankenia salina) by selectively infecting and suppressing the competitive dominant glasswort (Salicornia virginica) (Pennings and Callaway, 1996). Interaction with animals also plays an important role. In many tidal saline marshes across the globe, invasive plant species have started to replace the historic/native vegetation due to introductions of nonnative species, human alterations to the local hydrology, as well as increasing nutrient levels in estuarine waters (Thannheiser and Holland, 1994; Moyle, 1996; Castillo et al., 2000; Talley and Levin, 2001; Bertness et al., 2002; Vale´ry, 2004).
2.2. Benthic algae Benthic algae occur on the sediments below and adjacent to the emergent vegetation of tidal saline marshes, as well as on the culms of the emergent vegetation (epiphytic algae). Diatoms are universally present (Sullivan and Currin, 2000). Extensive cyanobacterial populations develop during the summer in Europe (Birkemoe and Liengen, 2000; Quintana and Moreno-Amich, 2002), as well as northeast and southwest coasts of the United States (Blum, 1968; Zedler, 1982). Green and brown algae reach the largest population sizes during seasons in which emergent vegetation is not dominant (Brinkhuis, 1977; Houghton and Woodwell, 1980; Sullivan and Currin, 2000). Distinct benthic algal communities are associated with different emergent vegetation communities and are related to differences in elevation, soil temperature, soil moisture, interstitial ammonium concentration, and canopy height (Sullivan and Currin, 2000).
2.3. Nekton Nekton can be divided into four categories based on marsh use (Peterson and Turner, 1994): (1) residents on the marsh surface that generally remain at low tide in pools and puddles, (2) regular visitors at high tide that retreat to fringing vegetation along the marsh edge at low tide, (3) individuals of larger species that associate strongly with the marsh edge as juveniles and penetrate only a few meters into the marsh at high tide, and (4) other subtidal species that rarely penetrate far onto the flooded marsh but may be associated with tidal creeks. Transient species may be found in marshes and include both fresh- and saltwater visitors as well as diadromous forms such as salmonids, clupeids, and anquillids that are present for one or more life history stages (Baltz et al., 1993; Dionne et al., 1999). In Louisiana marshes, tidal amplitude (~30 cm) is at the low end of the microtidal range and is easily dominated by winds making marsh flooding less predictable.
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Thus, interannual climatological variation influences the flooding duration and frequency of salt marsh habitats (Childers et al., 1990). In marshes with greater microtidal and mesotidal ranges (>1 m), inundation is more predictable and some fishes may spend as much as one-third of their time in flooded smooth cordgrass (Spartina alterniflora) (Hettler, 1989). For species that use intertidal zones as nurseries, interannual variation in habitat availability may have a strong influence on recruitment, particularly in microtidal systems (Childers et al., 1990; Baltz et al., 1993, 1998). For example, climatic conditions affect marsh accessibility for juvenile shrimp and are related to interannual variation in shrimp landings (Childers et al., 1990). Connolly (1999) reviewed study limitations that have hampered our understanding of the direct use of tidal saline marshes by nekton, primarily fishes, shrimps, and crabs, and suggested improvements to standardize sampling methods, overcome poor sampling designs, and improve assessment of flooding regimes and landscape structure. He also noted the uneven distribution of studies, largely limited to North America (90%), Europe (7%), and Australia (3%). Cattrijsse and Hampel (2006) recently reviewed intertidal marsh studies in Europe and contrasted floral, faunal, and physical patterns with North American coastal systems. In contrast to North American salt marshes, European salt marshes differ in having the lower limit of marsh vegetation defined by mean high water neap tides rather than mean tide level, vegetation typically dominated by sea purslane (Halimione spp.) rather than smooth cordgrass and a much higher stem density that inhibits nekton movement on the marsh surface. Laffaille et al. (2000) examined fish use of salt marsh vegetation in the macrotidal system of Mont Saint-Michel Bay, France, which is accessible to fishes for only a few minutes or hours during high spring tides (5–10% of tides). Due to this brief flood duration, no fishes are considered residents, but the annual pattern and three seasonal patterns of community structure are stable. Thirty-one fishes, netted in creeks as tides ebbed off of flooded marsh, included seven marine stragglers, 13 estuarine-dependent marine species, three catadromous species, and eight estuarine species. Most of the species can be characterized as euryhaline and eurythermal migrants. In Australia, Connolly et al. (1997) examined fish use of flooded saline marsh in the high intertidal zone that is generally separated from open water by fringing mangroves. Compared to tidal creeks draining the same marsh flats, the number of species and individuals caught in the high intertidal zone was lower; however, the density of fishes on marsh flats (1 per 23 m2) was higher than expected when creek numbers were parsed over drainage areas (1 per 134 m2). Only the two most abundant species found in creeks were captured on high marsh flats, glass goby (Gobiopterus semivestitus) and small-mouth hardyhead (Atherinosoma microstoma).
2.4. Reptiles Only a few species of snakes, turtles, and crocodilians can be considered common residents of tidal saline marshes (Neill, 1958). The salt marsh snake (Nerodia clarkia) is restricted to saline and brackish tidal wetlands along the Atlantic and Gulf coasts of North America. As a way to avoid predators, salt marsh snakes are nocturnal. The
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salt marsh snake does not drink saline water, and obtains water primarily from the foods it eats (Pettus, 1958). Diamondback terrapin (Malaclemys terrapin) is the only turtle restricted to tidal saline marshes (Greenberg et al., 2006). Special glands in the turtle’s eye region excrete excess sodium (Bentley et al., 1967). Juvenile and smaller male terrapins rely on the nearshore area where they forage on readily available prey such as clams, crabs, and small crustaceans. A large number of amphibian and reptile species are occasional visitors to tidal saline marshes, but are generally found in fresh to slightly brackish water (Greenberg et al., 2006).
2.5. Birds Typically, saline marsh avifauna is dominated, at least numerically, by large numbers of Anseriformes (waterfowl), Ciconiiformes (long-legged wading birds), and Charadriiformes (shorebirds, gulls, and terns) (Goss-Custard et al., 1977; Custer and Osborn, 1978; Bildstein et al., 1982; Erwin, 1996). In addition, Passeriformes (songbirds) feed and breed in saline marshes (Brown and Atkinson, 1996; Dierschke and Bairlein, 2004). In many marshes, avian populations increase considerably seasonally, not only during migratory periods, when large numbers of waterfowl and shorebirds congregate to feed and rest, but also during the breeding season, when wading birds congregate at traditional coastal colonies to nest. The total number of breeding bird species whose habitat primarily consists of tidal marshes has been estimated to be between 11 and 21, with only two songbird species that are entirely restricted to tidal saline marsh in North America: the seaside sparrow (Ammodramus maritimus) and the salt marsh sharp-tailed sparrow (Ammodramus caudacutus) (Greenberg et al., 2006). These species are adapted to nesting in cordgrass dominated tidal marshes (Benoit and Askins, 1999). Several subspecies of birds restricted to tidal marshes have wider distributions (e.g., a subspecies of the slender-billed thornbill Acanthiza iredalei rosinae) (Greenberg et al., 2006).
2.6. Mammals The total number of mammalian species, whose habitat primarily consists of tidal marshes, has been estimated to range from 13 to 26, with rodents predominant (Greenberg et al., 2006). Of these only the salt marsh harvest mouse (Reithrodontomys raviventris) is restricted to coastal marshes. In addition to wild mammals, domestic or feral mammals (primarily cattle, sheep, and horses) graze in some tidal saline marshes (Bakker, 1985; Turner, 1987).
3. INTERACTION AMONG C OMMUNITIES 3.1. Effects of animals on emergent vegetation distribution Burrowing organisms such as fiddler crabs (Uca spp.) increase tidal saline marsh plant production through moderation of soil stresses; for example, they increase soil
Ecosystem Structure of Tidal Saline Marshes
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aeration, oxidation–reduction potential, and in situ decomposition of belowground plant debris (Bertness, 1985), although the increase in production may depend on the severity of the stresses (Nomann and Pennings, 1998). In return, the crabs benefit from the shade and the shelter that the increased plant cover provides (Bortolus et al., 2002). Grazing by herbivorous waterfowl or cattle can significantly change the vegetation community in a tidal saline wetland (Ranwell, 1961; Bakker, 1985; Pehrsson, 1988). Grazing by breeding waterfowl in the Arctic can change the trajectory of plant succession by physically changing the environment. Moderate grazing preserves the dominance of grasses (Bazely and Jefferies, 1986), while heavy grazing can convert marsh to unvegetated mudflats (Handa et al., 2002). Intense grazing can also limit the distribution of plants along a maturation (time since plant establishment) gradient that is not related to elevation. Van der Wal et al. (2000) showed that the distribution of seaside arrowgrass (Triglochin maritima) in younger marshes is limited due to intense grazing by geese, hares, and rabbits. The distribution in more mature marshes is limited due to competition for light by taller species that are slower colonizers such as saltbush (Atriplex portulacoides). Field experiments indicate that the periwinkle (Littoraria irrorata) can overgraze otherwise healthy stands of smooth cordgrass and reduce them to bare mudflats when periwinkle predator density is low. Thus periwinkle predators including blue crabs (Callinectes sapidus) and terrapins are capable of exerting top-down control of smooth cordgrass (Spartina alterniflora) production (Silliman and Bertness, 2002). The density of grazers can be positively correlated with nitrogen availability (Bowdish and Stiling, 1998; Visser et al., 2006). The more stressful conditions in the low marsh may make plants at this elevation more palatable (Dormann et al., 2000) and grazing effects may be more pronounced in these more stressful environments (Kuijper and Bakker, 2005). Neighboring plant species may have both positive and negative effects on the level of grazing of a palatable species. Grazing on smooth cordgrass by an herbivorous crab (Neohelice (Chasmagnatus) granulata) is more intense when the plant grows with alkali bulrush (Scirpus maritimus) then when it grows in monotypic cordgrass stands (Costa et al., 2003).
3.2. Emergent vegetation as animal habitat Some tidal saline marshes such as the European Limonium, Sueda, and Salicornia marshes provide seeds eaten by several species of wintering songbirds (Brown and Atkinson, 1996). In addition, bulrush and cordgrass seeds are eaten by waterfowl (Mendall, 1949; Hartman, 1963; Landers et al., 1976; Gordon, 1998). Birds feeding on small aquatic organisms along the marsh edge occur more frequently in areas that have more open water in the form of marsh creeks and ponds (Craig and Beal, 1992). Both rodents and birds nesting in tidal saline marshes use the available vegetation as nesting material (Shanholtzer, 1974). Nests are generally constructed from blades of grass. Similar to forest species,
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breeding birds in tidal marshes are often associated with larger marsh tracts (Craig and Beal, 1992). Although tidal saline marshes probably function in a variety of ways to enhance growth and survival of a particular nekton species, the relative importance of food vs. refuge from predation is poorly understood (Boesch and Turner, 1984) and probably varies across species and life history stages. Both shallow water and vegetation in the marsh provide protection for small nekton from large predators, particularly from large piscivorous birds and fishes (Kneib, 1982a,b), and both provide a food-rich environment (Van Dolah, 1978; Gleason, 1986; Cyrus and Blaber, 1987; Gleason and Wellington, 1988). In experimental tests of predation as a factor determining the size-specific habitat difference between killifish (Fundulus heteroclitus) age classes, Kneib (1987) found that young killifish remained in high intertidal cordgrass habitat avoiding concentrations of larger piscivorous fishes in subtidal habitats, whereas habitat use by larger killifish is influenced by avian predators. Other field experiments indicate that predation pressure is lower and food availability is higher in vegetated than in unvegetated habitats (Rozas and Odum, 1988).
3.3. Nursery function Nurseries foster the growth and/or survival of early life history stages of fishes and macroinvertebrates (Beck et al., 2001). The particular environment used by a species may be characterized as nursery habitat if it can be shown that individuals are found at higher densities and experience enhanced survival and/or growth compared to nearby habitat types (Pearcy and Myers, 1974; Weinstein, 1979). It is difficult to separate the nursery function of flooded cordgrass marsh from that of adjacent habitat types. Baltz et al. (1993, 1998) examined the use of shallow open water and flooded smooth cordgrass as nursery habitat for a variety of resident and transient fishes in Louisiana estuaries. Small fishes and early life history stages of larger species often use shallow turbid water along the marsh edge at low tides and move onto flooded marsh at higher tides. Evidence suggests that the magnitude of fishery landings is correlated with the spatial extent of estuarine vegetation (Turner, 1977, 1992; Pauly and Ingles, 1988); therefore, extensive marsh loss in the northern Gulf of Mexico is a major concern for the sustainability of fisheries. However, the connection between fishery landings and marsh habitat loss in the northern Gulf of Mexico is not clear. Moreover, landings have increased for many species in spite of accumulating habitat alterations (Zimmerman et al., 1991). One hypothesis is that the marsh edge (i.e., the perimeter at the marsh–open water interface) is the essential habitat for many species, and that the nursery function and value will not decline and result in reduced landings until the quantity of marsh-edge perimeter begins to decline. During the process of marsh loss, the amount of marsh edge initially increases as solid marsh is converted to broken marsh and then it declines as broken marsh is converted to open water (Chesney et al., 2000). A temporary increase in marsh-edge perimeter, which occurs in the broken marsh phase, may
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be masking the ultimate effect of habitat loss on landings (Browder et al., 1985, 1989). An alternative hypothesis is that the marsh edge is not the essential habitat per se, but serves as access to flooded marsh, which is essential. However, neither hypothesis may be appropriate for all species, since a variety of species whose microhabitat use patterns often differ, occur in high densities in marshedge and flooded cordgrass habitat types (Zimmerman and Minello, 1984, Rakocinski et al., 1992; Baltz et al., 1993; Minello et al., 1994). These alternative hypotheses are testable in experiments that examine growth and/or survival along the marsh edge by contrasting sites with and without access to flooded marsh.
3.4. Saline marsh food webs Saline marsh food webs are probably far more complex than we recognize. Certainly the fishes and macroinvertebrates that form the upper trophic levels add to this complexity if we take their large variation in size into account. Because many species of fish and macroinvertebrates continue to grow throughout their lives, they can individually function as multiple “species” as their predator–prey relationships and habitat utilization patterns change with ontogeny (Livingston, 1988). The food webs are also complicated by interactions with subtidal ecosystem components. 3.4.1. Species interactions Kneib (1984) highlighted our lack of knowledge about the importance of complex species interactions in saline marsh communities that involve more than two trophic levels. Besides predation, there are many factors acting alone or in concert that may influence the distribution and abundance of invertebrates and their predators, primarily crustaceans, fishes, and birds. These include density-dependent processes, selective larval settlement or mortality, physical gradients that influence habitat selection, and both unpredictable and cyclical physical disturbances. Pennings et al. (2001) compared the palatability of northern and southern populations of smooth cordgrass and saltmeadow cordgrass (Spartina patens) to a variety of grazing insects. In 28 of 32 trials, the insects showed significant preferences for the northern plants and supported the biogeographic hypothesis that lower latitude plants are better defended from herbivory. Whether the preferences are based on toughness, nutrient, and mineral content, or secondary metabolites remains unclear and is likely to vary between plant–herbivore species pairs. Kneib (1991) explored the importance of indirect effects, particularly involving chains of predator–prey interactions, in soft-sediment communities. Indirect effects can influence primary producers, macrofauna, and meiofauna in marsh communities and are often implicated in counterintuitive outcomes of experiments intended to examine direct effects. Soft sediments reduce the feeding efficiency of predators on epibenthic meiofauna (Gregg and Fleeger, 1998). Large crustaceans are important predators in marshes. Grass shrimp have been well studied and are implicated as a connecting link between meiofaunal and nekton communities,
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especially as effective predators that focus their activities around the bases of smooth cordgrass stems (Gregg and Fleeger, 1998). In an elegant field-caging experiment, Silliman and Zieman (2001) demonstrated that periwinkle (Littoraria irrorata) grazing could exert top-down control of smooth cordgrass annual net primary production. The periwinkle is an important grazer in marshes, but it is also an important prey of a large number of predators that forage on marshes, so its ability to control smooth cordgrass is questionable where it suffers normal predation. 3.4.2. Primary producers Stable isotopes are useful for tracing the flow of primary production and nutrients through food webs (Fry, 2006). Stable isotopes of nitrogen, sulfur, and carbon, when used in combination, can greatly increase the power of the isotopic tracer approach in coastal food webs (Peterson et al., 1985) and help address questions about the role of cordgrass marshes in supporting marsh and estuarine consumers. Benthic microalgae and standing dead material may overshadow live cordgrass, macroalgae, and phytoplankton as sources of carbon (Sullivan and Moncreiff, 1990; Currin et al., 1995). Peterson et al. (1986) found that cordgrass detritus and phytoplankton were much more important than upland vegetation and sulfuroxidizing bacteria as carbon sources for marsh macroconsumers. Killifish and mud snails relied more on cordgrass while filter feeders typified by oysters and mussels relied on a combination of cordgrass and plankton. In addition to marsh and estuarine consumers, organic matter from marsh vegetation and estuarine plankton is exported to the offshore environment where it is consumed by marine organisms (Teal, 1962; Odum, 2000). Alber and Valiela (1994) provide evidence that microbial organic aggregates are more important in the nutrition of two marine mussels than particulate detritus or dissolved organic matter. Nitrogen fixation by the community of epiphytes growing on standing dead stems contributes significantly to total nitrogen fixation in marshes where senescent plants are not flattened by ice (Currin and Paerl, 1998). The rate for natural salt marsh is 2.6 g N/m2 stem surface per year and is comparable to sediment rates and about half of rhizospheric nitrogen fixation. Rates are comparable to cyanobacterial mats 2–8 g N/m2/year. While of little direct benefit to cordgrass, nitrogen fixation by epiphytes is important to animals that graze on them and to the nutrient cycle as a source of new biologically available nitrogen. Epiphytes are an important food resource for consumers (Currin et al., 1995). Diverse meiofaunal communities associated with algal epiphytes on smooth cordgrass stems are utilized by shrimp and fishes (Rutledge and Fleeger, 1993; Gregg and Fleeger, 1998). Three salt marsh macroinvertebrates, periwinkles (Littoraria irrorata), salt marsh coffee bean snails (Melampus bidentatus), and talitrid amphipods (Uhlorchestia spartinophila), that feed by shredding dead and senescing smooth cordgrass leaves, also benefit from ingesting fungi; snails, and amphipods also stimulate fungal growth (Grac¸a et al., 2000). These grazing macroinvertebrates contribute to nutrient cycling and connect microbial decomposers to higher-order consumers such as blue crab and Fundulus species that prey on them (Grac¸a et al., 2000).
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3.4.3. Indirect interactions among species A general and realistic qualitative loop model (Levins, 1966; Puccia and Levins, 1991) of a simplified marsh food web (Figure 1a) was compiled based in part on the literature reviewed above to illustrate the importance of indirect interactions. The model includes 20 nodes and covers several trophic levels. In this graphical model, direct positive effects of one node on another are indicated by a link terminating in an arrowhead ("), and direct negative effects are indicated by a link terminating in a filled circle (•). A loop is a path of interactions through nodes (variables) that returns to the node of origin without retracing itself through any node previously encountered. Model input is a community interaction matrix of direct interaction (i.e., –1, 0, or þ1) and output (i.e., the adjoint matrix) is an evaluation of the net number of positive or negative feedback loops influencing each node of the community (Dambacher et al., 2002a,b). Loop models can predict the outcomes of one or more perturbations on a system providing that the community structure is stable (Bender et al., 1984; Dambacher et al., 2002a). A press perturbation (sensu Bender et al., 1984), which may be positive or negative, is a sustained alteration of species densities (or environmental variables), and if the press is maintained, the unperturbed species (or variables) reach a new equilibrium. In essence, the elements of an adjoint matrix are the algebraic summation of the number of positive or negative feedback loops that contribute to the direction of change of a given variable. We have simplified the model output to a graphical representation of a positive press on the variables (Figure 1b). While the analysis of many similar models is necessary to reveal general truths (Levins, 1966) about marsh food webs, we can gain some insights from a single model. Our greatly simplified model with only 20 nodes has several million feedback loops and supports Kneib’s (1991) contention that indirect effects can influence primary producers, macrofauna, and meiofauna in marsh communities and are often implicated in counterintuitive outcomes of experiments intended to examine direct effects. As an illustration, we can revisit Silliman and Zieman’s (2001) cage experiment in which nitrogen and periwinkles (gastropods) were manipulated. In our model, a positive press on nitrogen or periwinkles predicts an enhancement and reduction of Spartina, respectively (Figure 1b). However, the experimental design also excluded major predators on periwinkles. In effect the cage experiment can be interpreted as simultaneous perturbations of nitrogen, periwinkles, and their predators, most importantly blue crabs (Grac¸a et al., 2000). A negative press on periwinkle predators, including blue crabs (–15,157 feedback loops) and raccoons (–13,493), greatly enhances the direct effect of augmenting periwinkles (–26,103) and offsets the enhancement due to nutrients (21,244). The overall effect on Spartina (–33,509) is strongly negative (Table 2). Moreover, the model indicates that virtually any positive or negative press, either alone or in concert, can strongly influence Spartina (Figure 1b). The often overlooked influence of indirect effects can be manifested by perturbations on all nodes in the model, and should be more generally evaluated in nature before assuming bottom-up or top-down control of ecosystems.
Bacteria (a)
Detritus
Fungi
N-fixers
Nutrients
Spartina
Mammals Birds
Phytoplankton
Large fish
Benthic algae Epiphytic algae
Small fish
Herbaceous zooplankton
Crabs
Bivalves
Gastropods
Meiofauna
Predaceous zooplankton Shrimp (b)
Nutrients
Fungi
Bacteria
Detritus
N-Fixers
Spartina
Phyto plankton
Benthic algae
Epiphytic algae
Herbaceous zooplankton
Bivalves
Meiofauna
Shrimp
Predaceous zooplankton
Gastropods
Crabs
Small fishes
Large fishes
Birds
Mammals
Figure 1 (a) A pictorial model of a Spartina marsh food web used to generate a qualitative loop analysis. Connections indicate positive (") and negative (•) direct interactions between nodes and result in many positive and negative feedback loops that amount to indirect interactions. (b) A simplified graphical representation of a positive press on the named variables [e.g., if nutrients are enhanced, five nodes ^ detritus, N-fixers, bivalves, predaceous zooplankton, and crabs ^ show negative responses (, •) and all others are enhanced (þ, *)]. A negative press on a variable of interest can be seen by simply reversing the signs of all other elements.
Nutrients (þ) Nutrients Fungi Bacteria Detritus N-fixers Spartina Phytoplankton Benthic algae Epiphytic algae Herbaceous zooplankton Bivalves Meiofauna Shrimp Predaceous zooplankton Gastropods Crabs Small fishes Large fishes Birds Mammals
54,740 7,559 41,055 –13,685 –13,685 21,244 28,852 13,188 13,188 4,279 –3,452 25,061 11,336 –3,266 25,937 –12,762 21,270 4,254 15,590 9,723
Gastropods (þ) 4,856 –27,317 3,642 –1,214 –1,214 –26,103 29,074 –20,000 –20,000 –11,900 –37,510 25,192 –21,520 13,502 58,276 –12,786 21,310 4,262 –17,258 7,980
Crabs (^) 10,556 –17,796 7,917 –2,639 –2,639 –15,157 14,415 –19,083 –19,083 –21,166 16,033 1,274 7,296 –15,321 43,509 –64,237 –11,867 –38,052 –30,756 –4,695
Mammals (^) –3,832 –12,535 –2,874 958 958 –13,493 –9,424 9,023 9,023 –2,659 2,562 5,689 –32,392 14,620 22,196 58,876 20,802 39,839 7,447 –94,759
Summation 66,320 –50,089 49,740 –16,580 –16,580 –33,509 62,917 –16,872 –16,872 –31,446 –22,367 57,216 –35,280 9,535 149,918 –30,909 51,515 10,303 –24,977 –81,751
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Table 2 Individual positive presses on nutrients and gastropods and negative presses on crabs and mammals result in a mix of responses by Spartina
The summation of all four presses, two positive and two negative, results in an overall reduction of Spartina, and all other variables in the model are also influenced to some degree. This is an extract of the adjoint matrix and negative presses for crabs and mammals were generated by reversing the signs on all variables to reflect the effects of reducing crab and mammal abundances.
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ACKNOWLEDGMENTS NOAA’s Coastal Ocean Program MULTISTRESS Award No. NA16OP2670 supported J. Visser and D. Baltz’s effort to prepare this chapter. This chapter benefited significantly from the extensive critical review by Susan Adamowicz, David Burdick, and John Portnoy.
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Pehrsson, O., 1988. Effects of grazing and inundation on pasture quality and seed production in a salt marsh. Plant Ecol. 74, 113–124. Pennings, S.C., Callaway, R.M., 1992. Salt marsh plant zonation: the relative importance of competition and physical factors. Ecology 73, 681–690. Pennings, S.C., Callaway, R.M., 1996. Impact of a parasitic plant on the structure and dynamics of salt marsh vegetation. Ecology 77, 1410–1419. Pennings, S.C., Siska, E.L., Bertness, M.D., 2001. Latitudinal differences in plant palatability in Atlantic coast salt marshes. Ecology 82, 1344–1359. Pennings, S.C., Selig, E.R., Houser, L.T., Bertness, M.D., 2003. Geographic variation in positive and negative interactions among salt marsh plants. Ecology 84, 1527–1538. Peterson, B.J., Howarth, R.W., Garritt, R.H., 1985. Multiple stable isotopes used to trace the flow of organic matter in estuarine food webs. Science 227, 1361–1363. Peterson, B.J., Howarth, R.W., Garritt, R.H., 1986. Sulfur and carbon isotopes as tracers of saltmarsh organic matter flow. Ecology 67, 865–874. Peterson, G.W., Turner, R.E., 1994. The value of salt marsh edge vs interior as a habitat for fish and decapod crustaceans in a Louisiana salt marsh. Estuaries 17, 235–262. Pettus, D., 1958. Water relationships in Natrix sipedon. Copeia 1958, 207–211. Puccia, C.J., Levins, R., 1991. Qualitative modeling in ecology: loop analysis, signed diagraphs, and time averaging. In: Fishwick, P.A., Lucker, P.A. (Eds.), Qualitative Simulation Modeling and Analysis. Springer-Verlag, New York, pp. 119–143. Quintana, X.D., Moreno-Amich, R., 2002. Phytoplankton composition of Emporda salt marshes, Spain and its response to freshwater flux regulation. J. Coast. Res. 36, 581–590. Rakocinski, C.F., Baltz, D.M., Fleeger, J.W., 1992. Correspondence between environmental gradients and the community structure of marsh-edge fishes in a Louisiana estuary. Mar. Ecol. Prog. Ser. 80, 135–148. Ranwell, D.S., 1961. Spartina salt marshes in southern England: I. The effects of sheep grazing at the upper limits of Spartina marsh in Bridgwater Bay. J. Ecol. 49, 325–340. Rozas, L.P., Odum, W.E., 1988. Occupation of submerged aquatic vegetation by fishes: testing the roles of food and refuge. Oecologia 77, 101–106. Rutledge, P.A., Fleeger, J.W., 1993. Abundance and seasonality of meiofauna, including harpacticoid copepod species, associated with stems of the salt-marsh cord grass, Spartina alterniflora. Estuaries 16, 760–768. Shanholtzer, G.F., 1974. Relationship of vertebrates to salt marsh plants. In: Reimold, R.J., Queen, W.H. (Eds.), Ecology of Halophytes. Academic Press, New York, pp. 463–474. Silliman, B.R., Bertness, M.D., 2002. A trophic cascade regulates salt marsh primary production. Proceedings of the National Academy of Sciences, USA. 99, 10500–10505. Silliman, B.R., Zieman, J.C., 2001. Top-down control of Spartina alterniflora production by periwinkle grazing in a Virginia salt marsh. Ecology 82, 2830–2845. Sullivan, M.J., Currin, C.A., 2000. Community structure and functional dynamics of benthic microalgae in salt marshes. In: Weinstein, M.P., Kreeger, D.A. (Eds.), Concepts and Controversies in Tidal Marsh Ecology. Kluwer Academic Publishers, Dordrecht, pp. 81–106. Sullivan, M.J., Moncreiff, C.A., 1990. Edaphic algae are an important component of salt marsh foodwebs: evidence from multiple stable isotope analyses. Mar. Ecol. Prog. Ser. 62, 149–159. Talley, T.S., Levin, L.A., 2001. Modification of sediments and macrofauna by an invasive marsh plant. Biol. Invasions 3, 51–68. Teal, J.M., 1962. Energy flow in the salt marsh ecosystem of Georgia. Ecology 43, 614–624. Thannheiser, D., Holland, P., 1994. The plant communities of New Zealand salt meadows. Glob. Ecol. Biogeogr. Lett. 4, 107–115. Turner, R.E., 1977. Intertidal vegetation and commercial yields of penaeid shrimp. Trans. Am. Fish. Soc. 106, 411–416. Turner, M.G., 1987. Effects of grazing by feral horses, clipping, trampling, and burning on a Georgia salt marsh. Estuaries 10, 54–60. Turner, R.E., 1992. Coastal wetlands and penaeid shrimp habitat. In: Stroud, R.H. (Ed.), Stemming the Tide of Coastal Fish Habitat Loss. Proceedings of the 14th Annual Marine Recreational Fisheries Symposium Baltimore, Maryland. Mar. Recreational Fish. 14, 97–104.
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Vale´ry, L., Bouchard, V., Lefeuvre, J.C., 2004. Impact of the invasive native species Elymus athericus on carbon pools in a salt marsh. Wetlands 24, 268–276. van der Wal, R., Egas, M., van der Veen, A., Bakker, J., 2000. Effects of resource competition and herbivory on plant performance along a natural productivity gradient. J. Ecol. 88, 317–330. Van Dolah, R.F., 1978. Factors regulating the distribution and population dynamics of the amphipod Gammarus palustris in an intertidal salt marsh. Ecol. Monogr. 48, 191–217. van Wijnen, H.J., Bakker, J.P., 1999. Nitrogen and phosphorus limitation in a coastal barrier salt marsh: the implications for vegetation succession. J. Ecol. 87, 265–272. Visser, J.M., Sasser, C.E., Cade, B.S., 2006. The effect of multiple stressors on salt marsh end-ofseason biomass. Estuar. Coast. 29, 331–342. Weinstein, M.P., 1979. Shallow marsh habitats as primary nurseries for fishes and shellfish, Cape Fear River, North Carolina. Fish. Bull. 77, 339–357. Zedler, J.B., 1982. Salt marsh algal mat composition: spatial and temporal comparisons. Bull. South. Calif. Acad. Sci. 81, 41–50. Zedler, J.B., Callaway, J.C., Desmond, J.S., Vivian-Smith, G., Williams, G.D., Sullivan, G., Brewster, A.E., Bradshaw, B.K., 1999. Californian salt-marsh vegetation: an improved model of spatial pattern. Ecosystems 2, 19–35. Zimmerman, R.J., Minello, T.J., 1984. Densities of Penaeus aztecus, Penaeus setiferus, and other natant macrofauna in a Texas salt marsh. Estuaries 7 (4A), 421–433. Zimmerman, R., Minello,T., Klima, E., Nance, J., 1991. Effects of accelerated sealevel rise on coastal secondary production. In: Bolten, H.S., Magoon, O.T. (Eds.), Coastal Wetlands, Coastal Zone ’91 Conference. American Society of Civil Engineers, New York, pp. 110–124.
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C H A P T E R
1 6
S ALT M ARSH B IOGEOCHEMISTRY – A N O VERVIEW Craig Tobias and Scott C. Neubauer
Contents 1. Introduction 2. Carbon 2.1. Exchanges 2.2. Internal cycling 2.3. Burial 3. Nitrogen 3.1. Exchanges 3.2. Internal cycling 3.3. Burial 4. Iron and Sulfur 4.1. Exchanges 4.2. Internal cycling 4.3. Burial 5. Phosphorus 5.1. Exchanges 5.2. Internal cycling 5.3. Burial 6. Marshes in Transition and Directions for Future Work Acknowledgments References
445 446 446 449 455 455 455 462 464 465 465 468 471 471 472 475 477 477 478 479
1. INTRODUCTION Salt marshes have long been considered important sources, sinks, and/or transformers of biologically important nutrients in the coastal landscape. Macrophyte production contributes autochthonous organic matter to soils. Marsh geomorphology and hydrodynamics lead to trapping of allochthonous organic and mineral particulates. The high rates of respiration in marsh soils create an electron-rich, chemically reduced environment proximal to oxidized surface waters. The subsurface redox environment is further modified by root-mediated Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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release of O2 into the rhizosphere, physical mixing by macro-organisms, and water movements caused by tidal infiltration and drainage, factors that vary over tidal to seasonal and interannual scales. As such, marshes provide highly reactive surfaces that modify water quality. High organic matter production rates, variable redox environments, and physically dynamic ecosystems that are regularly pulsed by tides are characteristic features of salt marshes, mangrove swamps, tidal freshwater wetlands, and seagrass systems. Salt marshes differ considerably from nonvegetated mudflats in terms of overall autotrophic productivity and impacts of macrophyte root production on the cycling of macroelements. While there are many commonalities between saline and tidal freshwater wetlands, the chemistry of seawater (notably the presence of sulfate) leads to significant biogeochemical differences in processes such as organic carbon (C) oxidation and the cycling of nitrogen (N), sulfur (S), iron (Fe), and phosphorus (P). The goal of this chapter is straightforward. We present an overview of the processing and cycling of the major biologically relevant elements (C, N, P, S, and Fe) in salt marshes. In keeping with the ecosystems theme of this book, and to the extent possible, the cycles are parsed into exchanges, pathways of internal cycling, and burial for each element.
2. CARBON 2.1. Exchanges 2.1.1. Photoautotrophy Marsh macrophytes serve as a dominant source of new C to marshes, play a key role in stabilizing marsh platforms, trapping sediments, aerating the soil through root O2 loss (ROL), and influencing biogeochemical cycling (Bodelier, 2003; Hines, 2006). Rates of net primary production (NPP) in Spartina alterniflora marshes range from 100 to >2,500 g/cm2/year (Mitsch and Gosselink, 1993; Dame et al., 2000), with highest values in south Atlantic and Gulf coast marshes (Mendelssohn and Morris, 2000). Belowground biomass accumulation often equals or exceeds that of aboveground tissues (Valiela et al., 1976; Schubauer and Hopkinson, 1984; Darby and Turner, 2008). The productivity of other marsh plants including Spartina patens, Distichlis spicata, and Juncus roemerianus can be comparable to that of S. alterniflora (Mitsch and Gosselink, 1993). NPP can vary within a single marsh as a function of anoxia, sulfide, and salinity stresses that affect nitrogen uptake and assimilation (King et al., 1982; Mendelssohn and Morris, 2000). Further, sea level anomalies lead to high interannual variability by affecting flooding frequency and thus the salinity of marsh porewaters (Morris, 2000; Morris et al., 2002). The majority of C fixed by plants is atmospheric CO2 although small amounts (<10% of atmospheric fixation) can come from porewater dissolved inorganic C (DIC) or CO2 that is recycled in lacunar spaces (Hwang and Morris, 1992). Relative to rates of macrophyte productivity, less is known about primary productivity by salt marsh benthic microalgae and macroalgae. Benthic microalgal
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Salt Marsh Biogeochemistry
Table 1 Net benthic microalgal (BMP) and vascular plant production (VPP) in tidal salt marshes, and the ratio of BMP to net primary production (NPP), where NPP = BMP þ VPP State, USA
Productivity (g C/m2/year) Microalgal a
Vascular plants
BMP NPP
Massachusetts New York Delaware
53 50 61–99
212a 292 NR
0.20 0.15 0.25
Virginia Virginia Virginia South Carolina Georgia
27.8 67b 235 98, 234c
254 161 831 534, 296c
0.10 0.29 0.22 0.16, 0.44c
180
732
0.20
Georgia Mississippi
208 28–151
1127 248–742a
0.16 0.90–0.38
Texas California
71 185–341
550–900 243–340
0.90–0.11 0.43–0.58
Reference
Van Raalte et al., 1976 Woodwell et al., 1979 Gallagher and Daiber, 1974 Anderson et al., 1997 Miller et al., 2001 Buzzelli et al., 1999 Pinckney and Zingmark, 1993 Pomeroy, 1959; Teal, 1962 Pomeroy et al., 1981 Sullivan and Moncreiff, 1988 Hall and Fisher, 1985 Zedler, 1980
Note that this calculation differs from the oft-reported ratio that expresses BMP as a fraction of plant production (i.e., BMP/VPP). Plant production was estimated from aboveground biomass in all studies except Miller et al. (2001) where a CO2-based gas flux model was utilized. NR, not reported. a Converted from original estimates, assuming that algal and/or plant biomass is 50% C. b BMP was measured in cleared zones within a dense Spartina patens/Distichlis spicata canopy; rates in these cleared zones likely overestimate rates beneath the canopy (Sullivan and Daiber, 1975). c Averages for tall and short Spartina alterniflora zones, respectively.
productivity ranges from 30 to 300 g C/m2/year (Table 1) and shows a similar latitudinal pattern as for macrophytes. The highest rates are observed along the southeast Atlantic and California coasts (Zedler, 1980). Benthic microalgal production generally accounts for 10–25% of total ecosystem (i.e., macrophyte þ algal) productivity but sometimes exceeds plant productivity (Table 1). Microalgal productivity is often higher during winter when the plant cover is at a minimum. Despite lower rates of productivity, benthic microalgal biomass is generally more labile than that of macrophytes and is preferentially assimilated by secondary consumers (Sullivan and Moncreiff, 1990; Currin et al., 1995; Sullivan and Currin, 2000). 2.1.2. Allochthonous C deposition Tidal flooding provides a mechanism for the delivery and deposition of water column suspended sediments and associated C. Rates of sedimentation depend on suspended sediment concentrations, tidal range, vegetation, creek proximity, and hydroperiod (Friedrichs and Perry, 2001). Thus, sedimentation and organic C deposition vary widely from marsh to marsh, laterally within a single site, and temporally within the year. As an example, deposition rates in Paulina Marsh, the
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Netherlands, ranged from <100 g/m2/spring–neap cycle near the marsh-upland border to 2,000 g/m2/cycle at the marsh–mud flat boundary (Temmerman et al., 2003). Across the entire marsh, deposition averaged 200 g/m2/cycle. Assuming an organic C content of 5–10% for suspended sediments (Middelburg and Herman, 2007), deposition across this marsh delivers 250–500 g C/m2/year. This rate falls within the range of deposition reported for other salt marshes (Cahoon and Reed, 1995; Salgueiro and Cac¸ador, 2007) and for tidal freshwater marshes (Chapter 19). Thus, sedimentation represents a source of C to salt marshes comparable to C fixation by benthic microalgae and chemoautotrophs (Section 2.2.3). Such inputs of allochthonous C have been suggested as a mechanism to explain the 9–12‰ depletion in soil organic matter d 13C values (relative to S. alterniflora biomass) that is often observed in mineral-dominated marshes (Middelburg et al., 1997). Marshes in sediment-starved regions or those that are infrequently flooded by tides will have lower inputs of allochthonous C relative to autochthonous C sources. Considerable quantities of sediment-associated C can be delivered by hurricanes and other large storms (Parsons, 1998; Turner et al., 2006a). Hurricanes Katrina and Rita (Gulf of Mexico coast, USA, August/September 2005) deposited 22.3 kg sediment/m2 (Turner et al., 2006a). Assuming that these sediments originated from offshore (as hypothesized by Turner et al., 2006a) with an average organic C content of 2.2% (Mayer et al., 2007), this deposition delivered 490 g C/m2 to the marsh surface. The mineral fraction can contribute to significant vertical marsh growth (Turner et al., 2006a), but the fate of the organic fraction is currently unknown. 2.1.3. Organic C export The exchanges of C and nutrients between marshes and tidal waters have been studied for decades, often in the context of the outwelling hypothesis (Kalber, 1959; Odum, 1968; see reviews by Nixon, 1980; Dame, 1994; Childers et al., 2000). High variability in terms of hydrology, basin age, and geomorphological setting, among other features, makes it a futile task to draw broad conclusions with respect to the direction and magnitude of particulate and dissolved organic C (POC and DOC) fluxes that will apply to all marshes at all times. Childers et al. (2000) summarized the salt marsh flux literature that has appeared since Nixon’s (1980) influential review of marsh–estuarine exchanges. In their compilation, three of eight studies showed a net annual POC export, with rates of 11–128 g C/m2/ year (the other five studies had POC imports of 3–140 g C/m2/year). In the North Inlet (South Carolina) system, the entire 3,200 ha marsh basin exported POC, whereas the geologically young Bly Creek sub-basin (66 ha) imported POC (Dame et al., 1986, 1991). In contrast to the high variability in the rates and direction of POC exchange, 11 of the 13 studies reporting DOC fluxes showed net DOC export (15–328 g C/m2/year versus <15 g C/m2/year for the two studies that showed DOC import). In North Inlet, the seepage and drainage of DOC-rich marsh porewaters is more important during summer, whereas external inputs related to freshwater discharge play a larger role during winter (Wolaver et al., 1986).
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449
There is some evidence that transient biota (e.g., fishes) can mediate significant C (and N) export from marshes. This mechanism is beyond the scope of this chapter (but see Deegan, 1993; Deegan et al., 2000; Teal and Howes, 2000). 2.1.4. Inorganic C export Mass balance calculations indicate that a significant fraction of marsh primary production (>90%) must be decomposed or otherwise exported in order to explain long-term C accumulation rates (Howes et al., 1985; Gardner, 1990; Middelburg et al., 1997). Gaseous CO2 emissions from salt marshes range from 240 to 720 g C/m2/year (Blum et al., 1978; Howes et al., 1985; Morris and Whiting, 1986; Morris and Jensen, 1998; Miller et al., 2001) with an additional loss of dissolved inorganic C of 120–240 g C/m2/year (Howes et al., 1985; Morris and Whiting, 1986; Nietch, 2000; Wang and Cai, 2004). Across saline and brackish marshes, the export of DIC accounts for 20–30 of total inorganic C losses (i.e., CO2 þ DIC). Tidal marshes can export significant amounts of DIC to adjacent coastal waters, influence the apparent metabolic state of these waters (i.e., net autotrophic vs. heterotrophic), and affect the magnitude and direction of CO2 exchange between coastal waters and the atmosphere (Cai and Wang, 1998; Cai et al., 1999, 2000; Raymond et al., 2000; Neubauer and Anderson, 2003; Wang and Cai, 2004; Borges, 2005). Similar estimates for tidal freshwater marshes are reported in this book (Chapter 19). The CO2 and DIC that accumulate (and can be lost) from salt marsh soils can be derived from the decomposition of plant roots or bulk soil organic matter. 13C isotopes indicate that root respiration can account for 21–90% of total soil respiration (Wang et al., in review.), with a higher fractional contribution of roots in more organic marshes. The mineralization of the large pool of soil organic matter can proceed at appreciable rates, even many years after in situ plant primary production has stopped (Morris and Whiting, 1986; Wang et al., in press.). Emissions of CH4 from salt marshes are generally lower than fluxes from freshwater wetlands (Bartlett et al., 1987; Bridgham et al., 2006). In a compilation of flux data from North American wetlands, the average CH4 flux from freshwater wetlands was 36.0 + 5.0 g C/m2/year (average + SE), whereas that from salt marshes was 3.6 + 2.3 g C/m2/year (Bridgham et al. 2006). Sulfate reduction coupled to anaerobic CH4 oxidation (Martens and Berner, 1977; Boetius et al., 2000) would further reduce CH4 emissions, but this process has not been explored in salt marsh soils.
2.2. Internal cycling 2.2.1. Aerobic mineralization Molecular oxygen is rapidly used in salt marsh soils as an electron acceptor for microbial respiration and an oxidant for reduced chemical species. Marsh O2 uptake (maximal at low tide) ranges from 5 to 65 mol O2/m2/year (Howes et al., 1984; Howarth, 1993; Cai et al., 1999). Many plants including S. alterniflora, Spartina anglica, and J. roemerianus can oxidize subsurface soils through ROL
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(Mendelssohn et al., 1995; Holmer et al., 2002; Maricle and Lee, 2002; Koretsky and Miller, 2008). The amount of O2 uptake due to advective air movement into drained pore spaces can be 50% of the diffusive O2 uptake across the marsh atmosphere boundary (Morris and Whiting, 1985). Filling pore spaces with fully aerated tidal water (rather than air) contributes a much smaller amount of O2 due to the lower concentration of O2 in water and decreased diffusion in water. Infauna can increase soil oxidation through burrowing, bioirrigation, and physical mixing of the substrate (Meile et al., 2001; Gribsholt and Kristensen, 2002; Kristensen and Kostka, 2005). Because there are no in situ methods for quantifying O2 added during ROL or faunal activity, estimates of aerobic respiration are likely to be seriously underestimated if only O2 fluxes across the marsh surface are considered. Howarth (1993) estimated that aerobic respiration accounted for 18–30% of total soil respiration but diagenetic simulation results suggest a much smaller contribution (1–5% of the total organic matter decomposition; Furukawa et al., 2005). Because organic matter can persist above the soil surface as standing dead stems and wrack, measurements of soil processes alone will underestimate the total role of aerobic decomposition. Fungi and bacteria carry out a significant amount of C mineralization subaerially and under aerobic conditions. For example, peak microbial respiration rates on senescent leaves of tall S. alterniflora coincided with peak fungal biomass (Buchan et al., 2003). Fungal degradation can remove up to 60% of the original aboveground organic matter (Newell and Porter, 2000) with aerobic degradation continuing to occur as the bacterial standing crop increases during later stages of decomposition (Benner et al., 1986; Newell et al., 1989). 2.2.2. Anaerobic mineralization Thermodynamic theory indicates that the availability of electron acceptors and competition between microbes for electron donors will govern the relative importance of different catabolic processes (Megonigal et al., 2004). Respiration using O2 as the electron acceptor has the highest energy yield. Following the depletion of O2, a predictable sequence of anaerobic processes follows: denitrification, Mn reduction, Fe(III) reduction, SO2 4 reduction, and methanogenesis (Ponnamperuma, 1972). There is some recent evidence that humic acids can also be significant electron acceptors. One result of the competition among microbes for electron donors can be vertical redox stratification that reveals itself as gradients in solid phase or porewater geochemistry (Griffin et al., 1989; Taillefert et al., 2007). However, due to fine-scale heterogeneity in distributions of electron donors and electron acceptors, multiple pathways can coexist within the same volume of soil (Højberg et al., 1994). Despite the importance of denitrification for NO 3 removal (Section 3.1.5), denitrification is not significant for organic C turnover in marshes due to relatively low NO 3 supply. It accounts for 1% of C mineralization in salt marsh soils (Table 2). Additions of NO 3 to Georgia marsh soils did not increase total C mineralization rates (Hyun et al., 2007). To our knowledge, Mn(III, IV) reduction rates have not been measured in salt marshes. However, voltametric (microelectrode) studies have identified dissolved Mn2þ in vegetated soils (Brendel and Luther, 1995) and in nonvegetated intertidal
Anaerobic Total
Sippewissett (MA) Jack Bay (MD) Skidaway Island (GA)
Sapelo Island (GA)
% of anaerobic respiration
Citation
NO 3 reduction
Metal reduction
SO2 4 reduction
Methanogenesis
0.82
0.2
“Negligible”
99.4
0.4
Howarth, 1993
ND ND
ND ND
2.7–51.4a 0–109b
48.6–95.2a 6–82b
0–2.0a ND
Neubauer et al., 2005b Kostka et al., 2002a
ND 0.95–0.99d ND 0.70
ND 0.9–1.0d ND 1.1
28–96c 1.7–61.6d 0–71.8e “Negligible”
4–72c 36.8–97.5d 22.1–95e 94.4
ND ND ND 4.4
Gribsholt et al., 2003 Furukawa et al., 2004 Hyun et al., 2007 Howarth, 1993
ND
ND
NDf
0–106f
ND
Kostka et al., 2002b
Salt Marsh Biogeochemistry
Table 2 Importance of anaerobic to total (aerobic þ anaerobic) metabolism in several salt marshes and the partitioning of anaerobic C decomposition through various pathways
The relative importance of each pathway is taken directly from each citation or calculated after summing the measured pathways and assuming that unmeasured pathways did not significantly contribute to total anaerobic metabolism. This is not always a valid assumption since the sum of individual metabolic rates does not always equal total CO2 þ CH4 production. Mn reduction rates have not been measured in salt marsh soils and sediments, so “metal reduction” refers to biological Fe(III) reduction only. “Negligible” is a quote from Howarth (1993). “ND” indicates that aerobic respiration was not reported or that rates for a specific metabolic pathway were not measured. a Range due to monthly variability in rates measured in June, July, and August. b Range due to spatial variability between a bioturbated, vegetated (Spartina alterniflora) levee site and an unvegetated, nonbioturbated creekbank. c Range due to spatial variability with increasing distance from Uca pugnax burrows (0–35 cm from burrow wall), as well as variations between rhizosphere and levee bulk soils. d Range due to spatial variability in modeled rates (integrated to 9 cm depth) between locations that are vegetated with S. alterniflora, containing abundant bioturbating fiddler crabs (U. pugnax) but no plants, and are unvegetated and lack U. pugnax. e Range due to lateral variability between tall S. alterniflora, short S. alterniflora, and creekbank zones, as well as depth-related differences (0–3 cm and 3–6 cm) within each zone. f Range due to lateral variability between a bioturbated unvegetated creekbank, a bioturbated vegetated (S. alterniflora) levee, and S. alterniflora mid-marsh zones, as well as depth-related differences (0–5 cm and 10–15 cm) within each zone. In this study, Fe(III) reduction was measured. Although no distinction was made between chemical and biological Fe(III) reduction, Kostka et al. (2002b) suggested that abiotic Fe(III) reduction dominated.
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creek bank sediments (Taillefert et al., 2007), indicating that Mn(III, IV) reduction was occurring. Based only on the presence/absence of reduced Mn, it is not possible to determine whether the Mn reduction was driven by biological (enzymatic) or strictly chemical reactions. Regardless, Mn(III, IV) reduction will probably be less important than Fe(III) reduction with respect to C mineralization since Mn is generally present at lower concentrations than Fe (Luther et al., 1992). However, Mn reduction can be a dominant pathway in nearsurface subtidal marine sediments (Canfield et al., 1993; Thamdrup, 2000). Iron(III) reduction can be an important pathway for organic matter turnover in salt marsh soils, accounting for 50–100% of anaerobic metabolism (Table 2). The relatively recent recognition that metal reduction can be a significant biological process contrasts with the historical view that SO2 4 reduction dominates anaerobic metabolism in salt marshes (Howarth and Hobbie, 1982; Howarth, 1993). Many early studies were conducted in short S. alterniflora or mid-marsh habitats where H2S accumulates due to low porewater turnover rates. Under these conditions, SO2 reduction should dominate metabolism and Fe(III) reduction will be a 4 primarily abiotic process where Fe oxides serve as an oxidant for sulfides: 8FeOOH þ H2 S þ 14Hþ ! 8Fe2þ þ SO2 4 þ 12H2 O
ð1Þ
Biological Fe(III) reduction will be more important than chemical reduction when amorphous Fe(III) oxides are plentiful and continually regenerated, or H2S production is low relative to the Fe(III) concentration (Jacobson, 1994). Indeed, marsh zones with heavy bioturbation activity (especially by fiddler crabs, Uca spp.) can have anaerobic metabolism dominated by biotic Fe(III) reduction (Gribsholt et al., 2003; Furukawa et al., 2004; Hyun et al., 2007). Plant-driven inputs of O2 (via ROL) and organic C can also lead to significant rates of C turnover coupled to Fe(III) reduction (Neubauer et al., 2005b) although this mechanism appears to be less important than bioturbation in some salt marsh soils (Furukawa et al., 2004). Further, Fe(III) oxides formed in the rhizosphere can be more amorphous (a factor that favors enzymatic Fe(III) reduction) than those in bulk soil (Weiss et al., 2004). Gribsholt and Kristensen (2002) suggested that the highest rates of organic matter decomposition and Fe(III) reduction will occur in substrates that are bioturbated (to regenerate Fe(III) oxides) and vegetated (with plants acting as an organic C source). Hydrology, including diurnal tidal cycles (Taillefert et al., 2007) and seasonal changes in tidal water level (Neubauer et al., 2005b) or groundwater (GW) discharge (Tobias et al., 2001b), may also be important in driving O2 penetration into marsh soils and sediments, regulating Fe(III) regeneration, and leading to significant rates of biotic Fe(III) reduction. Humic acids can be used as terminal electron acceptors for anaerobic metabolism with reasonable thermodynamic efficiency (Cervantes et al., 2000), in addition to their role as electron shuttles for Mn(III, IV) and Fe(III) reduction (Lovley et al., 1996). To date, the potential role of humic acid reduction as a pathway for organic C oxidation has not been directly measured but inferred from an inability to completely account for total rates of microbial respiration (i.e., CO2 þ CH4
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production). For example, the sum of individually measured anaerobic pathways accounted for only 43% of total metabolism in a Georgia salt marsh (Hyun et al., 2007) and 20–30% in a Maryland brackish marsh (Neubauer et al., 2005b). SO2 reduction is often a dominant pathway for anaerobic C metabolism 4 (Table 2; Howarth and Hobbie, 1982; Howarth, 1984; Hines et al., 1989; Kostka et al., 2002b). SO2 4 concentrations typically exceed several millimolar in flooding water so that SO2 4 availability will not limit reduction (Boudreau and Westrich, 1984; Roychoudhury et al., 2003b). Instead, competition from other electron accepting processes and the availability of labile organic C regulate the relative 2 importance of SO2 4 reduction. Based on considerations presented earlier, SO4 reduction should have the largest contribution to total anaerobic C metabolism in soils that are unvegetated, largely undisturbed (i.e., low bioturbation), and/or poorly flushed (allowing H2S to accumulate and minimizing advective O2 delivery) – this has been demonstrated in several recent studies (Kostka et al., 2002a,b, Furukawa et al., 2004). Further, SO2 reduction rates can be limited by an 4 imbalance between the temperature responses of microbes that hydrolyze and ferment complex organic substrates and SO2 4 reducers that utilize the resulting simple organic molecules (Weston and Joye, 2005). At low temperatures, the production of low-molecular weight dissolved organic carbon exceeds its consumption, resulting in accumulation. At high temperatures, however, SO2 4 reduction becomes limited by the rate at which low-molecular weight electron donors are generated (Weston and Joye, 2005) and by feedbacks from the accumulation of respiratory end products. There has been little work on rates of CH4 formation in salt marshes although existing evidence suggests that methanogenesis accounts for less than 5% of anaerobic respiration (Table 2). The paradigm is that methanogenesis in salt marshes is limited because methanogens are out-competed for electron donors (King and Wiebe, 1980a) although CH4 production can be substantial in some humus-rich salt marsh soils (Giani et al., 1996). In a comparison of tall and short S. alterniflora zones, King and Wiebe (1980b) found that rates of CH4 production were considerably greater in the short Spartina zone. At the same time, SO2 4 was relatively more depleted in the short Spartina zone, which allowed methanogenesis to increase in importance (King and Wiebe, 1980b). Given higher Fe(III) reduction in creek bank Spartina zones (Furukawa et al., 2004; Hyun et al., 2007), electron donor availability to methanogens may be limited by metal reducers along creek banks and SO2 4 reducers in the marsh interior. Despite the thermodynamic constraints against methanogenesis, some C are mineralized through this process, an indication of microzones in the soil matrix that is depleted in other electron acceptors and/or regions where substrates exist that can be converted to CH4 but cannot be used by SO2 4 reducers (e.g., methanol and methylated amines; Oremland et al., 1982). 2.2.3. Chemoautotrophy Reduced compounds such as Fe(II) and H2S contain considerable energy that can be released upon oxidation. When this oxidation occurs through a chemoautotrophic microbial process, the energy in the chemical bonds can be utilized to fix
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porewater DIC in marsh soils. Thus, from a C cycling perspective, this C fixation represents primary production. However, from an energetic perspective, chemoautotrophy is a form of secondary production (after Howarth, 1993; Chapin et al., 2006) since the chemical energy in the inorganic compounds was originally released from organic matter (or H2). Sulfides produced through SO2 4 reduction are rapidly incorporated into minerals such as iron monosulfides (FeS) and pyrite (FeS2). Given that >70% of the energy in organic compounds utilized by SO2 4 reducers ends up in these reduced sulfur compounds, it is apparent that FeS, FeS2, and H2S in marsh substrates represent a considerable energy source. Mass balance calculations indicate that >90% of the sulfides are ultimately reoxidized or exported (Howes et al., 1984; Gardner, 1990). Rates of S-mediated chemoautotrophy may range from 20 to 480 g C/m2/year (Howarth and Teal, 1980; Howarth, 1984, 1993), a rate that is comparable to benthic microalgal production. However, if these sulfides are oxidized chemically, the energy is diverted from chemoautotrophic production. Lithoautotrophic Fe(II)-oxidizing bacteria (FeOB) exist in salt marsh (Weiss et al., 2003) and subtidal marine environments (Emerson and Moyer, 2002; Edwards et al., 2003). Below, we calculate Fe(II) supply rates from the turnover of Fe–S minerals and oxidized Fe(III) to estimate potential rates of chemoautotrophic production by FeOB. In S. alterniflora soils, rates of SO2 reduction 4 typically range from 15 to 75 mol S/m2/year (daily rates extrapolated to an annual basis; see Table 2 of Kostka et al., 2002b). Since the majority of the iron sulfides produced by SO2 4 reduction are eventually oxidized, we estimate that 7–75 mol Fe(II)/m2/year are generated due to the turnover of reduced S compounds. In a Georgia salt marsh, Kostka et al. (2002a) reported Fe(III) reduction rates (and therefore Fe(II) production rates) of 0–190 mol Fe(II)/m2/year (extrapolated from daily rates). Combining these estimates of Fe(II) supply due to Fe–S mineral turnover and Fe(III) reduction, we estimate a total supply of Fe(II) of 7–265 mol Fe(II)/m2/year. If this Fe(II) is oxidized microbially with a growth yield of 0.7 g C/mol Fe(II) oxidized (Neubauer et al., 2002b), chemoautotrophic Fe(II) oxidation could contribute a maximum of 2–185 g C/m2/year of new production. However, the actual contribution of FeOB to chemoautotrophic production is likely to be considerably lower since FeOB are a very small component (<0.01%) of the total salt marsh microbial community (Weiss et al., 2003) and account for 50% or less of total Fe(II) oxidation (Neubauer et al., 2002b, 2007). 2.2.4. Carbonate mineral formation In most marine environments, Fe2þ rapidly precipitates with sulfides (Section 4.2.2); however, large concretions containing siderite (FeCO3) have been found in some salt marshes (Pye, 1981). The d13C evidence suggests that these carbonates are partially derived from the degradation of marsh organic matter (Pye et al., 1990), with siderite more likely to form in saline/brackish systems when rates of Fe(III) reduction are high relative to SO2 4 reduction due to low organic C availability (Adams et al., 2006). Under such conditions or when biological Fe(III) reduction is favored over chemical Fe(III) reduction (see above; Jacobson, 1994), low sulfide
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concentrations mean that Fe2þ and Mn2þ are more likely to complex with CO2 3 than with sulfides (Berner, 1969; Pye et al., 1990; Adams et al., 2006). However, Giblin and Howarth (1984) reported that porewaters of Great Sippewissett Marsh (Massachusetts) were highly undersaturated with respect to both siderite and rhodocrosite (MnCO2), providing strong evidence that formation of these minerals is limited.
2.3. Burial As determined using 137Cs and 210Pb radiotracers, rates of salt marsh C accumulation ranged from 100–200 g C/m2/year over the last 50–100 years (Turner et al., 2000; Chmura et al., 2003; Bridgham et al., 2006; Craft, 2007). This range is comparable to that for tidal freshwater marshes (Craft, 2007; Megonigal and Neubauer, 2009; Neubauer, 2008) although C accumulation may be lower in salt marshes than in tidal freshwater and brackish marshes (Craft, 2007). The organic matter that is buried long-term likely represents a combination of in situ production (litter, detritus, aquatic, and subsurface roots) and allochthonous organic materials associated with mineral sediments (Wolaver et al., 1988; Craft et al., 1993; Rooth et al., 2003; Nyman et al., 2006; Neubauer, 2008). Average accretion and accumulation rates decrease over the first 1,000 years following burial due to decomposition and compaction of deeper soil horizons (Neubauer et al. 2002a, Turner et al. 2006b). Ultimately, the magnitude of C (as well as N and P) burial depends on long-term marsh responses to rising sea level.
3. NITROGEN The total salt marsh N pool varies with marsh age and can be 5–30 times larger than the sum of all N cycling reactions within a year, leading to long turnover times for marsh N (Anderson et al., 1997; Rozema et al., 2000). The largest marsh N pool is in bulk soils (200–1,000 g N/m2; to a 30 cm depth), with plant biomass containing 1–22 g N/m2 (Hopkinson and Schubauer, 1984; DeLaune and Patrick, 1990; Morris, 1991; Anderson et al., 1997; Rozema et al., 2000; Hopkinson and Giblin, 2008). In porewaters, dissolved organic nitrogen (DON) and NHþ 4 are the dominant species (concentrations often 100 mM) while NO concentrations are generally low 3 ( 10 mM). Despite the smaller overall inventory of porewater DIN, it is a common intermediate reservoir for the numerous internal N cycling reactions.
3.1. Exchanges 3.1.1. N fixation N fixation may be important for building N stocks in young marshes, but its importance generally lessens as marshes trap more exogenous N and increase internal recycling. In mature marshes, N fixation ranges from <0.5 to 6.8 g N/m2/year
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(Teal et al., 1979; Rozema et al., 2000; Moisander et al., 2005). It is usually a small part (10%) of the total N inputs and is an order of magnitude lower than internal N recycling rates (Valiela and Teal, 1979; Anderson et al., 1997; Rozema et al., 2000). In young or restored marshes, rates can exceed 35 g N/m2/year and be of sufficient magnitude to support macrophyte N demand (Currin et al., 1996; Tyler et al., 2003). N fixation is concentrated in the marsh surface and performed principally by hetero- and nonheterocystous cyanobacteria (Currin and Paerl, 1998a,b; Piehler et al., 1998). N fixation patterns represent the sum total effects of irradiance, cyanobacterial diversity, tidal inundation, and O2 inhibition of nitrogenase (Ubben and Hanson, 1980; Joye and Paerl, 1994; Currin et al., 1996; Moseman, 2007). 3.1.2. Atmospheric deposition Atmospheric deposition, which is typically dominated by NO 3 and DON, can also occur as NHþ and particulate N (PN) (Russell et al., 1998; Paerl et al., 2002). The 4 dissolved inorganic N (DIN) concentrations in precipitation are comparable to those in coastal tidal waters (Russell et al., 1998; Paerl et al., 2002). However, because precipitation comprises only a small part of the marsh water budget (Morris, 1991; Lent et al., 1997), atmospheric deposition rates are a small part of the N budget. Along the Atlantic coast (USA), atmospheric deposition (<0.5–1.2 g N/m2/year) is 15% of the water column inputs (Valiela et al., 1978; Morris, 1991; Paerl et al., 2002; Seitzinger et al., 2002; Buzzelli, 2008); higher loadings (3 g N/m2/year) are seen in marshes of northwestern Europe (Rozema et al., 2000). There is a need for more studies in areas such as East Asia that can have even higher N deposition rates. 3.1.3. Groundwater inputs Groundwater can be an important route of N delivery to salt marshes that are adjacent to N-loaded watersheds with low evapotranspiration, conductive soils, and short flow paths. Discharge occurs at either the marsh-upland border, directly into tidal creek beds, or can flow under the marsh and discharge subtidally offshore (Howes et al., 1996; Portnoy et al., 1997). Groundwater fluxes range from <1.0 to 100 L/m2/day (Tobias et al., 2001a). Groundwater N fluxes to salt marshes (0.2–100 g N/m2/year; Table 3) reflect the net result of discharge, N concentration in groundwater, marsh area, and some influence of porewater drainage. Great Sippewissett Marsh overlies an N-rich, highly conductive aquifer; 50% of the new N to the marsh is supplied by GW (Valiela and Teal, 1979). However, Anderson et al. (1997) presented a marsh N budget where inputs of GW N were trivial despite close proximity to an N-rich aquifer. In other cases, GW N delivery is low due to N-poor GW (despite high flow rates) or because GW bypasses (i.e., flows under) marsh substrates (Portnoy et al., 1997; Tobias et al., 2001a–c). Because discharge typically decreases with distance from the upland-marsh border (Howes et al., 1996), the GW N flux can be less important in large marshes with short stretches of marsh-upland border than in fringing marshes. Flux estimates provided by watershed-scale numerical models
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Table 3
Selected groundwater nitrogen (GW N) fluxes estimated for a range of salt marshes
Marsh
GW N flux (gN/m2/year)
Dominant N species
Method
References
0.2
NO 3
Tobias et al., 2001a
<1.7
NO 3
Salt balance, hydraulic head Spring discharge
0.4
DON
Hydraulic head
Nauset Marsh, MA USA Nauset Marsh
0.4
NO 3
Simulation model
2–28
NO 3
Multiple methods
Great Sippewissett, MA Pamet River Estuary, MA North Inlet, SC
26
NO 3
Flood-ebb salt balance
Dame et al., 1991; Wolaver et al., 1988 Colman and Masterson, 2008 Giblin and Gaines, 1990; Nowicki et al., 1999 Valiela et al., 1978
23.8–99.4
DIN
Radium budget
Charette, 2007
12
NHþ 4
Radium budget
Krest et al., 2000
Ringfield, VA, USA Philips Creek, VA, USA Bly Creek, SC, USA
Anderson et al., 1997
Method denotes the technique used to calculate the water flux component of the GW N flux.
(Colman and Masterson, 2008) may provide a better-constrained estimate of GW-N flux when compared to small-scale direct measurements extrapolated to the whole ecosystem. The direct-measure flux estimates, particularly those based on geochemical tracers (e.g., radium budgets; Krest et al., 2000), include some component of recycled porewater N that can lead to an overestimate of GW-N delivery. More recent studies have used radium isotopes to parse out the contribution of “fresh” groundwater from that of recycled porewaters (Charette, 2007). Despite some of the uncertainties associated with radiumbased estimates of GW-driven exchanges of N, this approach operates at expanded scales not afforded by other techniques (Krest et al., 2000; Charette et al., 2003; Charette, 2007).
3.1.4. Tidal inputs/outputs Tidal exchange usually dominates the mass fluxes of N between salt marshes and adjacent ecosystems (Morris, 1991; Rozema et al., 2000). Thirty years of tidal flux studies show that marshes transform N, sometimes acting as a net importer or exporter of dissolved or PN. The magnitude, and in some cases the direction, of such exchanges can reverse on short timescales within a single marsh (Dankers et al., 1984; Wolaver et al., 1988; Whiting et al., 1989; Anderson et al., 1997) and it remains difficult to predict whether a particular marsh will import or export various N species at any given time.
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Table 4 A representative subset of reported tidal exchange fluxes (g N/m2/year) N fraction
Net import Range
NO 3 NHþ 4 DON PN
0.6–2.7 0.4–4.8 0.9–24.1 0.9–31
Net export
Mean
std
n
Range
Mean
std
n
1.6 2.9 7.4 8.7
0.8 1.9 11.2 15.2
8 6 4 4
0.2–6.3 1.6–8.7 0.3–9.2 4.5–42
2.1 4.2 3.3 10.7
2.1 2.8 17.3 16.1
11 7 4 5
The range, mean flux, number of studies, and standard deviation (std) of the flux magnitudes are reported. Exchange data from Jordan et al. (1983), Anderson et al. (1997), Chambers et al. (1994), and compilations of previous work presented in Dame (1994), Childers (1994), Childers et al. (2000), and Valiela et al. (2000).
Over twenty salt marsh N studies provide an accounting of N exchanges as DIN, DON, and particulate organic N (PON)/PN (Table 4). In this data set, 40% of the marshes were net annual importers of NHþ 4 , 35% imported NO3 , 27% imported DON, and 26% imported PON/PN. The exchanges (either import or export) of NHþ 4 tended to be larger than those of NO3 . The magnitude of import/export of DON and PON were higher than those of the DIN fractions but were highly variable. The developmental phase or “age” of the marsh seems to play a role in exchanges. Relative marsh age has been inferred by its location along the marsh estuarine continuum where young marshes lie closest to the upland border (Dame and Gardener, 1993; Dame, 1994) or have been defined by the ratio of vegetated to open water areas and the extent of tidal creek connections, where young marshes posses large areas of open water and few inlet/outlets (Valiela et al., 2000). When tidal exchanges are broken down as a function of marsh age (Dame, 1994; Valiela et al., 2000), younger marshes tend to import total N (although DON is often exported), whereas older marshes often export N as both DIN and DON. Some of the NO 3 export from these mature marshes may be explained by NO3 -rich groundwater discharge (Howes et al., 1996) or rapid oxidation (i.e. nitrification) of porewater that drains into exposed creekbeds during ebb tide. Mid-aged marshes generally import PN and export dissolved N (Dame, 1994). Tidal amplitude is superimposed on marsh age as an added control on tidal fluxes – younger marshes are more sensitive to changes in tidal amplitude – a response that reflects changes in hydraulic conductivity, porewater chemistry, and rates of porewater drainage as marshes age (Whiting and Childers, 1989; Childers, 1994; Howes and Goehringer, 1994). Using data exclusive of flume studies, Childers et al. (2000) showed that tidal amplitude alone accounted for 40% of the variance in NO 3 exchanges. At tidal amplitudes above 1.2 m, marshes tended to switch from NO 3 import to export. This switch is consistent with higher tidal ranges delivering more direct drainage of high NHþ 4 porewater that can be rapidly oxidized and exported as NO . 3 The largest PON (and POC) import is observed in younger marshes (Childers, 1994; Dame, 1994; Osgood, 2000; Valiela et al., 2000). Tidal range positively
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influences PON import in younger systems, but does not affect older marshes, which tend to be net exporters (Childers, 1994; Dame, 1994). Across a range of marsh types, net annual import of PON is observed about as often as net export of PON (Table 4). Stoichiometric PON flux estimates (POC flux N : CPOM) agree reasonably well with other direct estimates of PON import and range from 0.9 to 42 g N/m2/year. 3.1.5. Gaseous losses Denitrification is the primary route for losses of gaseous N from salt marshes. In addition to N2, several intermediate N gases are produced during denitrification (e.g., N2O). Denitrification in saturated soils/sediments usually yields <10% N2O as an end product (Smith et al., 1983; Seitzinger and Kroeze, 1998). Under exceptionally high NO 3 loads, Tobias et al. (2001c) reported a N2O : N2 ratio closer to 0.40 and suggested denitrification as a source of N2O in heavily N-loaded marshes. Although NHþ 4 is abundant in porewaters, soil pHs are typically low enough (NH3 pKa = 9.7) to prevent significant volatilization of NHþ 4 to NH3(g) (Morris, 1991). Organic carbon is the principal electron donor for denitrification; the importance of alternate electron donors (e.g., reduced sulfur or iron) to marsh denitrification remains open. Denitrification coupled to organic matter oxidation ½ðCH2 OÞy ðNH3 Þz þ 0:8yNO 3 !0:4yN2 þ ð0:2y zÞCO2 þ ð0:8y þ zÞHCO 3 þ ð0:6y zÞH2 O
ð2Þ
Denitrification rates range from 0 to 60 g N/m2/year (Table 5) and comparisons are hampered by methodological differences (see discussion in Seitzinger et al., 2006). Nevertheless, the differences in rates (Table 5) cannot be explained solely by different methods. Total denitrification depends on O2 concentrations in the subsurface, NO 3 supply, quantity and quality of organic carbon, and the presence of inhibitors (e.g., sulfide). The source of NO 3 for denitrification comes from allochthonous sources (direct denitrification) or is supplied from mineralization– nitrification (coupled denitrification; Figure 1). Seitzinger et al. (2006) suggested a breakpoint around 20 mM NO 3 in a variety of freshwater and saline environments where coupled denitrification dominates at lower NO 3 concentrations and direct denitrification dominates when NO exceeds 20 mM. This assertion has not been 3 well-tested in marshes. While O2 in high abundance inhibits direct denitrification, its delivery into the NHþ 4 -rich subsurface promotes the high rates of coupled nitrification–denitrification observed in close proximity to root channels and infaunal burrows (Howes et al., 1981; Dollhopf et al., 2005). Hammersley and Howes (2005) attributes several-fold higher coupled denitrification rates measured in situ versus laboratory rates to the enhancement of denitrification by plants. Seasonal variations in direct denitrification reflect changes in the external NO 3 loading (Koch et al., 1992). Coupled denitrification is enhanced by external inputs of
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Table 5 Select salt marsh denitrification rates Reference
Location
Ambient NO 3
0.6 1.7 0–44
Mid Atlantic, VA Gulf Coast, LA New England, MA
Philips Creek, VA Four League Bay, LA Great Sippewisset, MA
Low Moderate to high High
0.3–0.8 0.3–0.7 0–1.8 0–19.6 2.3–5.5 102 30 2–60 0.4–14.3 7
UK UK Southeastern NC Mediterranean Canada Mid Atlantic, VA New England, MA New England, MA East Coast, USA, Europe Gulf Coast
Colne Pt., UK Torridge Marsh, UK Newport River, NC Venice Lagoon, Italy St. Lawrence River, Canada Ringfield, VA Mashapaquit, MA Great Sippewisset, MA several Barataria Bay, LA
Moderate to high Moderate to high Low High Low to moderate Very high High High Variable Moderate
Rates encompass various laboratory and in situ techniques including N2 accumulation,
15
N2,
15
N2O dilution, isotope pairing, acetylene block, and mass balance methods.
Craig Tobias and Scott C. Neubauer
Anderson et al., 1997 Smith et al., 1985 Kaplan et al., 1979, Valiela and Teal, 1979 Aziz and Nedwell, 1986 Koch et al., 1992 Thompson et al., 1995 Eriksson et al., 2003 Poulin et al., 2007 Tobias et al., 2001b Hammersley and Howes, 2003 Hammersley and Howes, 2005 Compilation by Morris, 1991 DeLaune et al., 1989
Geographic Region
g N/m2/year
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N2, atm Tidal exchange ON, NO3–, NH4+
Watershed input ON, NO3–, NH4+
Mineralization
ON
NH4+
NO3–
ANAMMOX
Flux
Flux
NH4+
DNF DNF
NH4+
NO2–
NO3–
DNRA
NH4+
Mineralization
ON
Nitrification Immobilization
Immobilization
Figure 1 Summary of the major N cycling pathways in salt marshes. DNF = denitrification; DNRA = dissimilatory nitrate reduction to ammonium; ANAMMOX = anaerobic ammonium oxidation.The sizes of the arrows denote the relative magnitudes of the processes.
reduced N (Hammersley and Howes, 2005) and responds to increased NHþ 4 supply from mineralization. Positive relationships have been observed between mineralization and denitrification in freshwater systems (Seitzinger, 1994; Mulholland et al., 2008). This pattern is mirrored in some salt marshes where higher denitrification is coincident with seasonally high mineralization rates (Anderson et al., 1997). For marshes dominated by either type of denitrification, the highest rates are encountered in surface soils that are exposed at low tide, in closest contact with NO 3 in overlying water at high tide, nearest the zone of nitrification, and closest to nearsurface labile organic matter supplies (Koch et al., 1992; Tobias et al., 2001b). Because denitrification exports N, it plays an important role in overall N residence time. Competition for N between plants and denitrification is a regulator of long-term N retention, with denitrification reported at roughly 20% of plant uptake (White and Howes, 1994b). There are examples of marshes where plant growth far outcompetes denitrification for N (Anderson et al., 1997) and those where denitrification and plant N demand are roughly equal (Morris, 1991). Studies in several mature marshes put losses of N through denitrification approximately equivalent to inputs via N fixation, suggesting that the balance between these two pathways are important for overall marsh N balance (Valiela and Teal, 1979; Morris, 1991; White and Howes, 1994b; Anderson et al., 1997; Rozema et al., 2000). Denitrification ranges between 15% and 100% of N burial, with most marshes showing similar magnitudes for each pathway. There are only a few examples of systems where denitrification vastly exceeds burial and are likely explained by high NO 3 loading that fuels high rates of direct denitrification (Kaplan et al., 1979). As with all other input and removal pathways, the magnitude of denitrification is small relative to internal N cycling.
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Anaerobic ammonium oxidation (ANAMOX) is an alternate pathway by which N can be exported from ecosystems. ANAMMOX uses ammonium to reduce nitrite to produce N2. It is chemoautotrophic and, unlike denitrification, does not require organic carbon. ANAMMOX can rival rates of denitrification in a variety of marine sediments but is less important in high organic carbon systems (like salt marshes), where it accounts for less than 10% of the N2 production (Dalsgaard et al., 2005). It is not currently considered an important route of N loss from salt marshes.
3.2. Internal cycling 3.2.1. Photoautotrophy Except for some very young marshes, autotroph N demand exceeds inputs and primary production is fueled by recycled N. Macrophyte N uptake is one of the larger N flux terms ranging from 1 to 33 g N/m2/year (Gallagher et al., 1980; Hopkinson and Schubauer, 1984; Morris, 1991; Blum, 1993; Dai and Weigert, 1996; Anderson et al., 1997; Rozema et al., 2000; Hopkinson and Giblin, 2008). Microalgal uptake in marshes is 10–15% of emergent macrophyte demand (Anderson et al., 1997; Rozema et al., 2000; Tyler et al., 2003). Aboveground plant N content ranges from <0.2% to 3% (Buresh et al., 1980; Anderson et al., 1997). Nitrogen frequently limits plant production; plants respond to nitrogen additions with increased N content and enhanced biomass production (Deegan et al., 2007; Drake et al., 2008; and others). 3.2.2. Mineralization and immobilization The sum of new N inputs to salt marshes is on the order of 0.5% to <5% of the total N necessary to support macrophyte production (Hopkinson and Schubauer, 1984; DeLaune et al., 1989; Dame et al., 1991; Anderson et al., 1997). The difference is made up by the mineralization of organic N into the NHþ 4 pool that has a high rate of turnover (Figure 1). Both aerobic and anaerobic (e.g., sulfate reduction) mineralization pathways yield NHþ 4. Aerobic mineralization (y/z = C : N ratio of the organic matter) ðCH2 OÞy ðNH3 Þz H3 PO4 þ 106O2 ! yCO2 þ zNH3 þ H3 PO4 þ yH2 O
ð3Þ
Sulfate reduction ðCH2 OÞy ðNH3 Þz H3 PO4 þ 53SO2 4 ! yCO2 þ zNH3 þ H3 PO4 þ 0:5yS2 þ yH2 O
ð4Þ
Consequently, porewater NHþ 4 concentrations in well-developed marshes are high. Mineralization rates vary widely between marshes (3.0–122 g N/m2/year; Morris, 1991; Anderson et al., 1997; Rozema et al., 2000; Thomas and Christian, 2001; Tobias et al., 2001a). Mineralization rates decrease exponentially with depth due to decreasing organic matter quantity and lability (Bowden, 1984; Howes et al., 1985; Tobias et al., 2001a). For a variety of marshes, mineralization meets between 50% and 200% of autotrophic N requirements. Even in systems that appear to have mineralization
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463
rates that exceed uptake by N sinks (e.g., plants or denitrification), no large exports of N are observed (Anderson et al., 1997). Several lines of evidence indicate that microbial immobilization plays an important role in retaining a significant fraction (15–50%) of mineralized N (Anderson et al., 1997; Hopkinson and Giblin, 2008). Increases in the N content of macrophyte detritus during diagenesis are consistent with the immobilization of porewater NHþ 4 (Benner et al., 1991). Significant retention of 15N tracers within marsh soils over 100 days (93%; DeLaune et al., 1983) to 7 years (40%; White and Howes, 1994a,b) indicates that the NHþ 4 pool acts as a common substrate for exchanges among plant assimilation, microbial recycling, and removal via denitrification and burial. 3.2.3. Nitrification Nitrification is the oxidation of ammonium to nitrate. It consumes molecular O2 and links mineralization to N export via denitrification. Nitrification is the ratelimiting step for subsequent denitrification of mineralized N out of marshes. þ NHþ 4 þ 1:5O2 ! NO2 þ 2H þ H2 O NO 2 þ 0:5O2 ! NO3
ð5Þ
Nitrification occurs in the presence of adequate supplies of O2 and NHþ 4; because porewaters are typically rich in NHþ , nitrification is generally limited by 4 O2 availability. It is restricted either to surface soils or oxic microzones, declines with depth, and is influenced by tidal wetting and drying in close proximity to root channels or burrows (Howes et al., 1981; Tobias et al., 2001a; Eriksson et al., 2003; Dollhopf et al., 2005; Costa et al., 2007). Nitrification activity is inhibited by elevated sulfide levels (Joye and Hollibaugh, 1995), high salinities (Seitzinger et al., 1991; Rysgaard et al., 1999), and low pH (<4.50) characteristic of drained marshes undergoing acidification due to high rates of sulfur oxidation (Portnoy and Giblin, 1997). Under anaerobic conditions, NO 3 is quickly reduced so concentrations are low in the absence of large external NO 3 inputs. Annual rates of nitrification range from 0.26 to 52 g N/m2/year (Abd Aziz and Nedwell, 1986; Anderson et al., 1997; Tobias et al., 2001b; Eriksson et al., 2003; Hammersley and Howes, 2003; Dollhopf et al., 2005; Costa et al., 2007). Eighty percent of the reported rates are <10 g N/m2/year, with higher rates observed during warmer months (Thompson et al., 1995; Anderson et al., 1997). Some of the highest estimates come from marshes that are heavily N loaded (Eriksson et al., 2003; Hammersley and Howes, 2005) suggesting that competition for N can occur between nitrifiers and autotrophs. In marsh N budgets, nitrification is 4- to 20-fold lower than mineralization (Abd Aziz and Nedwell, 1986; Anderson et al., 1997; Neubauer et al. 2005a) and is roughly equivalent to coupled denitrification. In situ estimates of nitrification following a whole ecosystem 15NHþ 4 release showed that rapid nitrification on the marsh surface accounted for 30% of the NHþ 4 transformations (Gribsholt et al., 2005, 2006). This in situ technique yielded nitrification rates four- to eightfold higher than those measured in the laboratory and illustrated the importance natural gradients of O2 and NHþ 4 in maintaining this reaction. This type of approach shows promise in addressing nitrification (and other processes) at the ecosystem scale.
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3.2.4. Dissimilatory nitrate reduction to ammonium Dissimilatory nitrate reduction to ammonium (DNRA) retains N (Figure 1). The reaction is largely mediated by facultative anaerobic fermentative bacteria (Tiedje, 1988; Nijburg and Laanbroek, 1997; Megonigal et al., 2004) although chemolithoautotrophic DNRA has been proposed as an additional DNRA pathway (Brunet and Garcia-Gil, 1996). DNRA – complete stoichiometry – from Canavan et al. (2007) ½ðCH2 OÞyðNH3 Þz þ 0:5yNO 3 þ zCO2 þ ð0:5y þ zÞH2 O ! ð0:5y þ zÞNHþ 4 þ ðy þ zÞHCO3
ð6Þ
Temperature, salinity, sulfide, redox, labile organic carbon, and NO 3 concentration have all been suggested as potential controls on DNRA (King and Nedwell, 1985; An and Gardner, 2002; Laverman et al., 2007). DNRA appears to be favored at high- and low-temperature extremes (Kelly-Gerreyn et al., 2001) and seems to be enhanced by a high organic carbon : NO 3 ratio or when ample sulfide provides an alternate electron source (Gardner et al., 2006). DNRA can be further coupled to sulfur cycling through sulfide inhibition of denitrification which allows DNRA to dominate (Brunet and Garcia-Gil, 1996; Christensen et al., 2000; An and Gardner, 2002; Senga et al., 2006) or by inducing some sulfate-reducing bacteria þ to switch to NO 3 as the terminal electron acceptor and produce NH4 (Dalsgaard and Bak, 1994; An and Gardner, 2002). Different sulfide and carbon loads may also influence microbial community controls on DNRA and contribute to variations in observed rates (Burgin and Hamilton, 2007; Scott et al., 2008). DNRA effectively competes with denitrification for NO 3 across a range of “wet” ecosystems where it often accounts for over 50% of the NO 3 reduction (Kelly-Gerreyn et al., 2001; Megonigal et al., 2004). DNRA is currently believed to be most prevalent in marine sediments where rates range from 0 to 40 g N/m2/ year (Gilbert et al., 1997; Bonin et al., 1998; Christensen et al., 2000; Kelly-Gerreyn et al., 2001; An and Gardner, 2002; Gardner et al., 2006; Thornton et al., 2007). It has been measured only a few times in salt marshes (King and Nedwell, 1985; Tobias et al., 2001a,c) and fresh water wetlands (Bowden, 1986; Neubauer et al., 2005a; Scott et al., 2008). DNRA in mesohaline marsh soils, measured in the laboratory and in situ using a 15NO 3 tracer release, yielded “annualized” rates ranging from 1.2 to 92 g N/m2/year. These rates ranged from 0.3 to 2 times that of denitrification (Tobias et al., 2001a,c). DNRA appears to be an important fate for NO 3 in salty, high-sulfide, organic-rich sediments, yet this pathway remains understudied in salt marshes (Poulin et al., 2007). The positive relationship between DNRA and increasing salinity/sulfide suggests increasing N retention in tidal fresh and mesohaline marshes as seawater encroaches up estuaries due to sea level rise (SLR).
3.3. Burial Long-term nitrogen burial rates in salt marshes have been estimated directly using sediment traps, horizon markers, radioisotope profiles (210Pb, 137Cs), and extrapolated from SLR estimates (Hutchinson et al., 1995; Merrill and Cornwell, 2000;
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Craft, 2007; Goodman et al., 2007). N (and P) burial combines contributions from allochthonous suspended sediments and autochthonous organic matter. N is recycled multiple times within marsh soils, porewater, and biota before it ultimately becomes sequestered primarily in belowground organic matter (White and Howes, 1994b). Nitrogen burial rates along the Atlantic and Gulf coasts (USA) range from 1 to 23 g N/m2/year (Craft, 2007 and references therein). For Wadden Sea salt marshes, net sedimentation of N is on the order of 10–50 g N/m2/year (Rozema et al., 2000). For systems that lack direct measurements, a SLR proxy can be used to estimate N burial. With some exceptions (Chmura and Hung, 2004), there is good agreement between marsh accretion and SLR (Donnelly et al., 2004). By combining typical bulk density and N content estimates for marsh soils with local measures of SLR, one can broadly generalize that marshes keeping pace with SLR are burying N on the order of 2–6 g N/m2/mm SLR. In rapidly accreting marshes (>5 mm/year), N burial can be comparable to the rate at which new N is delivered. For most other systems, N burial is on the order of 50–60% of the total N inputs (Anderson et al., 1997; Neubauer et al., 2005a). Network analysis of three salt marshes (New England, Georgia, and Mid Atlantic, USA) indicated that burial is second only to tidal export as the fate of imported PN across these marsh types (Thomas and Christian, 2001). N burial is inversely related to salinity due to a lower soil N content at euryhaline sites and enhanced decomposition of belowground organic matter at higher salinity marshes (Craft, 2007).
4. IRON AND SULFUR The biogeochemical cycles of iron and sulfur are tightly linked in salt marsh soils (Figure 2), primarily through the abiotic reduction of Fe(III) by sulfides and via the reactions that control the formation and dissolution of Fe–S minerals such as iron monosulfides (FeS) and pyrite (FeS2). Further, both elements undergo active redox cycling, play a role in organic C catabolism (Section 2.2), and may also contribute to significant chemoautotrophic production (Section 2.2.3).
4.1. Exchanges 4.1.1. Dissolved Fe and S exchanges If water infiltrating marsh soils to replace the 5–20 L/m2/day lost to seepage and evapotranspiration (Morris, 1995) has near-oceanic SO2 concentrations 4 2 2 2 (28 mM SO2 ), there is a potential SO source of 4.5–18 g SO 4 4 4 -S/m /day. 2 These tidal SO4 inputs often exceed rates of sulfate reduction (Morris, 1995; Kostka et al., 2002b). SO2 4 delivery rates vary with soil depth, distance from tidal creeks, and location within the estuary. Sulfate concentrations in marsh porewaters are typically depleted (relative to a conservative tracer), indicating net SO2 4 utilization (Gardner et al., 1988; Hines et al., 1989; Hseih and Yang, 1997). We do not know of any estimates of dissolved Fe delivery to salt marshes although this flux is probably insignificant in the total marsh Fe budget.
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Figure 2 The Fe and S cycles are linked in salt marsh soils through a series of biologically mediated and abiotic reactions. This figure focuses on inorganic forms of Fe and S. With the exception of thiol (R-SH) formation during biological pyrite oxidation, the complexities of the organic Fe or S cycles are not considered here. For simplicity, the oxidation of FeS is not shown although the products of this reaction are similar to those of pyrite oxidation (i.e., SO2 4 , Fe(II), and/or Fe(III)). FeOB and SOB = Fe(II) and sulfide-oxidizing bacteria, respectively; FeRB and SRB = Fe(III) and SO2 4 -reducing bacteria, respectively. Oxidants are compounds including O2, NO 3 , MnO 2, and Fe(III) oxides that can oxidize reduced compounds. CH2O generically refers to organic C, which can be in the form of organic molecules that are catabolized by heterotrophic microbes such as FeRB and SRB or as ligands that play a key role in solubilizing and/or reducing Fe minerals (e.g., Luther et al., 1992; Carey and Taillefert, 2005).
The advection of marsh porewater and diffusive fluxes each play a role in exporting dissolved S and Fe. Estimates of the export of H2S (20 g S/m2/year) and thiosulfate (1,504 g S/m2/year) from Great Sippewissett Marsh are considerable, representing the energetic equivalent of 23% of total NPP (Howarth and Teal, 1980). Using a H2S concentration of 100 mM for tall S. alterniflora soils (King et al., 1982) and applying a porewater replacement rate of 5–20 L/m2/day (as above), we estimate an export of 6.4–22 g H2S-S/m2/year, a rate that is comparable to the value calculated by Howarth and Teal (1980). Luther et al. (1982b) observed that – the SO2 4 : Cl ratio over the tidal cycle decreased during ebb tide in two marsh creeks but increased dramatically in a third creek, indicating that total S import/ export depends on spatially distinct balances between sulfate reduction and oxidation of metal sulfides. Guo et al. (2000) reported a Fe2þ flux of 0.9–1.9 g Fe/m2/year
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from saline and brackish marsh microcosms. The loss of Fe2þ due to porewater drainage will be greater at creek banks than in mid- and high-marsh locations due to a higher Fe2þ supply [more Fe(III) reduction and a lower degree of pyritization (DOP)] and more rapid rates of porewater turnover in creek bank marshes. Given the kinetics of Fe(II) oxidation, it is likely that a significant fraction of Fe2þ in tidal waters will be rapidly reoxidized and may settle back onto marsh surfaces. However, reactions with organic ligands may keep the resulting Fe(III) in a soluble form (Luther et al., 1996) and allow Fe(aq) to be exported. 4.1.2. Atmospheric exchanges 2 2 The deposition of atmospheric SO2 4 (<1 g SO4 -S/m /year, NADP, 2008) and 2 Fe (0.1 g Fe/m /year; Duce et al., 1991) in the coastal zone is much lower than other inputs of these elements. However, atmospheric fluxes may be locally elevated in close proximity to industrial emissions (Lecoanet et al., 2001). Saline and brackish marshes emit H2S, dimethyl sulfide [DMS, (CH3)2S], carbonyl sulfide (COS), methanethiol (CH3SH), carbon disulfide (CS2), and dimethyl disulfide [(CH3S)2] (Morrison and Hines, 1990; DeLaune et al., 2002). Most salt marshes are net sources of S gases to the atmosphere (Cooper et al., 1987; Dacey et al., 1987; Morrison and Hines, 1990; Crozier et al., 1995; DeLaune et al., 2002) although there are reports of uptake of specific gases (e.g., COS; Morrison and Hines, 1990). DeLaune et al. (2002) reported an average total gaseous emission rate of 0.6 g S/m2/year for a salt marsh and 0.3 g S/m2/year for a brackish marsh. The production of sulfur-containing gases is positively correlated with SO2 4 concentrations (Crozier et al., 1995), a pattern that was observed for total gaseous S emissions along a salinity gradient (DeLaune et al., 2002). Because the production of these gases is biologically driven, emission rates are generally greater with higher plant biomass and growth (Morrison and Hines, 1990; DeLaune et al., 2002). DMS dominates emissions (>50%) in S. alterniflora marshes with low emissions as H2S (<10%; Cooper et al., 1987; DeLaune et al., 2002b). A S. patens brackish marsh emits primarily H2S and COS (DeLaune et al., 2002). High DMS emission rates from S. alterniflora versus S. patens brackish marshes reflect some differences in plant physiology; S. alterniflora contains high concentrations of dimethylsulfoniopropionate (DMSP), the precursor to DMS, whereas S. patens does not (Dacey et al., 1987). The differential emissions along salinity gradients also show the effect of sulfide sequestration in metal S complexes in euryhaline marshes. Despite emission of several S gases, it is not an important loss term for individual marshes relative to burial and tidal export. While S gas emissions are important for atmospheric chemistry and radiative balance (Shaw, 1983; Kelly and Smith, 1990), salt marshes contribute little as a global source/sink. 4.1.3. Sedimentary Fe and S deposition Recently deposited sediments and nearsurface marsh soils often have Fe contents in the range of 1–2% (mass basis; DeLaune et al., 1981; Gardner et al., 1988; Merrill and Cornwell, 2000) and S contents of 0.5–2% (although values up to 5% can occur; Cutter and Velinsky, 1988; Gardner et al., 1988; Giblin, 1988). Using these
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Fe and S concentrations and mass deposition rates for Paulina Marsh (Section 2.1.2), annual deposition of these elements is 50–100 g Fe/m2/year and 25–100 g S/m2/ year. These deposition rates roughly overlap with long-term accumulation (burial) rates of these elements (Section 4.3).
4.2. Internal cycling 4.2.1. Iron and sulfur reduction Iron(III) reduction can be biotic or abiotic (Figure 2). Microbial and geochemical analyses have suggested a significant, if not dominant, role for microbial Fe(III) reduction although there is spatial variability in the importance of each pathway (Section 2.2.2). Large populations of Fe(III)-reducing bacteria (FeRB) can be found in vegetated soils and around roots (Lowe et al., 2000; Kostka et al., 2002a; Koretsky et al., 2003; Weiss et al., 2003), proximal to SO2 4 -reducing bacteria and aerobic microbes (Koretsky et al., 2005). Microbial Fe(III) reduction accounts for 80–100% and 39–53% of total reduction in mineral- and organic-rich marsh soils, respectively (Kostka et al., 2002a; Gribsholt et al., 2003; Neubauer et al., 2005b; Hyun et al., 2007). Koretsky et al. (2003) proposed that seasonal changes in rates of H2S production cause oscillation between chemically dominated Fe(III) reduction in summer and biologically dominated Fe(III) reduction in other seasons. The reduction of SO2 4 is primarily a biological process coupled to the oxidation of simple organic compounds including volatile fatty acids and acetate (Hines et al., 1994, 1999; Boschker et al., 2001). The H2S produced tends to rapidly convert to pyrite although some fraction (<5–23%) appears as organo-sulfur compounds (Howarth, 1979; Neubauer et al., 2005b). Seasonal patterns of SO2 4 reduction tend to follow temperature patterns and are also influenced by marsh primary production (Howarth and Teal, 1979; Hines et al., 1999; Koretsky et al., 2003). Sulfate reduction decreases when plants allocate more C to flowering/ reproductive structures and less C leaks into the soil (Lytle and Hull, 1980; Hines et al., 1999). The relative abundance of SO2 4 -reducing bacteria (SRB) ribosomal ribonucleic acid (rRNA) associated with plant roots increased in parallel with changes in SO2 reduction activity (Hines et al., 1999). Similarly, the SRBs 4 Desulfobacter and Desulfovibrio were more prevalent in summer than winter (Koretsky et al., 2005). However, Hines et al. (1999) reported that the rRNA that could be ascribed to SRB did not change significantly over the year (relative to bacterial rRNA) even though rates of SO2 4 reduction did show seasonal variability. 4.2.2. Formation and oxidation of Fe–S minerals Solid phase sulfides are operationally divided into acid volatile sulfides (AVS, consisting of H2S and FeS) and chromium (II)-reducible sulfides (CRS: FeS2 and S0). Early work measuring SO2 reduction rates demonstrated that a large 4 fraction of the added 35SO2 tracer appeared as solid phase sulfides, indicating that 4 Fe–S minerals form rapidly in salt marsh soils (Howarth, 1979). In terms of concentration, pyrite is a 10-fold more important end product than iron monosulfides
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although the mechanisms of pyrite formation involve an FeS intermediary (Figure 2; Giblin and Howarth, 1984; Cutter and Velinsky, 1988; Gardner et al., 1988). The reaction of Fe(II) and H2S leads to the formation of a soluble FeS phase when suitable organic ligands are present (Taillefert et al., 2000; Carey and Taillefert, 2005) or to direct precipitation of FeS minerals. Reaction with additional reduced S (as H2S or polysulfides, S2 x ) then leads to FeS2 precipitation (Canfield et al., 1998; Rickard and Luther, 2007). Inorganic sulfide minerals are not always the dominant form of S in salt marsh soils (Figure 3a). Across a range of salt marshes, CRS accounts for 20–100% of total soil S [average CRS : TS (total sulfur) ratio = 0.57+0.25 SD]. The nonpyritic sulfur is likely dominated by organic S forms (Giblin, 1988). Organic S accounted for 45+19% of the total S pool in the top 30–50 cm of Great Marsh, Delaware (Cutter and Velinsky, 1988). Similarly, the soil C content is inversely correlated with the fraction of CRS-S (Figure 3b) in Great Marsh, Delaware (Cutter and Velinsky, 1988), and North Inlet, South Carolina (Gardner et al., 1988). In this same data set, there was a positive relationship between the DOP and the CRS : TS ratio for Great Marsh but not for North Inlet (data not shown). The DOP (Berner, 1970), an index of how much of the available Fe has reacted with sulfides to form pyrite, is calculated as DOP =
Fepyrite Fepyrite þ Fereactive
ð7Þ
where Fepyrite is the amount of Fe in pyrite and Fereactive is the amount of reactive or amorphous Fe in the sample. In salt marshes, DOP values for interior/high marsh sites are generally higher than for creekbank/levee sites (King et al., 1982; Gardner et al., 1988; Giblin, 1988; Otero and Macias, 2003; Roychoudhury et al., 2003a; 5.0 4.5 4.0 3.5 3.0 2.5 2.0 1.5 1.0 0.5 0.0 0.0
1:1 Cutter and Velinsky, 1988 Gardner et al., 1988
1:2
Giblin, 1988
1:5
1.0
2.0 3.0 Total S (%)
4.0
6.0
(b) 1.0 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0
CRS : TS = –0.049 × % C + 0.85 R 2 = 0.390
CRS : TS ratio
Pyrite-S or CRS-S (%)
(a)
0
2
4
8 10 6 Soil C (%)
12
14
Figure 3 Soil pyrite relationships in salt marsh soils. Most data are from Spartina alterniflora marshes although the Giblin (1988) study also included data from sites vegetated with Spartina patens, Scirpus, and Typha. Data points are individual samples and reflect temporal and spatial (lateral and depth related) variability in solid phase soil chemistry. (a) Chromium(II)reducible sulfur (CRS) versus total sulfur (TS) in marsh soils. Solid lines indicate contours where CRS = TS, CRS = 0.5TS, and CRS = 0.2TS. (b) Relationship between the CRS : TS ratio and the soil C content.The regression line is fit through all data.
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Roychoudhury, 2007). Sulfide production in interior marsh sites is 1–2 orders of magnitude greater than the rate of Fe delivery (Gardner et al., 1988), and DOP values can approach 1.0 at these sites (DOP 0.6–1.0). The lower DOP in creek bank marshes (0.2–0.7) indicates an enhanced supply of reactive Fe (Section 2.2.2), decreased sulfide availability (King et al., 1982), and/or processes that limit the formation/accumulation of pyrite. Gross rates of pyrite formation are considerably greater than net rates of pyrite accumulation (Gardner, 1990), indicating that pyrite is actively oxidized or exported from marsh soils. Net pyrite oxidation exists during the spring/summer growing season while net pyrite formation occurs in winter (Giblin and Howarth, 1984; Cutter and Velinsky, 1988). Variations in the timing and frequency of flooding of marsh surfaces can override some of this seasonal generality (see discussion in Giblin, 1988). Hseih and Yang (1997) speculated that roots play a major role in rates of pyrite accumulation by oxidizing the soil and providing labile organic C to fuel Fe(III) reduction and SO2 4 reduction. Below the rooting zone, pyrite concentrations do not show large seasonal or spatial variability, again emphasizing the role of plants in both FeS2 formation and oxidation (Cutter and Velinsky, 1988; Hseih and Yang, 1997). Diagenetic model results suggest that rhizosphere oxidation (with either O2 or organic compounds as oxidants) was responsible for the majority of FeS2 oxidation (Gardner, 1990). Bioturbation near creek banks can remove 18% of subsurface pyrite by bringing it to the surface where it could be oxidized and/ or washed away (Gardner et al., 1988). Advection of O2-rich tidal water through soils contributes to subsurface FeS2 oxidation, and the seepage of water out of creek banks provides advective removal of H2S. This tidal flushing, rather than effects of macro-organisms and plants, was most significant in affecting pyritization in creek bank soils (Roychoudhury et al. 2003a). Away from creek banks, where infaunal densities and porewater flushing are low, rhizosphere oxidation is likely to be the most significant mechanism for pyrite turnover. Thermodynamically, FeS2 can be oxidized anaerobically with NO 3, MnO2, or Fe(III) as oxidants. With some exceptions (Schippers and Jørgensen, 2002; Carey and Taillefert, 2005), the significance of these reactions has not been fully assessed. 4.2.3. Iron oxidation Oxidized Fe, in the form of root plaques or other accumulations (Luther et al. 1982a, Mendelssohn et al., 1995; Weiss et al., 2003), can serve as an electron acceptor for anaerobic metabolism (Lowe et al., 2000; Kostka et al., 2002a; Neubauer et al., 2005b) and as a sink for trace metals and PO3 4 (Scudlark and Church, 1989; Chambers and Odum, 1990; Weis and Weis, 2004; Neubauer et al., 2008). In tidal wetlands, the solid phase Fe in the upper portions of the soil profile is generally rich in Fe oxides whereas iron sulfide minerals dominate at lower depths (Griffin et al., 1989). In shallow soils, O2 is the likely oxidant that drives both chemical and biologically mediated Fe(II) oxidation. Microbial Fe(II) oxidation linked to anaerobic photosynthesis (Widdel et al., 1993), NO 3 reduction (Straub
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et al., 1996), and perchlorate reduction (Chaudhuri et al., 2001) may also play a role. Plant roots are hotspots for oxidation, forming Fe oxide-rich plaques tens of micrometers up to 0.4 cm in thickness (Taylor et al., 1984; Vale et al., 1990). Sundby et al. (2003) proposed that rapid rates of biological and/or chemical Fe(II) oxidation (and therefore plaque formation) occurred in spring as new highly active roots invaded anoxic sediments with high Fe2þ concentrations. Over multiple years, the oxidizing activity of plant roots can increase extractable Fe(III) pools relative to unvegetated sites (Kostka and Luther, 1995; Roden and Wetzel, 1996; Weiss et al. 2004). Tidal delivery of O2 also influences temporal patterns of Fe(II) oxidation (Neubauer et al., 2005b). Because Fe(II) availability is low in mid- and high-marsh areas (i.e., DOP is high, see above), overall rates of Fe(II) oxidation are likely to be lowest in these parts of marshes. However, lower pH (Gardner et al., 1988) may lead to biological Fe(II) oxidation being relatively more important in these areas since chemical Fe(II) oxidation rates decrease as pH decreases (Singer and Stumm, 1970).
4.3. Burial Howarth (1984) reported a net rate of S accretion of 13 g S/m2/year in the Great Sippewissett and Sapelo Island marshes although this estimate does not include the burial of organic-bound S. Assuming that the S is primarily buried as pyrite (but see below), Fe burial is estimated at 11 g Fe/m2/year. Data from organic-rich brackish and salt marshes show higher rates of total S accumulation (24–46 g S/m2/year; Krairapanond et al., 1992). Further, in these marshes, the majority of the S accumulated as C-bonded S (56–65%) and ester sulfates (21–23%) with <3% due to the preservation of pyrite. Based on soil data from Cutter and Velinsky (1988) and an accretion rate of 0.47 cm/year (Church et al., 1981), the S accumulation rate has averaged 29 g S/m2/year over the last 75–100 years at Delaware’s Great Marsh, with the accumulation split roughly evenly between pyrite and organic S. In the same data set, the long-term accretion rate of 21 g Fe/m2/year was driven by pyrite (62% of total Fe accumulation) and oxidized Fe (27%). The accumulation of pyrite increased with depth whereas that of reactive oxides generally decreased with depth. Data from other salt marshes are similar in terms of the magnitude of Fe burial, with rates ranging from 11 to 60 g Fe/m2/year (DeLaune et al., 1981; Gardner et al., 1988).
5. P HOSPHORUS Phosphorus (P) in salt marshes is diversely speciated in the solid and dissolved phases. P cycling is affected by iron and sulfur reactions and thus differs considerably in salt marshes versus tidal freshwater marshes. In salt marshes, less P is adsorbed to soils, and porewaters contain more dissolved inorganic P (DIP) (Paludan and Morris, 1999). N limitation of primary production in salt marshes versus P limitation in freshwaters arises in part from this P
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speciation. The resultant four- to fivefold drop in the DIN : DIP ratio of salt marsh porewater prevents P limitation of autotrophs although marsh bacteria remain P-limited due to scarcity of organic P substrates (Sundareshwar et al., 2003). Solid phases (including sorbed PO3 4 ) dominate P inventories. Soil-bound P (0.05–0.08%) plus that in plant detritus is the largest reservoir (30–100 g P/m2), exceeding that of live plant biomass (10–20 g P/m2; 0.05–2% by weight) and 2 porewater PO3 4 (<0.5 g P/m , 0–100 mM; Buresh et al., 1980; Sundareshwar and Morris, 1999; Zhou et al., 2007). Large fractions of total phosphorus (TP) are associated with organic matter; marshes low in organic content tend to show lower overall TP and porewater PO3 (Zhou et al., 2007). The various P 4 fractions in soils are defined as follows: (1) loosely exchangeable (salt extractable) which includes both the dissolved PO3 pool and the dissolved nonreactive 4 organic P (DNRP); (2) PO3 4 bound to oxidized iron and aluminum (of which the iron fraction is redox sensitive); (3) soil organic (humic) P which is removable following sequential NaOH/HCl treatment; (4) calcium (carbonate)-bound P which is liberated following acidification; and (5) residual refractory P (Sundby et al., 1998; Paludan and Morris, 1999; Coelho et al., 2004; Zhou et al., 2007). The amount of TP through the soil profile can either be relatively constant or decline with depth, but TP inventories change little throughout the year (Stribling and Cornwell, 2001; Coelho et al., 2004; Weston et al., 2006). Increases in the refractory P content with depth are consistent with organic matter diagenesis during burial (Lillebø et al., 2007). Approximately 20 to >40% of the bound P is organic (humic þ nonreactive þ refractory fractions; Coelho et al., 2004; Alva`rez-Rogel et al., 2007). Approximately 40% of the TP in soils is readily exchangeable and available for equilibrium maintenance of porewater PO3 4 (Lillebø et al., 2007). Spatial and temporal variability in porewater PO3 coincides with changes in biological activity (e.g., mineralization 4 and plant demand) and physicochemical conditions that affect (de)sorption to/ from mineral phases. The highest PO3 4 concentrations are seen in soils where mineralization is high and low Eh reduces PO3 sorption to iron oxides 4 (Chambers et al., 1992; Chambers, 1997; Sundareshwar and Morris, 1999; Stribling and Cornwell, 2001; Lillebø et al., 2007). Exchanges between solid and aqueous phases maintain porewater PO3 concentrations, and spatial/ 4 temporal differences represent adjustments to that dynamic equilibrium rather than large changes in net P import/export.
5.1. Exchanges Aside from some instances of groundwater inputs in areas of P-rich geology (Weston et al., 2006) and rainfall scouring of marsh soils (Mwamba and Torres, 2002; Cundy et al. 2007), tides are the principal routes for delivering and removing P. These exchanges are small, however, relative to the size of the marsh P reservoir (Paludan and Morris, 1999; Coelho et al., 2004). While marshes accumulate P over their lifespan as they accrete, they can act at any given time as sinks or sources for particulate and/or DIP.
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5.1.1. Tidal inputs/outputs Settling of particulate P (organic and inorganic bound) represents a major source of P for most salt marshes. There are robust examples of both net import and export of particulate P (Dankers et al., 1984; Childers, 1994; Dame, 1994). The magnitude and distribution of particulate P delivery is controlled by settling patterns of suspended P as a function of water column load, partitioning of P between organic and mineral phases, hydroperiod, and flooding frequency (Friedrichs and Perry, 2001). The mineral-attached fraction of particulate P depends on the geochemical composition of suspended sediments. Changes in ionic strength, pH, and availability of different sorptive minerals along salinity gradients cause differences in the magnitude and form of P delivery to mesohaline and euryhaline marshes (Sundby et al., 1992; Paludan and Morris, 1999; Jordan et al., 2008). Contributions of P associated with Fe and Al oxides are greatest in low salinity environments where Fe and Al minerals are in ample supply (Fox et al., 1986; Froehlich, 1988; Sundareshwar and Morris, 1999). At euryhaline salinities with higher pHs, the Fe- and Al-bound fractions decrease as does the TP adsorbed per unit soil. The calcite-bound/sorbed fraction of P becomes a more important source of settling P to euryhaline marshes (Coelho et al., 2004; Alva`rez-Rogel et al., 2007; Figure 4). Relative to tidal flux studies of POC or PON, there are fewer estimates of organic particulate P fluxes to marshes. However, one can derive estimates of particulate organic P fluxes from previously measured PON and POC fluxes. Using existing compilations of POC/PON flux data (Dame, 1994; Rozema
Oligohaline marsh POM(P)
Fe–P
Euryhaline marsh Ca–P
OM
POM(P)
Fe–P
Ca–P
OM
O2
O2 n
n
io
io
ct
ct
du
du
re
re
S2– + PO43– +
Fe
S2– + Fe-PO4(s)
FeS(s) / FeS2(s) Burial Low Eh dissolution
PO43– +
Fe
Fe-PO4(s)
FeS(s) / FeS2(s) Burial Low Eh dissolution
Figure 4 Summary of P dynamics in marsh sediments contrasting freshwater and salt marsh ecosystems. Salt marshes receive and contain more Ca ^ P relative to Fe ^ P. Less PO43 is sorbed to iron in salt marshes due to sulfur cycling.
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et al., 2000), we assigned a conservative C : P ratio of 120 to the POC fluxes and an intermediate N : P ratio of 20 (Craft, 2007) to reported PON fluxes. Using these stoichiometries, the range of net particulate organic P (POP) fluxes ranged from –0.2 to 0.79 g P/m2/year (from POC data, negative sign indicates export; Dame, 1994) and –3.0 to 2.0 g P/m2/year (from PON data; Rozema et al., 2000). For studies that showed net POP import, average rates were 0.6 g P/m2/year (from both POC and PON stoichiometry). The export of POP averaged 0.2 and 1.0 g P/m2/year from POC and PON data, respectively. Tidal flux estimates of net P import tend to be at the low end of measured long-term P accumulation rates (Section 5.3), probably because buried P contains mineral-bound P that is not included in the POP flux calculations. Younger to middle-age marshes are more likely to show net import of particulate P (Dame, 1994). Marsh flume studies conducted across systems with tidal amplitudes 1 m showed a close balance between P import and export (i.e., insignificant-low net particulate P exchanges). There was a small, but significant net export of sediment-bound P in higher tidal range environments (Childers, 1994). An opposite tidal amplitude effect was reported for organic particulate P. This disparity between inorganic bound and organic particulate P likely reflects asymmetry in flood versus ebb energy and/or density differences between organic versus inorganic particles. Export over mean tidal conditions probably underestimates the total annual particulate P efflux. While large import of sediments and associated P (and N) from offshore can result from storm events (Reed, 1989; Turner et al., 2007), rainfall scouring of exposed marsh sediments followed by export to coastal waters has also been observed (Dankers et al., 1984; Mwamba and Torres, 2002). Sporadic and/or extreme events remain difficult to adequately quantify. Net tidal exchanges of DIP tend to be two- to threefold smaller than particulate P fluxes. The majority of tidal marsh flux studies, particularly when they include measures of porewater drainage, indicate that salt marshes are small net exporters of PO3 (0.03–2.25 g P/m2/year; Chambers et al., 1992; Childers, 1994; Dame, 4 1994; Lillebø et al., 2007). Of the 11 studies on whole tidal creeks reviewed by Dame (1994), only two document a net import of PO3 4 although there can be seasonal reversals of import/export at individual sites (Jordan and Correll, 1991). Both marshes showing annual PO3 4 import were geologically/ecologically young 3 systems. Even young marshes will export PO3 4 when porewater PO4 is high and its exchange is dominated by advection (Osgood, 2000). Phosphate effluxes are most prevalent at the early stages of the flooding tide when low Eh porewater pools are flushed of PO3 4 recently produced from organic matter mineralization (Lillebø et al., 2004). At larger tidal ranges, some marshes respond with enhanced PO3 4 export (Childers, 1994), while others show a switch from export to import (Childers et al., 2000). These opposite patterns probably represent trade-offs at higher tides between enhanced tidal drainage and an expanded oxidized unsaturated zone favorable to PO3 4 sorption to iron oxides. Several factors contribute to seasonality in PO3 4 export. Mineralization is the 3 ultimate source of PO3 , and higher PO effluxes occur during warmer months 4 4 with up to fourfold increases reported in the summer (Lillebø et al., 2004). Seasonal 3 plant dynamics also influence PO3 4 exchange with lower PO4 effluxes from
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vegetated soils (Lillebø et al., 2004). Both vegetated and unvegetated regions of a Portuguese marsh were net sources of PO3 4 to the water column through the year except when autotrophic demand by plants and epiphytes was maximal (Lillebø et al., 2004, 2007). The plant modulation of PO3 4 efflux over diel and seasonal scales (Lillebø et al., 2004) is attributed to both plant assimilation and to O2 pumping into the soils that induces changes in Fe and S cycling and promotes mineral (Fe and carbonate) scavenging of P. The two- to fourfold greater PO3 4 adsorption capacity observed in vegetated soils translates to 50% less PO3 4 export compared to mudflats. Plant effects are seen on seasonal and diel timescales. Although the Ca-bound P pool dominates the mineral-bound P reservoir, water column-sediment PO3 4 fluxes may be controlled in just a thin layer of soil at the surface through redox-sensitive Fe oxide–P sorption dynamics (Chambers et al., 1992; Lillebø et al., 2007). It is not clear whether or how the Ca–P fraction is involved in PO3 4 fluxes. Theoretically, localized pH decreases, for example, due to sulfide oxidation in surface soils (Giblin and Howarth, 1984; Kostka and Luther, 1995) could release this Ca-bound fraction. The extent to which this mechanism might contribute surface P sorption dynamics and subsequent PO3 fluxes to 4 overlying water is not well characterized.
5.2. Internal cycling 5.2.1. Photoautotrophy and mineralization Is the import of P sufficient to support macrophyte production in mature salt marshes? Using the range of macrophyte N uptake presented earlier in this chapter and a plant N : P of 10–20, we estimate macrophyte P demand of 0.1–6.0 g P/m2/ year. Direct PO3 4 uptake is small, as most marshes export DIP, and particle settling supplies 0.6 + 0.8 g P/m2/year (Section 5.1.1). Comparing this supply rate with a median estimate of plant demand (3 g P/m2/year) shows that incoming particulate P is about 20% of plant P demand. Although some marshes can import P at rates nearly sufficient to support macrophyte production (Wolaver and Zieman, 1984), autotrophy in many marshes must be supported by internal recycling of P. Based on the size of the bioavailable P pool and macrophyte P demand, useable P turns over on the order of 15 years (Paludan and Morris, 1999). Mineralization of organic matter is an important source of PO3 4 , but its subsequent availability for plant uptake is controlled by geochemical speciation reactions. P release during mineralization is consistent with the rates of organic matter respiration (aerobic and anaerobic) and C : N : P stoichiometry presented earlier in this chapter. Using the range of reported N mineralization rates and N : P ratios for plants and marsh soils of 10–25 (Buresh et al., 1980; Craft, 2007; Zhou et al., 2007), P mineralization ranges from 0.3–12 g P/m2/year with the majority of mature marshes in the range of 1.0–8.0 g P/m2/year. These values are on the order of 5- to 10-fold higher than P import rates and are comparable to macrophyte demand. Bacteria strongly prefer organic P (Sundareshwar et al., 2003), so only a small fraction of any excess PO3 4 created during mineralization is microbially immobilized. Instead, it is geochemically sorbed and respeciated. These processes represent
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exchanges between sorbed P and porewater PO3 4 that help to regulate bioavailability and affect the resultant exchanges of DIP with tidal waters. 5.2.2. Geochemical cycling Bacteria perform the initial mineralization of organic matter to release PO3 4 . This PO3 4 sorbs readily to humics and oxidized mineral phases and gets incorporated into solid organic fractions of variable lability. Ranges of sorbed P are 0.15–1 mg P/gdw in both vegetated and nonvegetated marsh soils. Exchangeable fractions of sorbed P are on the order of 20–50% depending on mineral composition. Based on radiolabeled 32P experiments, the dissolved PO3 4 , DNRP, Al-, and Fe-bound P pools are potentially available for maintaining equilibrium with porewater PO3 4 and thus available for biological uptake but the Ca–P fraction shows little exchangeability (Jensen and Thamdrup, 1993). Salt marsh soils (relative to oligohaline systems) have lower organic N : P, higher porewater PO3 4 , less Fe oxide-bound P, and more Ca-bound P (Paludan and Morris, 1999; Coelho et al., 2004; Zhou et al., 2007). Despite the smaller contribution of Fe oxide-bound P, this fraction remains an important mediator of P dynamics in salt marshes. Zones of Fe-bound P occur in the uppermost soils in contact with the atmosphere and around oxidized root channels (Krom and Berner, 1980). These zones limit upward diffusion of PO3 4 (Chambers and Odum, 1990; Coelho et. al., 2004). The Fe-bound fraction is the most bioavailable of the sorbed P fractions because of its redox sensitivity and reactivity with sulfur (Figure 4). Salt marshes retain more P in macrophyte-dominated soils where high rates of photosynthesis enhance soil oxidation. Increased Fe oxide formation sorbs more P and 3 draws down porewater PO3 4 , leading to greater retention of PO4 in vegetated soils (Mendelssohn and Postek, 1982; Sundby et al., 1998; Lillebø et al., 2004). Maximum Fe oxide formation and P sorption was observed in a Delaware, USA, marsh in summer when photosynthesis was highest (Kostka and Luther, 1995). Coehlo et al. (2004) measured decreased daytime PO3 4 export in excess of that attributable to direct plant assimilation, and the formation of P-rich root plaques associated with Ca and Fe oxides in vegetated marsh soils has been widely observed (Caetano and Vale, 2002; Lillebø et al., 2007). Some P-rich plaques are retained on long timescales while others appear to be more transient. The reduction of Fe(III) oxides due to biological processes or chemical reactions with sulfides (Sections 2.2.2 and 4.2.1) can liberate PO3 4 and thus reduce the amount of potential P burial (Anschutz et al., 1998; Rozan et al., 2002). 2 PO3 4 –Fe interactions are heavily modified by S. Sulfides from SO4 reduction bind tightly to Fe forming pyrite and/or iron monosulfides (Figures 2 and 4), which 2 inefficiently sorbs PO3 4 . Additionally, SO4 competes for P-binding sites on Fe oxides. Iron sulfide minerals isolate Fe from further redox cycling and limit future binding with 3 PO3 4 (Sundby et al., 1992). The net result of a high S environment is that less PO4 sorbs to solid mineral phases and more remains in porewaters. Redox variations with seasons, tides, and daily photosynthesis generate patterns of Fe and S cycling that modify the P speciation and subsequent release of PO3 4 on the ecosystem scale (Giblin and Howarth, 1984; Scudlark and Church, 1989; Kostka and Luther, 1995).
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Ca-bound P represents a large fraction (40–80%) of the mineral-bound soil P, yet is less well characterized than other pools (Paludan and Morris, 1999; Coelho et al., 2004; Alva`rez-Rogel et al., 2007; Zhou et al., 2007). Sedimentation of Ca–P particles at the seaward end of estuaries traps P that might otherwise migrate upstream to more P-limited ecosystems. High Ca2þ (e.g., seawater) competes for P with organics. P-rich root plaques are also carbonate rich, and Ca–P is soluble at low pH. However, Ca–P is not considered readily exchangeable and is not thought to participate in exchanges with porewater PO3 4 . On the one hand, the Ca–P fraction appears to be dynamic, while on the other hand, it may represent an essential step toward P burial. Given the magnitude of P bound to Ca in salt marshes, its dynamics may represent a relatively unknown but potentially important component of salt marsh P dynamics.
5.3. Burial Burial is the dominant mechanism of P loss for all salt marshes. The amount of residual-bound P (0–0.26 mg P/gdw) provides a lower limit on the extent of P burial (Paludan and Morris, 1999; Coelho et al., 2004; Zhou et al., 2007). Assuming that marsh accretion reflects SLR, a conservative P burial rate of 0.2 g P/m2/ mm SLR can be estimated. If the Ca–P-bound fraction (assumed to be relatively bio-unavailable) is added to the burial inventory, the rate of P sequestration by salt marshes increases by a factor of 2–3. Measured rates of P burial for a variety of salt marshes range between 0.36 and 2.0 g P/m2/year (Craft, 2007 and references therein). These burial rates are generally lower than P burial in freshwater and oligohaline marshes due to higher sediment availability and a greater sedimentary P content in up-estuary environments, as well as increased decomposition rates in higher salinity systems (Merrill and Cornwell, 2000; Craft, 2007; Jordan et al., 2008).
6. M ARSHES IN T RANSITION AND D IRECTIONS FOR FUTURE W ORK Salt marshes are dynamic in terms of geomorphology, biogeochemistry, and their role in the coastal landscape. They are subject to the synergistic effects of global-scale climate forcings and local-scale human impacts. Our understanding needs improvement of how salt marshes respond to increasing CO2,atm, rising sea level, and increased continental nutrient loading in terms of C, N, and P cycling, storage, or burial. Salinity encroachment accompanying SLR will have its most pronounced effect on mesohaline marshes altering the balance of oxidants (e.g., O2 and SO2 4 ) and forms of N and P delivered to marshes. To a large extent, current marsh distribution, accretion, and biogeochemistry reflects the effects of SLR since the last glacial maximum, but future responses are now compounded by additional anthropogenic factors. Human loading of CO2 and nutrients to marshes induces responses in plant dynamics; however, the resultant effects on marsh
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biogeochemistry and long-term maintenance remain to be determined. The effect of combined global and local stressors on the capacity of marshes to transform biologically important elements and subsequently retain or export them within the coastal landscape is the biggest challenge for future salt marsh biogeochemistry work. We advocate an approach for examining coupled elemental cycling when possible to address these challenges. Based on this review, we offer a few specific directions for future work. • Improve characterization of the pathways through which organic C is decomposed. Despite decades of studying SO2 4 reduction (and, more recently, Fe(III) reduction), there are still instances where these “major” processes account for <50% of total C catabolism. Processes such as Mn(III, IV) reduction and the utilization of humic compounds may be more important than previously assumed. • Develop a better understanding of DNRA in marshes, its magnitude, and controls. In particular, a better characterization of its synergy with the sulfur cycle is needed. Changes in these pathways are of particular importance for N retention versus export in mesohaline marshes subject to increasing SO2 4 loads as SLR accelerates and freshwater withdrawls from rivers induce salinity encroachment upstream in estuaries. • Explore anaerobic oxidation pathways of Fe(II), H2S, and Fe–S minerals (FeS, FeS2). Thermodynamically, a variety of compounds including NO 3 , MnO2, and Fe(III) can be used as oxidants for various reduced Fe and S species, but the significance of these anaerobic reactions/processes have not been fully assessed. Because the reduction of these oxidants can also be coupled with organic matter catabolism, there are feedbacks and interactions between the cycling of C, Fe, and S. • Generate a more comprehensive picture of Ca–P dynamics. This is the largest inorganic-bound P fraction. It should be pH sensitive, but it is assumed to be recalcitrant. It may be more dynamic than previously thought, or it may be the pathway to P burial. An insufficient amount of information exists at present to conclude either. • Continue to pursue techniques that examine “nutrient” fluxes out of marshes to the coastal landscape at expanded spatial scales. Various geochemical tracer approaches (e.g., radium) show promise in this regard, but there needs to be a clearer picture of exactly which fluxes the tracers quantify.
ACKNOWLEDGMENTS We thank Robert Twilley, Don Cahoon, and Mark Brinson for their helpful reviews, which significantly improved this chapter. This is contribution no. 1484 from the University of South Carolina’s Belle W. Baruch Institute for Marine and Coastal Studies. This work was supported in part by NSF-DEB 0542635, the DOD-DCERP Program, and the UNCW Center for Marine Science.
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Vale, C., Caterino, F., Cortesa˜o, C., Cac¸ador, I., 1990. Presence of metal-rich rhizoconcretions on the roots of Spartina maritima from the salt marshes of the Tagus Estuary, Portugal. Sci. Total Environ. 97/98, 617–626. Valiela, I., Cole, M.L., McClelland, J., Hauxwell, J., Cebrian, J., Joye, S.B., 2000. Role of salt marshes as part of coastal landscapes. In: Weinstein, M., Kreeger, D.A. (Eds.), Concepts and Controversies in Tidal Marsh Ecology. Kluwer Academic Publishing, Dordrecht, pp. 23–38. Valiela, I., Teal, J.M., 1979. The nitrogen budget of a salt marsh ecosystem. Nature 280, 652–656. Valiela, I., Teal, J.M., Persson, N.Y., 1976. Production and dynamics of experimentally enriched salt marsh vegetation: belowground biomass. Limnol. Oceanogr. 21, 245–252. Valiela, I., Teal, J.M., Volkmann, S., Shafer, D., Carpenter, E.J., 1978. Nutrient and particulate fluxes in a salt marsh ecosystem: tidal exchanges and inputs by precipitation and groundwater. Limnol. Oceanogr. 23, 798–812. Van Raalte, C.D., Valiela, I., Teal, J.M., 1976. Production of epibenthic salt marsh algae: light and nutrient limitation. Limnol. Oceanogr. 21, 862–872. Wang, Z.A., Cai, W.-J., 2004. Carbon dioxide degassing and inorganic carbon export from a marshdominated estuary (the Duplin River): a marsh CO2 pump. Limnol. Oceanogr. 49, 341–354. Wang, W., Morris, J.T., Prokopenko, A., Neubauer, S.C., in review. Stable isotope signatures of CO2 evolution and carbon dynamics in East Coast salt marsh soils, USA. Limnol. Oceanogr. Weis, J.S., Weis, P., 2004. Metal uptake, transport and release by wetland plants: implications for phytoremediation and restoration. Environ. Int. 30, 685–700. Weiss, J.V., Emerson, D., Backer, S.M., Megonigal, J.P., 2003. Enumeration of Fe(II)-oxidizing and Fe(III)-reducing bacteria in the root zone of wetland plants: implications for a rhizosphere iron cycle. Biogeochemistry 64, 77–96. Weiss, J.V., Emerson, D., Megonigal, J.P., 2004. Geochemical control of microbial Fe(III) reduction potential in wetlands: comparison of the rhizosphere to non-rhizosphere soil. FEMS Microbiol. Ecol. 48, 89–100. Weston, N.B., Joye, S.B., 2005. Temperature-driven decoupling of key phases of organic matter degradation in marine sediments. Proc. Natl Acad. Sci. 102, 17036–17040. Weston, N.B., Porubsky, W.P., Samarkin, V.A., Erickson, M., Macavoy, S.E., Joye, S.B., 2006. Porewater stoichiometry of terminal metabolic products, sulfate, and dissolved organic carbon and nitrogen in estuarine intertidal creek-bank sediments. Biogeochemistry 77, 375–408. White, D.S., Howes, B.L., 1994a. Nitrogen incorporation into decomposing litter of Spartina alterniflora. Limnol. Oceanogr. 39, 133–140. White, D.S., Howes, B.L., 1994b. Long-term 15N-nitrogen retention in the vegetated sediments of a New England salt marsh. Limnol. Oceanogr. 39, 1878–1892. Whiting, G., Childers, D., 1989. Subtidal advective water as a potentially important nutrient input to southeastern USA salt marsh estuaries. Estuar. Coast. Shelf Sci. 28, 417–431. Whiting, G., McKellar, H., Spurrier, J., Wolaver, T., 1989. Nitrogen exchange between a portion of vegetated salt marsh and the adjoining creek. Limnol. Oceanogr. 34, 463–473. Widdel, F., Schnell, S., Heising, S., Ehrenreich, A., Assmus, B., Schink, B., 1993. Ferrous iron oxidation by anoxygenic phototrophic bacteria. Nature 362, 834–835. Wolaver, T.G., Dame, R.F., Spurrier, J.D., Miller, A.B., 1988. Sediment exchange between a euhaline salt marsh in South Carolina and the adjacent tidal creek. J. Coast. Res. 4, 17–26. Wolaver, T.G., Hutchinson, S., Marozas, M., 1986. Dissolved and particulate organic carbon in the North Inlet Estuary, South Carolina: what controls their concentrations? Estuaries 9, 31–38. Wolaver, T.G., Zieman, J., 1984. The role of tall and medium Spartina alterniflora zones in the processing of nutrients in tidal waters. Estuar. Coast. Shelf Sci. 19, 1–13. Woodwell, G.M., Houghton, R.A., Hall, C.A., Whitney, D.E., Moll, R.A., Juers, D.W., 1979. The Flax Pond ecosystem study: the annual metabolism and nutrient budgets of a salt marsh. In: Jeffries, R.L., Davy, A.J. (Eds.), Ecological Processes in Coastal Environments. Blackwell Scientific, Oxford, pp. 491–511. Zedler, J.B., 1980. Algal mat productivity: comparisons in a salt marsh. Estuaries 3, 122–131. Zhou, J., Wu, Y., Kang, Q., Zhang, J., 2007. Spatial variations of carbon, nitrogen, phosphorous and sulfur in the salt marsh sediments of the Yangtze Estuary in China. Estuar. Coast. Shelf Sci. 71, 47–59.
C H A P T E R
1 7
T HE R OLE OF F RESHWATER F LOWS ON S ALT M ARSH G ROWTH AND D EVELOPMENT Laurence A. Boorman
Contents 1. Introduction 2. Freshwater Routes in Salt Marshes 2.1. Stream flow 2.2. Groundwater flow 2.3. Rainfall 2.4. Surface flow 3. Associated Processes 3.1. Nutrient transport 3.2. Sediment transport 3.3. Organic matter transport 3.4. Pollutants 3.5. Salinity changes 4. Hydrological Impacts in Salt Marshes 5. Techniques for the Study of Marsh Hydrology 6. Implications of Freshwater Flows for Salt Marsh Management 7. Implications of Freshwater Flows for Salt Marsh Creation 8. The Ecohydrological Approach in Salt Marsh Studies 9. The Way Ahead – Problems and Challenges References
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1. INTRODUCTION Salt marshes are characterized by the presence of plants tolerant of both immersion in water for varying periods and some degree of salinity although freshwater inputs can also be a significant factor in many marshes (Boorman, 2003). Salt marshes have been shown to play a major role in many processes within the estuarine and coastal ecosystems (Boorman, 1999). There are various sources and routes of freshwater into a salt marsh. These can include river flow into the estuary, groundwater flow along a defined aquifer or channel or diffuse seepage, and also directly as a result of rainfall on the marsh and through surface flow from Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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adjacent slopes. However, in general terms, little attention has been paid to freshwater inputs and impacts on salt marshes except from the point of view of the effects of rainfall on the acceleration of seed germination of many plant species and the effect of river flow on the overall salinity of the water body at particular points in an estuary (Boorman and Hazelden, 2004). It has been shown that the tidally driven flow of saltwater can be the agent of transport to and from the marsh itself of sediment, mineral nutrients, pollutants, and particulate or dissolved organic carbon (Hazelden and Boorman, 1999). It would seem probable that where there are significant flows of freshwater in a salt marsh a similar effect may be expected. Additionally as excessive nutrient levels in an estuary (White et al., 2004) can affect the marsh plant communities, it would seem likely that nutrients brought in by freshwater could have a similar impact. Salt marshes are commonly developed on fine-textured sediments with the particle size in the clay and fine silt range, and consequently the permeability of such soil might be expected to be low. An inspection of many salt marshes shows that this is a simplification of the true picture. At low tide, water is seen to seep from the sides of marsh creeks from a variety of fissures and holes in the otherwise slowly permeable marsh clays and silts. These more permeable layers can be either physical or biological in origin. Physically, cracks and fissures develop when the soil dries out and the clay shrinks. Although these will close up on rewetting, they remain a permanent feature of the soil structure. In addition, coarse-textured horizons within the soil (perhaps sand and shell debris deposited during a storm) can give rise to more permeable layers within the soil (Boorman and Hazelden, 2005). Permeable layers, of biological origin, can result from the residual channels left after the death and decay of roots and other underground plant material (Boorman and Hazelden, 2005). They also result from the burrowing activities of varied intertidal fauna, from crabs and molluscs through to the many different groups in the meso- and microfauna. Water movement paths are also created in the marsh soil by the burial of layers of organic matter such as are created when the autumn fall of leaf material is buried by high rates of accretion during equinoctial tides. These layers are quite persistant as they are sometimes visible a hundred millimeters below the surface which, in a marsh with a mean annual accretion rate of around 3 mm, represents an age of the order of 30 years.
2. FRESHWATER R OUTES IN SALT MARSHES The dominant role of saline water (usually sea water) can be modified by the addition freshwater in various ways, and the magnitudes of the extent and impact of freshwater depend largely on the routes and quantities involved (Figure 1) (Boorman and Hazelden, 2005). The hydrodynamics of a marsh system will be determined by the topography of the marsh and its setting, by the variations in the porosity of the soil and the underlying strata, and by the tidal regimes of the area. However, the overall freshwater input to an area will be determined by meteorological conditions, principally the precipitation/evaporation balance. It should also
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Stream flow
Rain fall
Transpiration
Tidal flow Evapor -ation
Sheet flow Seepage
Ground water in
Ground water out
Figure 1 Water exchange routes associated with salt marshes.The background picture is of the marshes at Tollesbury, Essex, England. These are lowland estuarine marshes with a macrotidal regime (tidal range up to 5 m). The main estuary (River Blackwater) and the open sea are to the right of the picture while the gradients to the terrestrial ecosystems are to the left.
be noted that elevated temperatures and low humidity will not only reduce the effective rainfall but they will also tend to elevate the natural salinity by increasing evaporation. Furthermore, the impacts of any salinity changes will be at their greatest under favorable growing conditions while they will be reduced when soil and/or atmospheric temperatures fall to a level at which seed germination and plant growth is inhibited. The details of any freshwater input to a salt marsh will, however, depend on the route involved including direct static effects and dynamic effects through modified fluxes and exchanges. The biggest effects are generally through the long-term semipermanent flows above or below ground level in the form of river/stream flow or the flow of groundwater.
2.1. Stream flow Stream flow in this section includes any discrete aboveground flow of freshwater into a salt marsh; while it is very often in the form of a stream, this can range from the merest of trickles through to that of a major river flow into an estuary fringed with salt marshes. Whatever the size of the freshwater inflow, there are two distinct situations: the first being where the freshwater flows directly into a salt marsh system and the second where the freshwater input is less direct and through admixture with the sea water of an estuary. The distinction is thus between the situation where the freshwater impacts on the salt marsh directly and where the freshwater reduces the salinity of the inshore waters adjacent to the marsh system. In practice, the distinction may be blurred not only because in many salt marshes both routes can occur but also the definition of an individual salt marsh system is variable and flexible. Even where the bank of an estuary is fringed with salt marshes,
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these may collectively be regarded as part of the whole or as separate discrete systems, particularly when the marshes have their own distinct outflows to the estuary (Boorman, 2003). The impact of stream flow on a salt marsh will depend on the proportional contribution of freshwater and the resultant salinity at various points across the marsh. Particularly where the flow of freshwater is directly to the marsh, any seasonality in flow rates will affect the resultant impact with short-term high flow rates having a much smaller impact on the vegetation than the same total volume spread out over a long period. The setting of a salt marsh is the major factor determining the stream flow. Salt marshes in lowland situations, such as are found in the east and south of Britain, tend to have limited inflows of freshwater; in estuarine situations river flows are relatively constant and thus there is limited variability in the extent of the penetration of sea water. Very often the adjoining land is reclaimed land which was formerly salt marsh and in these cases the stream flow is largely a result of land drainage systems in the adjoining farm land. The small flow rates involved result in there being only a limited impact although nutrient and pollutant runoff from agricultural land can have serious environmental effects on salt marshes (Section 3). In mountainous upland situations such as those found in the west of Scotland, the salt marshes tend to be found at the head of the sea lochs and there can often be major stream flows into and across these marshes (Boorman, 2002). The flows involved can be relatively small in daily volume but with a high degree of seasonal variability. Although the catchment areas may be relatively small in area, the enhanced rainfall over high ground can result in these high short-term flow rates with their own special influence on salt marsh processes (Figure 2). The nature of a freshwater flow into and through a salt marsh can also be affected by the subsurface structure with a rather different range of impacts when High rainfall
Rapid streamflow
Periodic high riverflow
Figure 2 Fluxes associated with upland salt marshes, Loch Beag, Isle of Skye, Scotland.The Isle of Skye has extensive sea lochs with limited deposits of often rather coarse sediment in rock basins extending up to 15 km from the open sea and there are tidal ranges of around 4 m. The whole area has a high rainfall (up to 1.0 myr ^1), and with largely impermeable substrates, this results in high runoff with periodic flushing of the salt marsh by high volumes of fresh water.
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there is a porous substrate as compared to marshes underlain by solid impervious rock or other impermeable substrates. The water routes created when there is a porous subsurface layer will be considered more fully in the next section. When stream flows through salt marshes have high short-term flow rates, the bed material of the stream itself will be bedrock rather than marine sediments or terrestrial alluvium as is the case when there are consistently low flow rates, which is the situation in the majority of lowland salt marshes. The pattern of exchanges in these upland marshes is markedly difference from that in lowland marshes where creek water velocities are relatively low and stable (Boorman, 2002). The pattern of water flow can be further modified by the magnitude of the tidal range interacting with the relative volume of freshwater discharge (Fitzgerald et al., 2002).
2.2. Groundwater flow While stream flow over the land surface is the visible input of freshwater into salt marshes, the flow of groundwater below the soil surface may in fact have greater impact at least in some salt marshes. Two conditions need to be fulfilled before groundwater flow into salt marshes can occur: the existence of permeable layer or layers within the substrate and a sufficient water pressure to initiate the flow. The critical factor for the latter is the presence of higher ground adjacent to the salt marsh (Boorman and Hazelden, 2005). Typically, this situation is found where marsh-fringed estuaries are bordered by higher ground such as that of the Stour River in north Essex, England, where local reductions in soil salinity marked by the invasion of Phragmites australis reveal the local upwelling of freshwater (Figure 3).
Figure 3 Scattered clumps of Phragmites communis are found along the upper margins of the macrotidal salt marshes on the estuary of the River Stour, Essex, England, marking the occurrence of local upwelling of fresh ground water moving down from higher ground to the south (right-hand side).
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Not all underground flows will be visible anywhere on the marsh surface or will indeed have any significant direct effect on the marsh. If the permeable soil layers are continuous under the marsh, freshwater may partially or totally pass under the marsh and discharge directly into the sea. In this case, the impact of nonsaline groundwater will be virtually inseparable from that of the stream flow contributing only through a limited reduction in estuarine salinity. The situation described above of estuaries being bordered by higher ground relates to the lowland situation. In upland situations, the occurrence of higher ground adjacent to a salt marsh is more or less universal; however, the underlying material is very often impermeable bedrock. Even when there is a degree of subsoil permeability, the relatively high rainfall tends to result in overground stream or sheet flow being the dominant route of drainage water of terrestrial origin. Wilson and Gardner (2006) demonstrated the importance of tidal activity in maintaining the flow of groundwater through salt marsh sediments. They demonstrated that groundwater flow principally occurred in the creek bank and even when the marsh surface was covered by the tide. However, the porosity of the marsh sediments were an important controlling factor for groundwater fluxes. A sandy layer or layers of underlying sandy strata were shown to have a big influence in groundwater flow (Figure 4). Groundwater movements were less in muddy sediments particularly with decreasing particle size, but the sediment porosity was also affected by the degree of soil compression. While the major movements of groundwater in a salt marsh are in the form of discrete groundwater flows there are also the diffuse movements of pore water through the soil. As with major groundwater movements, the seepage of pore water is driven by pore water pressure gradients but relatively small quantities of
Figure 4 Seepages of pore water during low tide from the creek bank of a salt marsh at Mill Bay, County Down, Northern Ireland. The main seepage is associated with a layer of slightly coarser sediment with a higher porosity over a more compact and less porous layer of fine sediment, but usually some seepage occurs right down the creek bank soil profile.
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water are involved and then over shorter distances. While the generally small particle size found in salt marsh sediments results in limited soil permeability, the occurrence of pore water movement can be seen through the way that a pit dug in the salt marsh can collect water long before it is reached by the rising tide and by the seepage visible along the creek side during low tide. Subsurface water flow can induce complex water table dynamics particularly in relation to the distance from the nearest creek (Ursino et al., 2004). They also showed that, often, there was a persistent unsaturated zone below the soil surface which remained aerobic facilitating root respiration and growth. This persistent unsaturated zone was enhanced by the presence of plants and moreover the presence of pioneer plants along the marsh edges increased the availability of soil oxygen facilitating the development and growth of subsequent plant communities. The dynamics of pore water seepage in marsh sediments has been the subject of a recent paper by Gardner (2005). He developed a mathematical model in which he showed that pore water discharge from the marsh occurred predominantly through the creek face rather than the channel bottom. Furthermore, he suggested that these exchanges took place mainly within a few meters of the creek itself and that the pore water dynamics of the creek bank environment differs markedly from that of the distal areas of the marsh. The situation is, however, likely to be different in marshes based on sandy substrates or where permeable sandy layers underlie the marsh. Less information is available regarding the movement of fresh or brackish as opposed to saline pore water in salt marshes, but this has been studied by Gardner et al. (2002) who showed that these groundwater movements were subject to change through evapotranspiration, seepage, and tidal movements as well as rainfall and surface flow (q.v.).
2.3. Rainfall Atmospheric precipitation provides the input of freshwater into virtually all salt marshes and although the water volumes are generally small compared with the saline water reserve in the marsh soil, the effect of the rainwater is emphasized through its effect being mainly limited to the surface layers of the marsh soil. The effects of rainfall on the marsh are twofold. Firstly there is the reduction of the soil salinity and secondly, there are the effects of the water flow itself across the surface of the marsh. The effects of salinity are mainly on seedling germination and plant growth and this will be considered under Section 3.5. While these are the most studied effects of rainfall on a salt marsh, the appearance of colored sediment-laden water running off the marsh surface into marsh creeks during periods of heavy rainfall suggests that rainfall and the resultant sheet flow can entrain and transport sediment. Mwamba and Torres (2002) have shown that the impact of water droplets rapidly mobilizes recently deposited sediments during low-tide rain storm events facilitating subsequent transport by sheet flow of the released sediment. They concluded that although the freshwater flux over salt marshes is negligible compared to that of the tidal prism, the rainfall may facilitate the redistribution of disproportionately large volumes of sediment. The effect of rainfall appears to be indirect as well as direct. Tolhurst et al. (2006) showed that rain
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showers during low tide were correlated with a general reduction in the erosive threshold of the intertidal cohesive sediments thus facilitating the tidal recycling of sediment.
2.4. Surface flow The flow of sea water to and fro across a salt marsh is perhaps the defining feature of the salt marsh. Not only does the incoming tide bring the sediment necessary for marsh building but also the salinity of the seawater controls the composition of both the flora and the fauna which together characterize the marsh itself. Section 2.3 described the flow of freshwater resulting from rainfall over the marsh surface. While both freshwater and saltwater flow across the surface of the marsh start as sheet flow this is broken down by the marsh vegetation into discrete microchannels that coalesce ultimately into the complexities of the system of salt marsh creeks. The wide-scale deposition of sediment across the whole salt marsh is modified by the development of the creek system which is itself the result of the role of salt marsh vegetation in determining the patterns of accretion within the marsh. Much of the sediment initially deposited is subsequently reworked by the surface flow and the channeled flow of water. Together these are modified by the over-riding influence of the salt marsh vegetation in determining patterns of accretion (Boorman et al., 1998)
3. A SSOCIATED P ROCESSES Both the runoff of freshwater from inland to the salt marsh as well as the flow of groundwater provide a direct route and mechanism for the transport of various materials to the salt marsh and also through the marsh to the sea. The material can be transported both in the dissolved and the suspended solid states although unless the groundwater flows are through distinct channels much of the suspended material is likely to be filtered off. The substances involved can include plant nutrients, dissolved organic matter, potentially toxic pollutants, and suspended sediments; in addition, the freshwater itself can have an important effect in diluting dissolved substances in the marsh pore water.
3.1. Nutrient transport The freshwater runoff, direct or through groundwater, from land which has been subjected to urban or agricultural development often contains elevated levels of nitrate and particulate organic nitrogen relative to marsh surface waters (Page et al., 1995). The nitrate imports from the land are utilized both for the primary production on the marsh as well as being exported to the estuarine waters either as nitrate or ammonium nitrogen (Figure 5). The dynamics of pore water nutrients have not so far been studied in great detail in salt marshes, but it is clear from work on the pore water of an intertidal sand flat
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Heavy metals
Nutrients
Salinity
Uptake nutrients Release
Nutrients
Sediment Plant growth Pesticides Herbicides
Figure 5 Exchanges of plant nutrients, pollutants, and sediments in salt marshes. The background picture is of the marshes at Tollesbury, Essex, England. The main estuary (River Blackwater) is to the right with tidal inputs of mineral nutrients and fine sediments.To the left are gradients to the terrestrial ecosystems, mainly agricultural land, where fertilizers, pesticides, and herbicides are used, while in the background there are some boatyards, formerly a significant source of heavy metals.
that nutrient concentration gradients can generate diffusive fluxes to and from the deeper sediments and that the increased oxygenation during emersion affected nitrification and nitrate reduction rates (Kuwae et al., 2003). Microbial nitrate reduction occurred in the deeper subsurface sediments, and this process was supported by the downward diffusive flux of nitrate from the surface sediment. It might be expected that in the less porous salt marsh soils similar processes might occur at a slower rate; however, even in the more porous sandy sediments, both the soil water content and the levels of the water table change little during immersion suggesting that porosity was not a particularly important controlling factor. In order to understand the effect of freshwater flows on the nutrient economy of salt marshes, it is necessary to consider the effects of the normal tidal fluxes affected by sea water. The fluxes of mineral nutrients between salt marsh and estuary are dependent on the external nutrient loading of the estuary, the degree of eutrophication, and on the release of nutrients by the decay of organic matter within the salt marsh. Overall, in studies on a range of Western European estuaries, there appeared to be a net export of dissolved nitrogen out of the salt marshes (Boorman, 2000). The picture regarding available phosphorus was less clear although it did appear that, at least at certain times of the year, there were net exports of phosphate. The concentrations of both nitrogen and phosphorous in salt marsh creeks will depend on the balance between the supply of that component from inside and outside the marsh and the rate of uptake of the component by the growth of salt marsh vegetation. Nitrogen is generally considered to be the critical factor for algal production in salt water (Doering et al., 1995) but this is not always the case with higher plants. It appears that, at times, low concentrations of phosphorus can be the
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main limiting factor in determining the productivity of salt marsh vegetation with eutrophic estuaries enhancing vegetation growth (Boorman, 2000). While generally the release of nitrogen and phosphorus from the salt marsh occurs during the processes of the decomposition of organic matter, direct losses by the leaching of nitrogen, phosphorus, and also carbon from live plant tissues can also occur (Turner, 1993). The amounts released are high enough to account for significant increases in the activity of the estuarine plankton community and thus are of potential significance for many other estuarine communities. A discharge of nutrient-rich groundwater not only affects the salt marsh, it can also have a significant role in the primary production of adjacent coastal waters. It is very likely that a proportion of these nutrients could feed back and enhance the productivity of the salt marsh itself (Slomp and van Cappellen, 2004). Freshwater flows with a nutrient load equivalent to that of sea water would not significantly affect nutrient fluxes except in situations where enhanced total water volumes were involved thus augmenting flux opportunities. However, in many cases, the runoff or discharge from urban or agricultural areas is eutrophic with enhanced levels of nitrogen or phosphorus. The concentrations of these macronutrients in the groundwater flowing into a salt marsh influence the growth and development of the marsh vegetation (Wolanski et al., 2004). There is a fine line between the marginally enhanced growth of salt marsh plants in this way and the gross changes which result from the input of highly eutrophic freshwater. This can significantly affect the composition of the salt marsh vegetation through the enhanced growth of the less salt-tolerant species and the competitive exclusion of the less vigorous halophytes (Alexander and Dunton, 2006). Not all nitrogen coming into the marsh is taken up there; a proportion is exported to the estuarine waters although often nitrogen coming into a marsh as nitrate nitrogen via the ground water is converted to ammonium nitrogen before being exported to the estuary (Page et al., 1995). Thus the salt marsh is, in effect, a processor facilitating and transforming the exchange of nutrients between land and sea (Boorman, 2000). The changing nature of this role has been illustrated by Jickells and Andrews (2000) who showed that as the extent of salt marsh in the Humber Estuary, England, was reduced by reclamation, the trapping of nitrogen and phosphorus was also reduced, and they suggest that this process could be reversed by large-scale salt marsh creation. Mention has already been made of the serious effects groundwater pollution can have on salt marshes but while efforts are being made to reduce the inputs of excessive nutrient levels as yet there has been little work on controlling groundwater pathways. However, in a situation where the reverse problem has occurred, that of saline intrusions in fresh groundwater, a degree of groundwater control has been achieved by modeling the discharge matrix and then by the selective drawdown through controlled pumping (Zhou et al., 2003).
3.2. Sediment transport Sediment can be transported equally well by freshwater as by salt, but for many marshes, the freshwater inputs are small in comparison with the daily saline tidal flow and thus only limited sediment inputs are possible. Higher relative
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freshwater flows can be found in the loch head tidal marshes in upland areas but the peak stream flows occur with high water velocities and most bed load and suspended sediment goes straight through to the sea. The major input of freshwater sediment occurs in salt marshes associated with major river systems such as the Mississippi, USA, or the Humber, England, where the river flow is of very turbid water with a high sediment load. This is in marked contrast to other estuaries where the sediment load is mainly of marine origin. It has been shown that even in cases where marsh accretion from marine sources is insufficient to compensate for rising sea levels, the introduction of river water with the associated suspended sediment load can enhance marsh accretion and stability thus reversing the rate of wetland loss (DeLaune et al., 2003). The benefits appear to result from a combination of sediment input, lowered salinity, and enhanced levels of extractable phosphorus. Sediment transport in salt marshes is mainly thought of in terms of tidally driven sedimentary processes; however, freshwater in the form of rainfall can also make a significant contribution. It has been shown that rainfall can decrease the erosive threshold and thus remobilize recently deposited sediments (Tolhurst et al., 2006). This will further add to the recirculation of sediment brought in by the tide and initially trapped by the salt marsh vegetation (Boorman et al., 1998). The accretion of the salt marsh surface is thus affected by the combination of sediment input from the river and from the sea, by sediment recirculated from the low-level intertidal mud flats, and by internal processes in the marsh itself (Figure 6).
Uplands
Terrestrial erosion Rain fall
River flow
SALT MARSH
Sediment recirculation
Tidal flats
Internal erosion
Tidal exchange Estuarine water column
Marine erosion
Figure 6 Sediment sources and sediment transport routes in a salt marsh. Boxes indicate locations while ellipses indicate the processes involved. Arrows indicate the major routes for sediment exchanges. “Uplands” is a relative term implying both hill country and lowlands draining to marshes and estuaries.
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3.3. Organic matter transport As well as the ionic transport of plant mineral nutrients, there can be significant fluxes of dissolved inorganic carbon with significant contributions from the degradation of organic carbon (Cai et al., 2003). These studies showed that the groundwater in the marsh in South Carolina are mixtures of sea water and freshwater and that the end-members are modified by the input of CO2 from the degradation of organic matter. Furthermore, the work demonstrated that there were significant groundwater fluxes of dissolved inorganic carbon from the land to the sea via the salt marshes. While much attention has been paid to fluxes of carbon associated with dissolved and fine particulate organic matter in the water column, at times there can also be significant fluxes of coarse floating organic matter comprising plant remains, including the seeds of salt marsh plants, lifted from the surface of the marsh by spring tides. Depending on the wind direction, this material may be carried out to sea when it is dispersed with little visible effect, or with an on shore wind, it can be blown ashore at the upper margin of the marsh in the form of a visible drift line of plant debris (Figure 7).
3.4. Pollutants As well as the loading of plant macronutrients, freshwater flowing into a salt marsh can also contain significant levels of various agricultural chemicals including herbicides such as atrazine and organochlorine insecticides (Boorman, 2003). It is
Figure 7 The wind-driven transport of floating organic matter can lead to significant accumulation of coarse organic matter along the high-tide line particularly when there are spring tides and on-shore winds. This drift line of plant debris is primarily composed of dead leaves and partially decayed remains of vascular plant together with macroalgae and it often includes the seeds of salt marsh plants. This example of the drift line was photographed at Tollesbury, Essex, England.
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difficult to assess the impact of these chemicals on the biota as they only occur at relatively low levels and most of the available data on their specific effects on plant and animal species relate to much higher levels but given their persistence there is at least a need for some caution. Various industrial chemicals such as polychlorinated biphenyls (PCB) also occur widely in the environment and have been detected in salt marsh sediments. It is likely that water flow of terrestrial origin is the main source of input and given the possibility for the redistribution of these pollutants following the reworking of sediments, there would also appear to be a potential for damaging effects. This reworking of sediments can result from cyclical changes in the patterns of salt marsh growth and this process is likely to be enhanced by the impact of rising sea level with the accompanying risks of dangerous pollutants being brought back into circulation (Boorman, 1999, 2003). Various heavy metals have also been shown to occur in salt marsh sediments including cadmium, lead, chromium, and mercury (Windham et al., 2001). While there is no evidence that at the levels recorded there is any effect of these metals on the salt marsh plants which can absorb and accumulate quite high concentrations, it is considered that plant feeders, including grazing stock, could well be deleteriously affected. Studies at the managed realignment site at Tollesbury, Essex, England, show, however, that vegetated salt marsh can developed despite the presence metallic contaminants (Chang et al., 2001). In climatically warmer and drier areas, there is a tendency for marshes to become hypersaline through enhanced evaporation (Hughes et al., 1998). This process can be reduced by the inflow of freshwater; however, river flows are frequently dramatically reduced to provide water supplied to urban developments and for agriculture. When attempts are made to restore river flow levels using treated waste water effluent, the degree of eutrophication can significantly affect the composition of the salt marsh vegetation through the enhanced growth of the less salt-tolerant species and the competitive exclusion of the less vigorous halophytes (Alexander and Dunton, 2006).
3.5. Salinity changes 3.5.1. Seed germination While the flora of the older and higher salt marsh is composed mainly of perennial species which set seed infrequently, the colonizing margin and the lower areas of marsh are dominated by annual species replaced by the germination and establishment of seed each year (Adam, 1990). The dispersal of the seeds of salt marsh species is primarily by the sea, nevertheless, the germination of many of the salt marsh species is facilitated by a reduction in salinity. It has been suggested that this may be of adaptive advantage to reduce losses through seeds germinating while they are still floating in sea water. The lowest areas colonized by salt marsh species are at a level not reached by the smallest of neap tides over a period of a few days; thus the surface salinity can fall if there is rainfall during this period and soil temperatures are high enough. This typically occurs in the early spring although increasingly seed germination can occur in salt marshes during mild periods in the winter months.
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The germination of the seeds of salt marsh plants is at a minimum either during high spring tides which cover the whole marsh with fully saline sea water or during hot dry periods in the summer, even in temperate areas. Under these circumstances, evaporation from shallow pools in the upper marsh left by the retreating tide can raise the salinity to levels considerably in excess of those in sea water resulting in the inhibition of seed germination and the formation and spread of areas devoid of vegetation through the inhibition of seed germination (Adam, 1990). This effect is even more marked in tropical and subtropical areas. However, although the reduction in soil salinity is essential for the germination and growth of most salt marsh plants, it has been shown that exposure to sea can be a prerequisite for the subsequent freshwater-enhanced germination of certain salt marsh species notably Limonium spp. (Boorman, 1971). 3.5.2. Seedling growth Like seed germination, the growth of the seedlings themselves is inhibited by high salinities and there may even be significant seedling mortality before they become better adapted to withstand full exposure to sea water (Adam, 1990). It is also worth noting that the duration of immersion itself can reduce the growth rates of pioneer salt marsh species tolerant of the salinity of sea water. However, other factors are also involved. For example, the burial of the root system by the deposition of fresh sediment can enhance seedling growth of both pioneer and upper marsh species (Figure 8). Clearly, there is a complex of other factors as well as the level of salinity which determines the rate of seedling growth in respect of both root and shoot parameters (Boorman et al., 2001). For seedlings growing in a mobile habitat such as a pioneer salt marsh, the rate of growth, particularly root growth, can be critical to ensure that the plant can withstand tidal water movements and not be washed away (Adam, 1990).
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3.5.3. Plant growth and reproduction While the germination and seedling growth of many salt marsh plants are enhanced by a reduction in salinity, the majority of the perennial species of the more mature higher level salt marshes are little affected by fluctuations in salinity and are well able to withstand regular immersion in sea water. Many salt marsh species can grow and thrive in nonsaline soils, but their ability to withstand saline conditions appears to be at the expense of reduced competitive vigor. Salt marsh plants do not occur naturally in nonsaline soils because they are unable to withstand competition from the more aggressive nonsalt-tolerant plant species. It is sometimes possible to find individual salt marsh plant species growing on the landward side of sea banks, but this almost invariably indicates points of seepages of sea water through the sea bank. There is, however, a small group of perennial plant species which have a degree of salt tolerance although less than that of the majority of salt marsh plant species. Typical of these are species such as P. australis, Bolboschoenus spp., and Schoenoplectus spp., and these species characteristically provide a fringe between salt marsh and adjacent freshwater marshes on the landward side (Adam, 1990). Perhaps more common than this situation is one where there are visible clumps of, for example, P. australis along the edge of the upper marshes marking localized freshwater seepages generally associated with groundwater movement from adjacent high ground (Boorman and Hazelden, 2005). While generally such invasions are kept in check by full salinity in other parts of the marsh, there have been examples reported where the invasion of P. australis has become a serious problem in the management of salt marshes particularly where human activities have resulted in salinity reduction over wide areas (Bart et al., 2006). In salt marshes in temperate, relatively damp and cool, areas the majority of plant species are little affected by the general level of salinity. In drier and warmer areas of the world, the input of freshwater becomes of increasing importance to the salt marsh. In South Africa, it has been shown that the salt marsh plants are only in active growth during the winter rainfall period (Bornman et al., 2002). During the dry season, plants are dependent for their survival on access to saline groundwater. The occurrence of winter rainfall ensures the replenishment of the saline groundwater with freshwater both decreasing the depth of the water table and reducing its salinity thus facilitating plant growth.
4. H YDROLOGICAL IMPACTS IN SALT MARSHES While surface and groundwater flow can provide necessary plant nutrients, excessive nutrient loading can result in hypereutrophic conditions with major effects on the biodiversity. It has been shown that groundwater flows can cause the transport of these nutrients over considerable distances (Mayer et al., 2000). Groundwater dynamics along forest marsh transects were studied by Gardner et al. (2002) using long-term piezometric transects coupled with salinity measurements. This study also showed that there were mechanisms by which excessive nutrient levels could be transported through to nearshore sediments with possible effects on
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marine habitats. This is an extreme situation and more generally salt marshes can be regarded as sinks which control the eutrophication of coastal waters by removing excessive nutrients from the system (Teal and Howes, 2000). As well as affecting the concentrations and fluxes of nutrients, organic matter, and sediment associated with a salt marsh, the hydrology of the marsh can also affect the physical conditions within a marsh. It has been shown that variability in evapotranspiration and tidal flooding can affect the soil volume and consequently the precise level of the surface of the marsh (Paquette et al., 2004). This effect is of primary importance in making accurate measurements of accretion/erosion in marsh development. Such measurements are crucial both in the study of salt marsh processes and in the monitoring of success in salt marsh creation. Undetected changes in marsh levels could also have significant consequences for physical and biological processes on the surface of the marsh, in particular on the patterns of seed dispersal and germination and thus the subsequent resultant patterns of plant colonization.
5. T ECHNIQUES FOR THE STUDY OF M ARSH HYDROLOGY Groundwater flows, with the possibilities of their transporting nutrients over considerable distances, necessitate the use of special techniques to determine their source. In one study, involving the leakage of partially treated sewage, the molecular marker coprostanol was used to assess nutrient inputs to a marsh (Mayer et al., 2000). Radioisotopes have also been used to trace groundwater pathways. Routes and flux rates of submarine groundwater discharge in a Massachusetts salt marsh were determined using four radium isotopes (Charette et al., 2003). These workers also showed that under drought conditions seawater–sediment interactions were important in delivery of certain dissolved substances to coastal waters. In another study in North Carolina, the isotopic composition of dissolved inorganic carbon was used to define a component of the surface water–groundwater system (Gramling et al., 2003). The work demonstrated that, when precipitation was low, artesian groundwater discharge accounted for virtually all the freshwater input to the marsh while in wet periods there was a negligible groundwater contribution. Studies continue to collect long-term real-time data on the ecohydrology of salt marshes and to develop mathematical models to interpret the various processes involved (Crowe et al., 2004). More is known about groundwater dynamics in wet coastal grasslands, enabling the prediction of changes (Mohrlok, 2002). Reeves and Fairborn (1996) installed extensive instrumentation to enable the development of a numerical model to study the groundwater dynamics of the forest marsh interface. Crucius et al. (2005) studied groundwater flow and discharge in a small estuary using radon and salinity measurements as markers and constructed a box model. The model was used to aid the understanding of both rates and locations of discharge and the best fit was achieved on the assumption that groundwater discharge is fresh and it occurs in the channel or adjacent areas of the marsh. The authors concluded that more data were needed for the results to be representative of
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long-term averages. Wilson and Gardner (2006) used numerical simulations to show the effect of tidal activity as the driver of groundwater flow and solute exchanges and underlined the importance of adequate data on sediment porosity in controlling the magnitude of these flows and exchanges. The next major step will be to integrate these various models, possibly through the use of a decisionbased support system, in such a way that for any given salt marsh the underlying ecological processes, including the magnitude and direction of the various fluxes, can be understood sufficiently to develop effective management techniques.
6. I MPLICATIONS OF FRESHWATER FLOWS FOR SALT M ARSH M ANAGEMENT The most direct effect of groundwater on salt marshes is the opportunity it offers for the transport of pollutants into the salt marsh ecosystem. Salt marshes adjacent to intensively used farm land can have significant concentrations of selective herbicides (Fletcher et al., 2004). The transport was by both surface and subsurface routes. While it was not possible to demonstrate a detectable effect on the vegetation, the residual herbicide concentrations measured in this study were above the UK environmental safety guidelines. The implications of groundwater quality for the management of salt marshes can also be inferred indirectly. Studies in Japan showed that the use of excessive fertilizer could affect the use of the water for irrigation (Fujiwara et al., 2002). Seawater intrusion into the aquifer was also shown to be having an impact on water quality but the situation was complicated by the activity of cation exchange phenomena. The relation of the salt marsh and freshwater flows is often seen simply in terms of a stream or river flowing to the sea, through an area of salt marsh, and measurement of the incoming river flow will thus be considered to characterize the freshwater input. However, in practice, salt marsh areas often have many freshwater inputs from a number of distinct areas with very different types of land cover and land use. Such was the case in a study of salt marshes in South Carolina, USA, where, through the development of a conceptual model, it was shown that the monitoring of creek headwaters could give early warning of possible harmful effects on tidal areas with serious implications both for conservation and economically important activities (Holland et al., 2004).
7. IMPLICATIONS OF FRESHWATER FLOWS FOR SALT MARSH C REATION The re-creation of salt marshes on land which was originally salt marsh would in the hydrological sense seem fairly straightforward. However, there can be problems caused by the changes that will have taken place to the sediment and soils while the land has been used for agriculture or grazing (Hazelden and
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Boorman, 2001). The most obvious physical change is the “ripening” of the soil; this is the irreversible drying of the sediment by evapotranspiration during which the bulk density increases and porosity decreases. Soil structure, a semipermanent network of cracks throughout the soils delineating soil “peds,” also develops, and the salt (NaCl) will have been leached from at least the upper layers of the soil. On some newly created salt marshes, the old agricultural soil is rapidly buried by the accumulation of new sediment, which provides a good medium for the germination and growth of salt marsh plants. However, where this does not happen, the establishment of salt marsh vegetation may be hindered in these dry, dense soils. The physical properties of reclaimed marsh soils are little altered by the reversion of the land to salt marsh and their burial by new sediment. However, the relatively dense subsurface layer can affect subsequent creek development. Drainage patterns established on a site prior to its reversion to salt marsh will, to a great extent, control those that subsequently become established (Figure 9). In some sites, salt marsh re-creation may be complicated by other factors. It has been shown that some grassland communities of saline areas are very much dependent on the up-welling of groundwater through a saline peat layer (Beyen and Meire, 2003). In order to compensate for the loss of such areas, it was necessary to make detailed hydrological studies, albeit on a fairly local scale, to locate the relative rare occurrence of sites suitable for this type of habitat creation. Even when there are no such special conditions for the re-creation of salt marsh, the changes in the soil hydrological regime which occurred while the marsh was under agricultural use, and no longer subject to regular tidal flooding, are considerable. The effects of the changes in tidal level were limited to small changes in
Figure 9 A linear creek in the salt marshes at North Fambridge, Essex, England. These marshes were re-created naturally following a break in the sea wall during a major storm over a hundred years ago. It is generally considered that the unnaturally straight form of most of the creeks reflects the lines of the man-made drainage system that had been constructed while the land was under agricultural use. This pattern is still visible today despite the intervening accumulation of more than a meter of new sediment on the former land surface.
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the level of the underlying water table (Blackwell et al., 2004). Consequently, there were major adjustments following the return of tidal flooding. Not only was there the direct effect of the immersion in saltwater but there was also a wide range of changes in both physical and chemical soil properties. Changes in soil water table resulted in the soil environment changes from an oxidizing to a reducing environment. In the short term, there were changes in soil pH, with the topsoil water becoming markedly acid. There were also large decreases in the rates of decomposition of organic matter. All of these effects have serious implications for the establishment of salt marsh vegetation and subsequent salt marsh management. The sustainable long-term management of created salt marshes must be a key part of any such program, and there is a range of issues involved (Boorman and Hazelden, 2004). While the successful establishment of vegetation cover may only take a few years, much longer time periods are needed before anything like full ecosystem function is achieved. A recent study of the rate of ecosystem development in created Spartina alterniflora marshes (Craft et al., 2003) showed that, while most of the functional ecological attributes have achieved equivalence to those in nearby natural marshes in 5–15 years, the levels of pools of organic carbon and nitrogen are still lower than in the natural marshes even 28 years after the marsh creation. This work involved the study of a wide range of ecological processes, and this may not always be possible to achieve on economic grounds when there is extensive and widespread marsh creation. It is important, however, to note that simpler methods of assessment may give misleading results. Studies in a range of healthy and impaired salt marshes in Louisiana showed that the state of the aboveground biomass was not a good indicator of marsh health (Turner et al., 2004). However, the work did show that marshes under stress have a reduced belowground biomass which could be detected long before there was any detectable effect on the vegetation aboveground, thus giving the possibility of applying appropriate management techniques.
8. T HE ECOHYDROLOGICAL APPROACH IN SALT M ARSH STUDIES The case for salt marsh creation in order to compensate for lost or degraded marshes has been well made but at present the remedial measures suggested are considered to be inadequate to restore fully the ecological processes of a healthy robust estuary or to reinstate the full beneficial functions of the estuarine ecosystem (Wolanski et al., 2004). These authors consider that the successful management of estuaries and coastal waters requires an ecohydrology-based catchment-wide approach to ensure the survival of the full range of marine coastal, brackish, and freshwater habitats. This will necessitate changing present practices which are based on local administrative units and on the narrowly focused approaches of managers of specific activities (including fisheries, water resources and urban development). Without this change in thinking and in management concepts, estuaries and coastal waters will continue to degrade whatever management plans are put in place.
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9. T HE W AY AHEAD – PROBLEMS AND C HALLENGES It should be clear from the preceding text that flows of fresh water can and frequently do have a distinctive role in many of the components of the ecological functioning of a salt marsh. The precise details of this do seem to vary considerably with the geological, chemical, and biological parameters of individual salt marshes. Because of this, it is often difficult to extrapolate from the results of detailed studies on specific marshes to the general situation. The implications of the various fresh water-induced effects have also to be seen both from the standpoint of a fundamental understanding of salt marsh function and from the more practical aspects of salt marsh creation and management (Section 7; Broome and Craft, 2009). The advancement of the understanding of salt marsh function depends both on the collection of good long-term sets of real-time data on the levels and fluxes of each of the significant plant nutrients and on the development of functional models of the magnitude and direction of all major marsh fluxes. The achievement of these aims also depends on the availability of adequate data on sediment porosity and potential water pathways. Quite a lot has already been revealed by the various models and simulation of salt marsh fluxes, and logically the next step will be to try and integrate the various models and resolve any anomalies which may appear. Reference has been made to the importance of acquiring adequate data sets to represent the annual variations in the levels and fluxes of important plant nutrients. However, in the times of global climatic change, data sets which were previously regarded as adequate for modeling purposes may need to be re-examined in relation to setting values for climatic extremes.
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Hydrology and the Management of Water Resources, National University of Ireland, Galway, pp. 241–256. Boorman, L.A., 2003. Salt marsh review. An overview of coastal saltmarshes, their dynamic and sensitivity characteristics for conservation and management. Joint Nature Conservation Committee Report No. 334. JNCC, Peterborough, pp. 1–116. Boorman, L.A., Garbutt, A., Barratt, D., 1998. The role of vegetation in determining patterns of the accretion of salt marsh sediments. In: Black, K.S., Patterson, D.M., Cramp, A. (Eds.), Sedimentary Processes in the Intertidal Zone. Geological Society, London. Spec. Publ. 139, 389–399. Boorman, L.A., Hazelden, J., 2004. The sustainable management of biodiversity in natural and created salt marshes. In: Green, D.R. (Ed.), Developing Sustainable Coasts: Connecting Science and Policy. Proceedings of Littoral 2004, Aberdeen, Scotland, pp. 508–513. Boorman, L.A., Hazelden, J., 2005. The importance of groundwater and other ecohydrological impacts in the management of salt marsh plant communities. In: Herrier, J.-L., Mees, J., Salman, A., Seys, J., van Nieuwenhuyse, H., Dobbelaere, I. (Eds.), Proceedings “Dunes and Estuaries” – International Conference on Nature Restoration, Practices in European Coastal Habitats, Kokzijde, Belgium, 19–November 23, 2005. VLIZ Spec. Publ. 19, 335–343. Boorman, L.A., Hazelden, J., Boorman, M., 2001. The effect of rates of sedimentation and tidal submersion regimes on the growth of salt marsh plants. Cont. Shelf Res. 21, 2155–2165. Bornman, T.G., Adams, J.B., Bate, G.C., 2002. Freshwater requirements of a semi-arid supratidal and floodplain salt marsh. Estuaries 25, 1394–1405. Broome, S.W., Craft, C.B., 2009. Tidal marsh creation. In: Perillo, G.M.E., Wolanski, E., Cahoon, D.R., Brinson, M.M. (Eds.), Coastal Wetlands: An Integrated Ecosystem Approach. Elsevier Science, Amsterdam, pp. 715–736. Cai, W.J., Wang, Y.C., Krest, J., Moore, W.S., 2003. The geochemistry of dissolved inorganic carbon in a surficial groundwater aquifer in North Inlet, South Carolina, and the carbon fluxes to the coastal ocean. Geochim. Cosmochim. Acta 67, 631–639. Chang, Y.H., Scrimshaw, M.D., MacLeod, C.L., Lester, J.N., 2001. Flood defence in the Blackwater Estuary, Essex, UK: the impact of sedimentological and geochemical changes on salt marsh development in the Tollesbury Managed Realignment Site. Mar. Pollut. Bull. 42, 470–481. Charette, M.A., Splivallo, R., Herbold, C., Bollinger, M.S., Moore, W.S., 2003. Salt marsh submarine groundwater discharge as traced by radium isotopes. Mar. Chem. 84, 113–121. Craft, C., Megonigal, P., Broome, S., Stevenson, J., Freese, R., Cornell, J., Zheng, L., Sacco, J., 2003. The pace of ecosystem development of constructed Spartina alterniflora marshes. Ecol. Appl. 13, 1417–1432. Crucius, J., Koopmans, D., Bratton, J.F., Charette, M.A., Kroeger, K., Henderson, P., Ryckman, L., Halloran, K., Colman, J.A., 2005. Submarine groundwater discharge to a small estuary estimated from radon and salinity measurements and a box model. Biogeosciences 2, 141–157. Crowe, A.S., Shikaze, S.G., Ptacek, C.J., 2004. Numerical modeling of groundwater flow and contaminant transport to Point Pelee Marsh, Ontario, Canada. Hydrol. Process. 18, 293–314. DeLaune, R.D., Jugsujnda, A., Peterson, G.W., Patrick Jr., W.H., 2003. Impact of Mississippi River freshwater reintroduction on enhancing marsh accretionary processes in a Louisiana estuary. Estuar. Coast. Shelf Sci. 58, 6531–6662. Doering, P.H., Oviatt, C.A., Nowicki, B.L., Klos, E.G., Reed, L.W., 1995. Phosphorus and nitrogen limitation of primary production on a simulated estuarine gradient. Mar. Ecol. Prog. Ser. 124, 271–287. Fitzgerald, D.M., Buynevich, I.V., Davis, Jr., R.A., Fenster, M.S., 2002. New England tidal inlets with special reference to riverine-associated systems. Geomorphology 48, 179–208. Fletcher, C.A., Scrimshaw, M.D., Lester, N., 2004. Transport of mecoprop from agricultural soils to an adjacent salt marsh. Mar. Pollut. Bull. 48, 313–320. Fujiwara, T., Ohtoshi, K., Tang, X., Yamabe, K., 2002. Sequential variation of groundwater quality in an agricultural area with greenhouses near the coast. Water Sci. Technol. 45, 53–61. Gardner, L.R., 2005. A modeling study of the dynamics of pore water seepage from intertidal marsh sediments. Estuar. Coast. Shelf Sci. 62, 691–698. Gardner, L.R., Reeves, H.W., Thibodeau, P.M., 2002. Groundwater dynamics along forest-marsh transects in a southeastern salt marsh, USA: Description, interpretation and challenges for numerical modeling. Wetl. Ecol. Manage. 10, 145–149.
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Gramling, C.M., McCorkle, D.C., Mulligan, A.E., Woods, T.L., 2003. A carbon isotope method to quantify groundwater discharge at the land-sea interface. Limnol. Oceanogr. 48, 957–970. Hazelden, J., Boorman, L.A., 1999. The role of soil and vegetation processes in the control of organic and mineral fluxes in some western European salt marshes. J. Coast. Res. 15, 15–31. Hazelden, J., Boorman, L.A., 2001. Soils and “managed retreat” in south-east England. Soil Manage. 17, 150–154. Holland, A.F., Sanger, D.M., Gawle, C.P., Ledberg, S.B., Santiago, M.S., Riekerk, G.H.M., Zimmerman, L.E., 2004. Linkages between tidal creek ecosystems and the landscape and demographic attributes of their watersheds. Exp. Mar. Biol. Ecol. 28, 151–178. Hughes, C.E., Binning, P., Willgoose, G.R., 1998. Chacterisation of the hydrology of an estuarine wetland. J. Hydrol. 211, 34–49. Jickells, T.J., Andrews, J.E., 2000. Nutrient fluxes through the Humber, UK estuary – past, present and future. Ambio 29, 130–135. Kuwae, T., Kibe, E., Nakamura, Y., 2003. Effects of emersion and immersion on the pore water nutrient dynamics of an intertidal sand flat in Tokyo Bay. Estuar. Coast. Shelf Sci. 57, 929–940. Mayer, T., Bourbonniere, R.A., Crowe, A.S., 2000. Assessment of sewage derived phosphorous input to Point Pelee marsh. Int. Wetl. Rem. Conf., 205–214. Mohrlok, U., 2002. Prediction of changes in groundwater dynamics caused by relocation of river embankments. Hydrol. Earth Sys. Sci. 7, 67–74. Mwamba, M.J., Torres, R., 2002. Rainfall effects on marsh sediment redistribution, North Inlet, South Carolina, USA. Mar. Geol. 189, 267–287. Page, H.M., Petty, R.L., Meade, D.E., 1995. Influence of watershed run-off on nutrient dynamics in a southern California salt marsh. Estuar. Coast. Shelf Sci. 41, 163–180. Paquette, C.H., Sundberg, K.L., Boumans, R.M.J., Chmura, G.L., 2004. Changes in salt marsh surface elevation due to variability in evapo-transpiration and tidal flooding. Estuaries 27, 82–89. Reeves, H.W., Fairborn, L.W., 1996. Application of an inverse model, SUTRAP, to a tidally driven groundwater system. IAHS Publ. 237, 115–123. Slomp, C.P., van Cappellen, P., 2004. Nutrient inputs to the coastal ocean through groundwater discharge: controls and potential impact. J. Hydrol. 295, 64–86. Teal, J.M., Howes, B.L., 2000. Salt marsh values: retrospection from the end of the century. In: Weinstein, M.P., Kreeger, D.A. (Eds.), Concepts and Controversies in Tidal Marsh Ecology, Kluwer, Dordrecht, pp. 9–19. Tolhurst, T.J., Friend, P.L., Watts, C., Wakefield, R., Black, K.S., Paterson, D.M., 2006. The effect of rain on the erosion threshold of cohesive intertidal sediments. Aquat. Ecol. 40, 533–541. Turner, R.E., 1993. Carbon, nitrogen and phosphorus leaching rates from Spartina alterniflora salt marshes. Mar. Ecol. Prog. Ser. 92, 135–140. Turner, R.E., Swenson, E.M., Milan, C.S., Lee, J.M., Oswald, T.A., 2004. Below-ground biomass in healthy and impaired salt marshes. Ecol. Res. 19, 29–35. Ursino, N., Silvestri, S., Marini, M., 2004. Subsurface water flow and vegetation patterns in tidal environments. Water Resour. Res. 40, W05115. White, D.L., Porter, D.E., Lewitus, A.J., 2004. Spatial and temporal analyses of water quality and phytoplankton biomass in an urbanized versus a relatively pristine salt marsh estuary. J. Exp. Mar. Biol. Ecol. 298, 255–273. Wilson, A.M., Gardner, L.R., 2006. Tidally driven groundwater flow and solute exchange in a marsh: numerical simulations. Water Resour. Res. 42, W01405. Windham, L., Weis, J.S., Weis, P., 2001. Patterns and process of mercury release from the leaves of two dominant salt marsh macrophytes, Phragmites australis and Spartina alterniflora. Estuaries 24, 787–795. Wolanski, E., Boorman, L.A., Chicaro, L., Langlois-Saliou, E., Lara, R., Plater, A.J., Uncles, R.J., Zalewski, M., 2004. Ecohydrology as a new tool for sustainable management of estuaries and coastal waters. Wetl. Ecol. Manage. 12, 235–276. Zhou, X., Chen, M., Liang, C., 2003. Optimal schemes of groundwater exploitation for the prevention of seawater intrusion in the Leizhou Peninsular in southern China. Environ. Geol. 43, 978–985.
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T IDAL F RESHWATER W ETLANDS Dennis F. Whigham, Andrew H. Baldwin, and Aat Barendregt
Contents 1. Introduction 2. Hydrogeomorphic Setting 3. Biodiversity 3.1. Plants 3.2. Animals 4. Primary Production and Nutrient Cycling 5. Threats and Future Prospectus References
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1. INTRODUCTION Tidal freshwater wetlands (hereafter referred to as TFW) are the orphans of coastal wetland ecosystems. In many places, they are not recognized as a distinct type of coastal wetland and in most developed parts of the world they have been historically heavily impacted by human activities, resulting in their destruction or degradation. One consequence of their orphan status is that the literature on the distribution and ecology of TFW is relatively scant compared to the large number of publications on saline and brackish coastal wetlands. There are, however, current efforts to remedy this situation by compiling and summarizing the literature on TFW. First, Barendregt et al. (2006) recently summarized information on TFW in Europe and North America. In North America, many publications appeared in the 1980s, particularly from studies of TFW along the Atlantic coast (e.g., Odum, 1988; Odum et al., 1984 and references in Yozzo and Steineck, 1994). Another synthesis, relying heavily on the material compiled in Odum et al. (1984), for North American TFW appeared in Mitsch and Gosselink (2000). Prior to the Barendregt et al. (2006) summary, only one European review (Meire and Vincx, 1993) was available. The most recent effort to summarize the work on TFW will be an edited volume by Barendregt et al. (2009). The overview that we present in this chapter is based on the earlier reviews cited above, information summarized in Barendregt et al. (2006), and on selected materials Coastal Wetlands: An Integrated Ecosystem Approach
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from the forthcoming book. A recent book focuses on tidal swamps of the southeastern United States (Conner et al., 2007). This chapter has five sections. We begin with an overview of the hydrogeomorphic settings in which TFW occur. We describe where TFW are known to occur, with the knowledge that a description of their global distribution is incomplete because a global inventory is lacking. In Section 3 we describe elements of biodiversity. TFW often have high plant species biodiversity and high community diversity, but few species are known to be restricted to TFW. Because of the importance of annual species, vegetation is also very often dynamic and we consider the relationship between the diversity of the seed bank and the diversity of extant vegetation in one well-studied TFW. Compared to plants, fewer studies have focused on animals in TFW. In this review, we focus on fish, birds, and mammals. Ecological processes are the theme of Section 4. Some of the highest levels of net annual primary production in temperate zone wetlands have been measured in TFW, but the level of productivity varies widely depending on geographic location and within-wetland habitat variation. TFW have also been shown to provide important water quality functions. In Section 5 we examine threats to TFW. Given their location near the upper limit of tide estuaries (i.e., also historically often the upper limit of navigation), TFW have been almost completely destroyed in some countries and the remaining areas are now viewed as being important and worthy of intervention to assure their survival or restoration. In other parts of the world, human impacts have been minimal and the major threats to TFW are increasing stresses associated with global environmental changes, such as sea-level rise and intrusion of brackish water into areas that are currently tidal freshwater habitats. In Section 5 we also consider approaches that have been and are being used to conserve and restore TFW, a topic treated in detail in Baldwin et al. (2009). We end with a prospectus.
2. HYDROGEOMORPHIC SETTING TFW are almost always restricted to the upper limit of tide where coastal brackish water meets freshwater flow from nontidal rivers (Figure 1), resulting in a tidal freshwater zone where there is bidirectional flow of freshwater. These conditions primarily exist when there is sufficient freshwater flow from a river, where there is a relatively flat and long gradient from the ocean inland, and where there is a tidal range of 0.5 m or more (Odum et al., 1984; Mitsch and Gosselink, 2000; Barendregt et al., 2006). The tidal freshwater zone probably occurs in most rivers with an appropriate geomorphic setting but the extent of the zone would vary seasonally in response to annual rainfall patterns. For example, the tidal freshwater zone varies seasonally in estuarine systems in arid Mediterranean climates. During dry periods freshwater flows are so low that brackish or saline water extends to the upper limit of tide. In wet periods, freshwater flows are large enough to create a tidal freshwater zone with in the tidal portion of the river. Patterns of sedimentation within tidal freshwater zones also influence the development and dynamics of TFW (Pasternack, 2009). Pasternack described two
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Average annual salinity
Marsh type Nontidal freshwater
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<0.5 ppt Oligohaline <5.0 ppt Mesohaline <18.0 ppt
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Figure 1 Distribution of wetlands along a salinity gradient from the open ocean to a nontidal river. Tidal freshwater wetlands occur in the tidal freshwater zone. Source: Odum et al. (1984).
landscape positions in which TFW form, deltaic and fringe. TFW develop on dynamic deltaic deposits that form at the mouth of tidal basins where the sediment carrying capacity of river has been exceeded. Fringing TFW occur at any location in the tidal freshwater zone where the local supply of sediment is greater than the transport capacity of the water. Studies of sediment cores from TFW demonstrate that they are a recent landform, ranging in age from a little more than 100 years to almost 4,000 years (Pasternack, 2009). The influences that humans have had on sediment dynamics in the tidal freshwater zone have been especially important in recent history. Khan and Brush (1994) analyzed sediment cores from a TFW on the Patuxent River (Maryland, USA) and found that sedimentation rates between 630 and 1603 years AD ranged from 0.05 to 0.08 cm/year. Rates of sedimentation increased dramatically after European settlement; which signaled the onset of land clearing. The range of sedimentation rates for periods of time between 1690 and 1990 AD varied from 0.13 cm/year between 1686 and 1694 to a high of 0.77 cm/year in
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1972–1973. In only one time period (1924–1927) did Khan and Brush measure sedimentation rates lower than 0.40 cm/year. There has never been a global inventory of TFW and consequently there are no estimates of their worldwide extent. In North America, TFW are abundant along the mid-Atlantic Coast from southern New England to Florida (Odum et al., 1984). With the exception of the St. Lawrence Estuary (Glooschenko et al., 1993) TFW along the New England coast are fewer and smaller in extent because there are few geomorphic settings that are appropriate for their development (Leck and Crain, 2009). TFW occur along the Gulf Coast of the United States but they are hydrologically and geomorphically distinct from the definition for TFW given earlier. TFW on the Gulf Coast typically having a tidal amplitude that is less than 0.5 m, they occur far inland in areas where there is little slope to the land, and they are not associated with specific river systems (Sasser et al., 2009). Along the west coast of the United States, TFW are not abundant in areas where a Mediterranean climate dominates river hydrology (Leck et al., 2009). In Mediterranean climates, the tidal freshwater zone in rivers can be extensive during the rainy season but it disappears or is very narrow during the dry season. As a result, saline and brackish waters intrude far into river systems during the dry season and the wetlands are typically dominated by species that are associated with brackish tidal wetlands. TFW are more extensive along the larger rivers in the Pacific Northwest (e.g., Columbia) and British Columbia (Fraser), but there have been few detailed studies of TFW in those areas (Leck et al., 2009). The largest extent of tidal freshwater habitat in the United States occurs in Alaska (Hall, 2009) where acreage is probably greater than the estimate for all TFW in the lower 48 states. We assume that extensive TFW also exist in northeastern Russia where the landscape is very similar to Alaska, but we know of no assessment of TFW for Russia or any other part of eastern Asia. One of us (D.F.W.) has seen TFW dominated by herbaceous plant species on Hokkaido Island in northern Japan and forested TFW on Iriomote Island in southern (subtropical) Japan, but we are unaware of any assessments of either their extent or ecology in Japan. TFW were historically common in northwestern Europe but many have been destroyed during centuries of human activities (Barendregt et al., 2006) and some of the remaining areas continue to be used for cultural activities (Figure 2). In The Netherlands, TFW were diked and drained in ancient times, using some of the first techniques to control water movement (Barendregt et al., 2006). Port development and diking eliminated most of the original TFW habitats in Germany, Belgium, and England. In The Netherlands, the massive Delta project resulted in the elimination or deterioration of most of the remaining TFW and they only remain in Belgium because the Scheldt estuary was not closed as part of the Delta project. Along the river Elbe remnants of many TFW still occur, although they suffer from the deepening of the channel for shipping (Garniel and Mierwald, 1996). The presence and abundance of TFW in other parts of the world are poorly documented. Junk (1983) briefly described the presence of TFW on the Atlantic Coast of South America, but he offered no details on their locations or extent. Characteristics of TFW in the Rı´o de la Plata Estuary in Argentina have been described (Kandus and Malva´rez, 2004; Pratolongo et al., 2007).
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Figure 2 Tidal freshwater wetland on the Oude Maas (The Netherlands) in winter. On the right side of the tidal stream is a coppiced stand of osier (Salix). In historical and modern times, managed osier beds are sources of stems used for a variety of purposes (e.g., basketry, mats used in dike construction and maintenance). Source: A. Barendregt.
3. BIODIVERSITY 3.1. Plants Plant diversity was often a topic for study by researchers of TFW on the Atlantic Coast of eastern North America. where factors associated with tides, such as increased soil aeration, combined with lack of salt water stress result in high species diversity and high primary production (the latter discussed in Section 4) (Odum et al., 1995). TFW almost always have a higher diversity of plants than brackish or saline tidal wetlands (Odum et al., 1984). TFW in Europe seem to have lower diversity than their counterparts in North America, most likely due to high rates of sedimentation and the highly eutrophic conditions in most European TFW, a condition that often results in dominance by fewer species (Barendregt et al., 2006). While overall plant species diversity is high in TFW, diversity varies from one habitat to another and the variation can be explained by differences in the relationship between habitat setting and hydrology. A typical cross section through a TFW in the United States is shown in Figure 3 and similar zonation patterns occur in European TFW (Barendregt, 2005; Barendregt et al., 2006). Vegetation in the open water, low marsh, and high marsh habitats is dominated by herbaceous species with diversity increasing from the open water to high marsh habitats (Simpson et al, 1983a). At the upper extreme of tide, TFW are often dominated by woody species with areas being dominated by shrubs or by trees (Barendregt et al., 2006; Conner et al., 2007; Leck et al., 2009). Extensive lists of species for TFW vegetation can be found in Leck et al. (2009) and Odum et al. (1984). In general, subtidal habitats and low marsh areas that are exposed briefly at low tide (Figure 4) are dominated by species with relatively large leaves that are held above the water (e.g., Nuphar lutea
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Bald cypress–Black gum Wax-myrtle
Wild rice Giant cutgrass 4 Cattail Sedges–Rushes Big cordgrass Rose mallow Jewelweed 3 Bur marigold Tearthumb Smartweed 2
Arrow arum Pickerelweed
Spatterdock 1 Rooted aquatics MHW MLW
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Figure 3 Cross section of a typical tidal freshwater wetland showing major habitats and distributions of species. Source: Odum et al., (1984). Scientific names are as follows: bald cypress =Taxodium distichum, black gum = Nyssa sylvatica, wax myrtle = Morella (Myrica) cerifera, wild rice = Zizania aquatica, giant cutgrass = Leersia oryzoides, cattail =Typha spp., sedges ^ rushes = Carex spp. ^ Juncus spp., big cordgrass = Spartina cynosuroides, rose mallow = Hibiscus moscheutos, jewelweed = Impatiens capensis, bur marigold = Bidens laevis, tearthumb = Polygonum arifolium and Polygonum sagittatum, smartweed = Polygonum punctatum, arrow arum = Peltandra virginica, pickerelweed = Pontederia cordata, spatterdock = Nuphar lutea, rooted aquatics = for example, Myriophyllum spicatum, Vallisneria americana, MLW = Mean Low Water and MHW = Mean HighWater.
(L.) Sm., Peltandra virginica (L.)Schott, and Pontederia cordata L. in North America). In northwestern Europe, most open water systems have no aquatic plants due to the high sedimentation rates and eutrophication. Low marshes in northwestern Europe with extensive mudflats that are flooded twice a day are dominated at the upper border by Schoenoplectus triqueter (L.), Schoenoplectus lacustris (L.) Palla, and Bolboschoenus maritimus (L.) Palla. The low marsh in the United States has many of the same species that occur in open water areas but it is also the habitat in which Zizania aquatica L. and Polygonum punctatum Elliott are often abundant. The creek bank associated with the low marsh commonly has several low-growing species (Callitriche heterophylla Pursh, Gratiola neglecta Torrey, Lindernia dubia (L.)Pennell, Ludwigia palustris (L.)Ell.) that form groundcovers. Figure 5 shows the transition zone between a low marsh and a high marsh habitat along the Nanticoke River (Maryland, USA). The high marsh habitat has the highest species diversity in both the United States and Europe. In the United States, high marsh habitats consist of a diversity of annual (e.g., Ambrosia trifida L., Bidens laevis (L.)BSP, Impatiens capensis Meerb., Pilea pumila
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Figure 4 Mudflat with Schoenoplectus lacustris in the Elbe River (Germany). Source: A. Barendregt.
(L.)A. Gray, Polygonum arifolium L., Polygonum sagittatum L.) and perennial (Acorus calamus, Leersia oryzioides (L), Swartz, Peltandra virginica, Typha spp.) species. The lowest diversity on the high marsh occurs when clonal perennials form dense patches in which few other species become established. Examples of patch-forming perennials are species of Typha and Phragmites australis (Cav.)Trin. ex Steudel. In Europe, perennials (e.g., Lythrum, Phalaris, Epilobium, Typha, Symphytum, Valeriana, Sparganium) dominate high marsh habitats (Barendregt et al., 2006).
Figure 5 Low marsh to high marsh transition on the Nanticoke River, Delaware (USA). The dominant species in the low marsh (left side of photograph) is Nuphar lutea. The dominant species in the high marsh (right side of photograph) is Acorus calamus. In the high marsh annual species become dominant toward the end of the growing season. Source: A.H. Baldwin.
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Figure 6 Forested tidal freshwater wetland on the Nanticoke River, Maryland (USA). Note relatively open forest canopy and diverse assemblage of herbaceous plants in the understory. Source: A.H. Baldwin.
TFW habitats dominated by shrubs and trees also can also have high species diversity (Peterson and Baldwin, 2004) because, in addition to trees and shrubs and a few herbs that rarely occur in more open habitats (e.g., Osmunda regalis var. spectabilis (Willd.) Gray), they contain many of the herbaceous species that occur on the high marsh. Examples of shrub and tree species (e.g., Acer rubrum L., Viburnum dentatum L., Fraxinus pennsylvanica Marsh) in the United States (Figure 6) and Europe can be found in Barendregt et al. (2006), Odum et al. (1984), Rheinhardt (1992), and Barendregt (2005). Living and fallen trees and shrubs are often the focal points for the development of mounds or hummocks that are the preferred habitat for a variety of herbs that are less tolerant of flooding, including species of Carex, grasses (Cinna arundinacea L.), ferns (Osmunda cinnamomea L., O. regalis, Thelypteris palustris Schott), and Viola cucullata Aiton (Rheinhardt, 1992; Leck et al., 2009). Almost all plant species in TFW also occur in non-TFW. In the United States only one TFW plant species (Aeschynomene virginica (L.) BSP) has been listed as endangered (Griffith and Forseth, 2003). In Europe, almost all TFW plant species also occur in other types of freshwater wetlands and there are few species that have been identified as threatened or endangered. In TFW of the Elbe estuary close to Hamburg there are two endemic species, Oenanthe conioides Lange and Deschampsia wibeliana (Sond.) Parl (Burkart, 2001). Both species are listed in Germany and incorporated into the EU Habitats Directive to preserve the species. In TFW in The Netherlands and Belgium, a variety of Caltha palustris L. (var. araneosa), occurs that produces roots on the nodes below the flower that after the breaking of the stem can be transported by the tides permitting dispersal to almost all TFW in the region (van Steenis, 1971). An endangered European species that occurs in low marsh habitats that experience some erosion is Schoenoplectus triqueter (Deegan and Harrington, 2004). This species is distributed from the Elbe in Germany to the Gironde in the south of France.
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There have been few studies of rarity in TFW. However, long-term studies in New Jersey (USA), based on monitoring of seed bank and vegetation in a created TFW in the Delaware River that is adjacent to a natural TFW, provide insights into the dynamics of TFW vegetation including rare species (Leck et al.,1988; Leck and Leck, 2005 and references cited therein). In 1988, 426 species were reported in the study area (Leck et al., 1988); by 2005 the number had increased to 875 with a number of rare (29) and endangered (8) species mostly from the created wetland (Leck and Leck, 2005). The increased number of taxa was the result of continued exploration of the study area, disturbances of the natural wetlands due to road construction, and inclusion of vegetation in upland habitats within the marsh complex. From a wetland perspective, one of the most interesting results was the number of rare species that appeared in the constructed wetlands attributed to the availability of new substrates for colonization. Over a 5-year period, 177 species emerged from soil seed bank samples from the constructed wetland, compared to 96 species from soils in the natural wetland over more than 15 years. Eighty-three of these species only occurred in soils in the constructed wetland, an indication of the potential for dispersion of rare species within the tidal freshwater zone of the river. In both the constructed and natural wetlands, the number of established plant species was much lower than the number of species that emerged as seedlings from the soil samples. Leck and Leck (2005) suggested that the differences were due, in part, to the absence of suitable field germination sites in both types of wetlands. The presence of a relatively high number of rare and endangered species at the constructed wetland, which was only one small portion of a larger tidal freshwater zone in the Delaware River, suggests the importance of maintaining a diversity of TFW habitats to assure the persistence of a diverse flora, especially species requiring open habitats with limited competition.
3.2. Animals Animals associated with TFW have received less attention than flowering plants, as have other groups of plants (e.g., algae, bryophytes, ferns) as well as fungi and microorganisms. Much of the information on animals is adapted from Barendregt et al. (2006), Odum et al. (1984), Mitsch and Gosselink (2000), and Swarth and Kiviat (2009) and we focus on three groups of animals, fish, mammals, and birds. In general, benthic invertebrates may be less diverse in TFW compared to brackish and saline tidal wetlands, but the diversity of terrestrial invertebrates is higher (Barbour and Kiviat, 1986; Ysebaert et al., 1998, 2003; Barendregt et al., 2006). Similar to plants, few animals are restricted to TFW, but beyond species identification, few animal groups have been examined in detail. Many animals that occur in TFW are wide ranging and also are common in brackish and saline wetlands or in non-TFW (Odum et al., 1984; Mitsch and Gosselink, 2000). Examples of wide-ranging fish, mammals, and birds in TFW in the United States are the yellow perch (Perca flavescens Mitchill), the predaceous river otter (Lutra canadensis Schreber), and herbivorous mammals such as the common muskrat (Ondatra zibethicus L.) and beaver (Castor canadensis Kuhl). Examples of widespread bird species are the great blue heron (Ardea herodias L.) and osprey (Pandion haliaetus L.).
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Odum et al. (1984) listed 125 fish species for TFW, but only 59 were regular components of the fish community. The families with the greatest number of species were the Cyprinidae, Centrarchidae, and Ictaluridae. The fish fauna of TFW includes nonnative species such as the common carp (Cyprinus carpio L.). Cyprinid species of killifish (e.g., Fundulus heteroclitus L., F. diaphanous Lesueur) are examples of abundant forage fish (Lippson and Lippson, 1997). Several fish species that are commercially important spawn in the tidal freshwater zone or as juveniles forage in that zone. Striped bass (Morone saxitalis Walbaum), yellow perch (Perca flavescens Mitchill), and American shad (Alosa sapidissima Wilson) are all abundant at one or more life history stages. Striped bass and American shad spawn in tidal freshwater zone. Yellow perch are potanadromous, migrating only within coastal rivers. They spawn in nontidal freshwater portions of rivers but larvae, juveniles, and adults forage in the tidal freshwater zone (Piavis, 1991). The fish community has been described for many estuaries in Europe (Elliot and Dewailly, 1995), including the Minho, Lima, and Gironde (Lobry et al., 2003), Loire, Scheldt (Maes et al., 1998), Rhine, Meuse, and Elbe (Thiel and Potter, 2001), and Forth and Tyne (Pomfret et al., 1991). Similar to North America, some marine species that enter the estuary migrate through the tidal freshwater zone on a seasonal basis, either as adults or juveniles. Freshwater fish that occur in European TFW habitats also occur in nontidal freshwater habitats. Diadromous fish (anadromous and catadromous) that spend part of their life cycle at sea and part in nontidal portions of rivers use TFW habitats during migrations, and a few species, for example Allis shad (Alosa alosa L.) and Twait shad (Alosa fallax Lace´pe`de), are protected at a European level, since they are listed in the EU Habitats Directive. Odum et al. (1984) listed 10 mammals that are common in TFW. The most obvious mammals are the species that have visual impacts on the vegetation. The common muskrat builds lodges (up to 2 m high and 1–3 m wide) that are composed mostly of mounds of vegetation (Figure 7). They also construct feeding stations (Figure 7) that are not as large as lodges but are also distinct features within the vegetation mosaic. One consequence of lodge and feeding station construction is that muskrats apparently harvest more aboveground biomass than belowground biomass even though rhizomes of several species are preferred food (Lynch et al., 1947). Muskrats, however, appear to have little impact on plant diversity, but feeding activities alter soil nitrogen dynamics (Connors et al., 1999). Beavers, eradicated throughout much of the Atlantic coast of the United States, have made a remarkable recovery in recent decades and are now common in TFW where they build lodges and consume large amounts of woody biomass. In some situations, beaver lodges occur on the high marsh, but they are most often found in areas where dams have been placed across shallow tidal areas, often located near food sources (Figure 8). In addition to the larger mammals, some smaller species (e.g., marsh rice rat (Oryzomys palustris Harlan)) can impact vegetation through their feeding activities. One of the authors (D.F.W.) has observed marsh rice rats consuming seedlings and juveniles of wild rice on numerous occasions, to the point where population size was reduced on a small scale. Birds are also conspicuous components of TFW. Species may nest in TFW vegetation, forage on vegetation, or hunt animal prey. The most common birds that
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Figure 7 Muskrat lodges at the Jug Bay National Estuarine Reserve on the Patuxent River, Maryland (USA). Several lodges can be seen in the photograph (dark mounds), as well as a lower feeding station to the left of the lodge in the foreground. Two people standing in the marsh provide scale. Source: A.H. Baldwin.
Figure 8 Beaver dam across a tidal freshwater creek at the Jug BayWetlands Sanctuary on the Patuxent River in Maryland (USA). The dam can be seen running diagonally across the photograph from right to left in the foreground; the low marsh plant Nuphar lutea is visible in the background on the far side of the small pond created by the dam. In the foreground, to the right of the dam is a tidal creek and associated tidal freshwater wetland. To the left of the dam, the wetlands no longer experience any significant tidal influence. Source: A.H. Baldwin.
nest in TFW vegetation on the Atlantic Coast are the least bittern (Ixobrychus exilis Gmelin), Canada goose (Branta canadensis L.), Virginia rail (Rallus limicola Vieillot), king rail (Rallus elegans Audubon), marsh wren (Cistothorus palustris Wilson), common yellowthroat (Geothlypis trichas L.), and red-winged blackbird (Agelaius phoeniceus L.).
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In North America, the American black duck (Anas rubripes Brewster) is the most abundant waterfowl species, especially in the winter (Swarth and Burke, 2000) but many other species of waterfowl use TFW for resting and feeding during migration. Gulls are often abundant throughout the year. Large numbers of gulls. Larus argentatus (Pontoppidan, 1763, Denmark), L. atricilla (Linnaeus, 1758, Bahamas), L. delawarensis (Ord, 1815, Philadelphia) congregate in TFW at low tide (Wondolowski, 2001) and Chris Swarth (personal communication) has observed up to 12,000 L. atricilla resting in Maryland (USA) TFW prior to continuing on to summer breeding grounds. TFW are used by a large number of wintering songbirds that roost individually or in small-to-large flocks and forage activity in all types of habitats. Red-winged blackbirds and common grackles (Quiscalus quiscula L.) are two species that congregate in enormous flocks, often roosting in tall emergent vegetation (Meanley, 1965). Red-winged blackbirds and bobolink (Dolichonyx oryzivorus L.) specialize on wild rice seeds (Figure 9) in the late summer (Meanley, 1993 as cited in Swarth and Kiviat, 2009).
Figure 9 The annual grass wild rice, Zizania aquatica, with inflorescences. Source: A.H. Baldwin (shown in the picture).
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The European TFW are rich in bird species, with nesting birds in the reedbeds, marshlands, and tidal forests during summer and additional migrating birds in the winter season (Ysebaert et al., 2000; Barendregt et al., 2006). Ducks and waders especially in the winter period are important. In some locations, their numbers are so great that TFW are of extreme conservation value, supporting 1% of the world population for many species.
4. P RIMARY P RODUCTION AND NUTRIENT C YCLING TFW are one of the most productive types of wetlands in the temperate zone, but the level of biomass production varies among species and habitats, with a range of approximately 400–2500 g/m2 for aboveground biomass (Whigham et al., 1978; Odum et al., 1984; Mitsch and Gosselink, 2000; Barendregt et al., 2006). Open water and low marsh habitats are also less productive because those sites are inundated for longer periods compared to high marsh and shrub- and tree-dominated habitats. The most productive habitat appears to be the high marsh (Neubauer et al., 2000) where annual net biomass production of more than 3,000 g/m2 has been measured for individual species (e.g., Sickels and Simpson, 1985). An interesting feature of many high marsh habitats is that there is less annual variation in aboveground production compared to brackish and saline tidal wetlands. Whigham and Simpson (1992) reported results from an 11-year study of a TFW in the Delaware River estuary. TFW had a lower coefficient of variation in annual production compared to brackish and saline tidal wetlands. They suggested that the low annual variation was due to the three factors. Plants in TFW are not stressed by salinity, nutrients levels are high in TFW because most of them are located near urban and suburban areas with high nutrient loading rates, and they have a high diversity of annual species. The high diversity of annual species allows for compensation among species resulting fairly constant levels of biomass production even though the abundance and growth of one or more species may vary considerably from year to year. This feature of TFW appears to be unique among tidal wetland ecosystems. Most of the in situ organic matter produced by plants in TFW flows through the detritus food chain and leaves of most species have high decomposition rates (Odum and Heywood, 1978; Findlay et al., 1990). Internal cycling of nutrients seems to be sufficient to support the high rates of primary production (Morris and Lajtha, 1986; Bowden et al., 1991) and an experiment to test the hypothesis that production is nitrogen limited did not result in an increase in above ground biomass, an indication of the relatively high N status of many TFW (Chambers and Fourqurean, 1990; Bowden et al., 1991; Morse et al., 2004). Sediment deposition is also an important source of nutrients in TFW (Orson et al., 1990; Darke and Megonigal, 2003; Morse et al., 2004; Pasternack, 2009) and sediment inputs may enable surface elevations in TFW to keep pace with an accelerated rate of sea-level rise. Because they are accreting environments, TFW substrates also
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accumulate heavy metals (Khan and Brush, 1994), resulting in elevated concentrations in plant tissues (Simpson et al., 1983b). High rates of primary production in many habitats and high rates of sedimentation both indicate that TFW would be net sinks for nutrients. The few nutrient-budget studies that have been conducted on TFW, suggest that there is a net accumulation of nutrients during the growing season and a net release of nutrients during the fall and winter months (Simpson et al., 1978, 1983b). Their primary contribution to coastal estuarine systems seems to be as sites for nutrient transformation, with particulate forms of nutrients dominating flood tides and dissolved nutrients dominating ebb tides (Odum et al., 1984; Bowden et al., 1991). Bowden et al. (1991) concluded that the nitrogen budget of a TFW in Massachusetts (USA) was “largely independent of the nitrogen budget of the river”. In Europe, the TFW appeared to be the essential link between the rivers and the estuaries, where nutrients and suspended matter are transformed to detritus. The silica cycle appeared to be especially important in TFW (Barendregt et al., 2006).
5. T HREATS AND FUTURE P ROSPECTUS Barendregt et al. (2006) described the fate of many TFW in northwest Europe and the Atlantic coast of the United States. The location of TFW near the upper limit of tide in major river systems resulted in their destruction, especially in European estuaries, as cities and associated port facilities developed. A small-scale example of the long-term effects of human activities in the United States can be observed in the Anacostia River, a tributary of the Potomac River within the city of Washington, DC. Most of the original 1,000 ha of TFW have been destroyed by dredging and filling and the sites that were not destroyed are now highly degraded (Baldwin, 2004). Ongoing efforts are currently directed toward protection and restoration of TFW on the Anacostia (Baldwin, 2004). An important component of the restoration activity is a watershed-level effort to improve water quality. The responses of existing TFW to improvements in water quality will be interesting to document because in recent history, TFW typically occur in areas that are rich in nutrients and sediments. Improvement in water quality will result in a decrease in nutrients and a reduction in sediment inputs. These changes are likely to result in shifts in species abundances within TFW habitats. Restoration of TFW is showing promise as a tool in reducing losses of TFW and restoring habitat and species diversity (see Baldwin et al., 2009). In other parts of the United States, different activities were responsible for the historical losses of TFW. In New England, the placement of dams near the upper limit of tide was responsible for the losses of TFW (Leck and Crain, 2009). In South Carolina, large areas of TFW were diked and converted into rice fields and only recently have there been efforts to restore them to their original condition (Whigham et al., 2009). Diking and filling were also responsible for losses in the
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Sacramento–San Joaquin river Delta (California, USA) and restoration efforts are underway to restore the important ecological functions associated with the TFWs (Jassby and Cloern, 2000; Hammersmark et al., 2005); however, loss of organic matter due to aeration has lowered substrate levels exacerbating flooding and negatively affecting restoration efforts. In the United States, national and state regulations have resulted in the protection of most coastal wetlands and wetland losses in the costal zone have been reduced dramatically (Dahl, 2006), but degradation continues. Restoration of TFW in The Netherlands and Belgium, where there had been significant historical losses is currently under consideration (Barendregt et al., 2006). Technical procedures for restoration of TFW in Europe are well established and the ecosystems become well established within a few years when the conditions are optimal (e.g., Zonneveld, 1999). A range of restoration projects are planned or even in the implementation phase (Storm et al., 2005; Van den Bergh et al., 2005). However, in North America a suite of restoration techniques have been attempted, with varying degrees of success in establishing ecosystem structure and function comparable to undisturbed TFW (Baldwin et al., 2009). On a global scale, as indicated in Section 2, there are undoubtedly large TFW areas that have not been heavily impacted by human activities. The extensive TFW that exist in Alaska, for example, do not face any immediate threat. Similar conditions probably prevail in other northern areas (e.g., Siberia) where human impacts have been minimal. In those areas the greatest threats are undoubtedly associated with the consequences of global climate change. In Alaska, increasing temperatures are causing glaciers to melt at a faster rate, resulting in increased sediment input to coastal estuaries. The long-term impacts of increased sediment loading are unknown and the consequences can be either positive or negative. Increased sediment input will enable TFW to increase their relative surface elevation and thus keep pace with rising sea levels. Too much sediment, however, can result in negative impacts of vegetation. In the Kenilworth Marsh in Washington, DC, for example, sediment was placed at a higher elevation in one of the cells that was constructed for restoration purposes. The cell became dominated by invasive species as a result of the higher surface elevation in the cell. Threats associated with global climate change may impact TFW in other ways (Neubauer and Craft, 2009). Increasing rates of sea-level rise may result in the intrusion of brackish water into tidal freshwater portions of rivers, resulting in the replacement of TFW by brackish wetlands. In situations where it is not possible for TFW to migrate upstream (e.g., near the upper end of tide in most rivers on the Atlantic coast of the United States, due to placement of dams, fault lines, and development), they will eventually be eliminated or their area will decrease significantly. Ongoing efforts to protect and restore TFW present a paradox against the backdrop of the potential effects associated with global climate change. We strongly recommend that these efforts around the world be undertaken in the context of the dynamic location of TFW within the coastal zone. Effective conservation, restoration, and management will require vigilance and commitment by governmental and nongovernmental organizations.
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REFERENCES Baldwin, A.H., 2004. Restoring complex vegetation in urban settings: the case of tidal freshwater marshes. Urban Ecosyst. 7, 125–137. Baldwin, A.H., 2009. Restoration of tidal freshwater wetlands in North America. In: Barendregt, A., Whigham, D.F., Baldwin, A.H. (Eds.), Tidal Freshwater Wetlands. Backhuys Publishers, Leiden. Baldwin, A.H., Hammerschlag, R.S., Cahoon, D.R., 2009. Evaluation of restored tidal freshwater wetlands. In: Perillo, G.M.E., Wolanski, E., Cahoon, D.R., Brinson, M.M. (Eds.), Coastal Wetlands: An Integrated Ecosystem Approach. Elsevier Science, Amsterdam, pp. 801–832. Barbour, S., Kiviat, E., 1986. A survey of Lepidoptera in Tivoli North Bay (Hudson River Estuary). In: Cooper, J.D. (Ed.), Polgar Fellowship Reports of the Hudson River National Estuarine Research Reserve Program, 1985. Hudson River Foundation, New York, NY, USA, pp. IV.1–IV.26. Barendregt, A., 2005. The impact of flooding regime on ecosystems in a tidal freshwater area. Int. J. Ecohydrol. Hydrobiol. 5, 95–102. Barendregt, A, Whigham, D.F., Baldwin, A.H. (Eds.), 2009. Tidal Freshwater Wetlands. Backhuys Publishers, Leiden. Barendregt, A., Whigham, D.F., Baldwin, A.H., van Damme, S., 2006. Wetlands in the tidal freshwater zone. In: Bobbink, R., Beltman, B., Verhoeven, J.T.A., Whigham, D.F. (Eds.), Wetlands: Functioning, Biodiversity Conservation, and Restoration. Springer-Verlag, Berlin, Germany, pp. 117–148. Bowden, W.B., Vo¨ro¨smarty, C.J., Morris, J.T., Peterson, B.J., Hobbie, J.E., Steudler, P.A., Moore III, B., 1991. Transport and processing of nitrogen in a tidal freshwater wetlands. Water Resour. Res. 27, 389–408. Burkart, M., 2001. River corridor plants (Stromtalpflanzen) in Central European Lowland: a review of a poorly understood plant distribution pattern. Global Ecol. Biogeogr. 10, 449–468. Chambers, R.M., Fourqurean, J.W., 1990. Alternative criteria for assessing nutrient limitation of a wetland macrophyte (Peltandra virginica (L.) Kunth). Aquat. Bot. 40, 305–320. Conner, W., Doyle, T., Krauss, K., 2007. Ecology of Tidal Freshwater Swamps of the Southeastern United States. Springer, Dordrecht. Connors, L.M., Kiviat, E., Groffman, P.M., Ostfeld, R.S., 1999. Muskrat (Ondatra zibethicus) disturbance to vegetation and potential net nitrogen mineralization and nitrification rates in a freshwater tidal marsh. Am. Midl. Nat. 143, 53–63. Dahl, T.E., 2006. Status and Trends of Wetlands in the Conterminous United States 1998 to 2004. US Department of the Interior, Fish and Wildlife Service, Washington, DC, USA. 112pp. Darke, A.K., Megonigal, J.P., 2003. Control of sediment deposition rates in two mid-Atlantic coast tidal freshwater wetlands. Estuar. Coast. Shelf Sci. 57, 259–272. Deegan, B.M., Harrington, T.J., 2004. The distribution and ecology of Schoenoplectus triqueter in the Shannon estuary. Proc. R. Ir. Acad., B, 104, 107–117. Elliot, M., Dewailly, F., 1995. The structure and components of European estuarine fish assemblages. Neth. J. Aquat. Ecol. 29, 397–417. Findlay, S., Howe, K., Austin, H.K., 1990. Comparison of detritus dynamics in two tidal freshwater wetlands. Ecology 71, 288–295. Garniel, A., Mierwald, U., 1996. Changes in the morphology and vegetation along the humanaltered shoreline of the Lower Elbe. In: Nordstrom, K.F., Roman, C.T. (Eds.), Estuarine Shores – Evolution, Environments and Human Alterations. John Wiley & Sons, Chichester, pp. 375–396. Glooschenko, W.A., Tarnocai, C., Zoltai, S., Glooschenko, V., 1983. Wetlands of Canada and Greenland. In: Whigham, D.F., Dykyjova´, D., Hejny´, S. (Eds.), Wetlands of the World: Inventory, Ecology and Management, vol. 1. Africa, Australia, Canada and Greenland, Mediterranean, Mexico, Papua New Guinea, South Asia, Tropical South America, United States. Kluwer Academic Publishers, Dordrecht, pp. 415–515.
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Odum, W.E., Heywood, M.A., 1978. Decomposition of intertidal freshwater marsh plants. In: Good, R.E., Whigham, D.F., Simpson, R.L. (Eds.), Freshwater Wetlands. Ecological Processes and Management Potential. Academic Press, New York, NY, pp. 89–97. Odum, W.E., Odum, E.P., Odum, H.T., 1995. Nature’s pulsing paradigm. Estuaries 18, 547–555. Odum, W.E., Smith III, T.J., Hoover, J.K., McIvor, C.C., 1984. The Ecology of Tidal Freshwater Marshes of the United States East Coast: A Community Profile. FWS OBS-83-17, US Fish and Wildlife Service, Washington, DC, 177pp. Orson, R.A., Simpson, R.L., Good, R.E., 1990. Rates of sediment accumulation in a tidal freshwater marsh. J. Sedimentol. Petrol. 60, 859–869. Pasternack, G.B., 2009. Hydrogeomorphology and sedimentation. In: Barendregt, A., Whigham, D.F., Baldwin, A.H. (Eds.), Tidal Freshwater Wetlands. Backhuys Publishers, Leiden. Peterson, J.E., Baldwin, A.H., 2004. Variation in wetland seed banks across a tidal freshwater landscape. Am. J. Bot. 91, 1251–1259. Piavis, P.G., 1991. Yellow perch (Perca flavescens). In: Funderburk, S.L., Mihursky, J.A., Jordan, S.J., Riley, D. (Eds.), Habitat Requirements for Chesapeake Bay Living Resources. National Oceanic and Atmospheric Administration, Annapolis, MD, pp. 14–1 to 14–15. Pomfret, J.R., Elliott, M., O’Reilly, M.G., Phillips, S., 1991. Spatial and temporal patterns in the fish communities in two UK North Sea estuaries. In: Elliott, M., Ducrotoy, J.P. (Eds.), Estuaries and Coasts: Spatial and Temporal Intercomparisons. Olson & Olson, Fredensborg, pp. 277–284. Pratolongo, P., Kandus, P., Brinson, M.M., 2007. Net aboveground primary production and soil properties of floating and attached freshwater tidal marshes in the Rı´o de la Plata estuary, Argentina. Est. Coasts 30, 618–626. Rheinhardt, R., 1992. A multivariate analysis of vegetation patterns in tidal freshwater swamps of lower Chesapeake Bay, U.S.A. Bull. Torrey Bot. Club 119, 192–207. Sasser, C.E., Gosselink, J.G., Holm, G.O., Visser, J.M., 2009. Freshwater tidal wetlands of the Mississippi River delta. In: Barendregt, A., Whigham, D.F., Baldwin, A.H. (Eds.), Tidal Freshwater Wetlands. Backhuys Publishers, Leiden. Sickels, F.A., Simpson, R.L., 1985. Growth and survival of giant ragweed (Ambrosia trifida L.) in a Delaware River freshwater tidal wetland. Bull. Torrey Bot. Club 112, 368–375. Simpson, R.L., Good, R.E., Leck, M.A., Whigham, D.F., 1983a. The ecology of freshwater tidal wetlands. BioScience 33, 255–259. Simpson, R.L., Good, R.E., Walker, R., Frasco, B.R., 1983b. The role of Delaware River freshwater tidal wetlands in the retention of nutrients and heavy metals. J. Environ. Qual. 12, 41–48. Simpson, R.L., Whigham, D.F., Walker, R., 1978. Seasonal patterns of nutrient movement in a freshwater tidal marsh. In: Good, R.E., Whigham, D.F., Simpson, R.L. (Eds.), Freshwater Wetlands. Ecological Processes and Management Potential. Academic Press, New York, NY, pp. 243–258. Storm, C., Van der Velden, J.A., Kuijpers, J.W.M., 2005. From nature conservation towards restoration of estuarine dynamics in the heavily modified Rhine-Meuse estuary, The Netherlands. Arch. Hydrobiol. Supplement 155 (Large Rivers) 15, 305–318. Swarth, C.W., Burke, J., 2000. Waterbirds in Freshwater Tidal Wetlands: Population Trends and Habitat Use in the Non-breeding Season. Technical Report of the Jug Bay Wetlands Sanctuary, Lothian, MD, 37pp. Swarth, C.W., Kiviat, E., 2009. Animal communities – North America. In: Barendregt, A., Whigham, D.F., Baldwin, A.H. (Eds.), Tidal Freshwater Wetlands. Backhuys Publishers, Leiden. Thiel, R., Potter, I.C., 2001. The ichthyofaunal composition of the Elbe Estuary: an analysis in space and time. Mar. Biol. 138, 603–616. Van den Bergh, E., Van Damme, S., Graveland, J., De Jong, D., Baten, I., Meire, P., 2005. Ecological rehabilitation of the Schelde Estuary (The Netherlands-Belgium; Northwest Europe): linking ecology, safety against floods and accessibility for port development. Restor. Ecol. 13, 204–214. van Steenis, C.G.G.J., 1971. De zoetwatergetijdedotter van de Biesbosch en de Oude Maas, Caltha palustris L. var. araneosa, var.nov. Gorteria 5, 213–219. Whigham, D.F., Baldwin, A.H., Swarth, C., 2009. Conservation of tidal freshwater wetlands in North America. In: Barendregt, A., Whigham, D.F., Baldwin, A.H. (Eds.), Tidal Freshwater Wetlands. Backhuys Publishers, Leiden.
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C H A P T E R
1 9
B IOGEOCHEMISTRY OF T IDAL F RESHWATER W ETLANDS J. Patrick Megonigal and Scott C. Neubauer
Contents 1. Introduction 2. Carbon Biogeochemistry 2.1. Carbon inputs 2.2. Carbon outputs 3. Processes Governing Organic Carbon Metabolism 3.1. Anaerobic respiration 3.2. Processes regulating methane production, oxidation, and emission 4. Nitrogen Biogeochemistry 4.1. Nitrogen exchanges 4.2. Nitrogen transformations 4.3. Nutrient regulation of plant production 5. Phosphorus Biogeochemistry 6. Silicon Biogeochemistry 7. Biogeochemical Effects of Sea-Level Rise 8. Concluding Comments Acknowledgments References
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1. INTRODUCTION By virtue of their unique position in coastal landscapes, tidal freshwater wetlands (TFWs) are hot spots of biogeochemical transformation and exchange (Pasternack, 2009). As with all tidal wetlands, they occur in geomorphic settings that promote exchanges of water, solutes, solids, and gases with adjacent terrestrial and aquatic ecosystems, groundwater, and the atmosphere. However, many features of biogeochemical processes in TFWs are distinct from the nontidal freshwater and saline ecosystems that lie within the coastal landscape because of a unique combination of flushing by tides, the chemical milieu of freshwater, and their position at the limit of tidal influence. Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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Tidally driven hydrology is expected to produce more open element cycles in TFW than nontidal freshwater wetlands, and is the basis of the Outwelling Hypothesis (Kalber, 1959; Odum, 1968) that has informed a great deal of research in saline tidal wetlands. Low concentrations of sulfate in fresh water make TFWs stronger sources of CH4 (a potent greenhouse gas) than saline tidal wetlands (Bartlett et al., 1987). Exports of dissolved inorganic carbon (C) and alkalinity from TFWs have relatively dramatic effects on water chemistry because adjacent estuarine waters are relatively poorly buffered compared to saline waters closer to oceans. The broad outlines of biogeochemical cycles in TFWs are influenced by many of the same factors that constrain element cycles in nontidal and saline tidal wetlands. TFWs support herbaceous marshes, swamp forests, and shrub lands that differ with respect to primary productivity, root–leaf–wood C allocation, and C quality. They occur on both mineral and organic soils, the chemical composition of which is expected to affect a host of ecosystem processes, including the contribution of Fe(III) to anaerobic microbial respiration (Neubauer et al., 2005b). TFWs are found on eutrophic and oligotrophic rivers that place limits on nutrient availability and productivity. This review focuses on the exchanges, transformations, and storage of the major elements, recognizing that these processes govern the contribution of TFWs to the metabolism of coastal landscapes.
2. CARBON BIOGEOCHEMISTRY The most complete C budget of a tidal freshwater wetland is for Sweet Hall Marsh on the Pamunkey River in Virginia, USA (Neubauer et al., 2000, 2002; Neubauer and Anderson, 2003; Figure 1). One strength of this C budget is that it is based, in part, on repeated measurements of ecosystem-level CO2 and CH4 exchange. This approach avoids several problems with estimating C input from biomass harvests (i.e., how to account for biomass turnover and translocation; see discussion in Neubauer et al., 2000), and is especially insightful for understanding ecosystem-level C cycling. We present the Sweet Hall C budget as a case study and a heuristic device for organizing our review of C cycling in TFWs.
2.1. Carbon inputs Gas exchange studies are particularly useful for understanding the integrated metabolism of ecosystems. Two-thirds of annual C inputs to Sweet Hall Marsh come from in situ gross primary production (GPP), which represents the photosynthetic CO2 assimilation of macrophytes and microalgae (Figure 1). About 37% of macrophyte GPP is consumed in growth and maintenance respiration, leaving 625 g C/m2/year in net primary production (NPP), but also root exudates and possibly mycorrhizae. In a given year, NPP is supported by a combination of currentyear photosynthates and C that was translocated from storage organs such as rhizomes. Accounting for both photosynthesis and translocation, macrophyte NPP
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Atmosphere
CO2
1062 1135 996 66
72 CH4 1207
Plants
Primary producers
DIC 197 Sediment C
Algae
Soil CO2 and CH4
517 Soil organic matter 229
Tidal waters
Soil
Figure 1 Carbon budget for Sweet Hall Marsh, USA, showing the major pools and fluxes. All fluxes are in units of g C/m2/year. Ecosystem boundaries for the purpose of this budget are the soil ^atmosphere interface, the soil ^ tidal water interface, and the 30 cm depth contour (which is based on the 1963 137Cs peak). Most of the belowground biomass and the most active zone of biological activity lies within the top 30 cm of the marsh.
at Sweet Hall Marsh was 557–736 g C/m2/year (1,150–1,500 g biomass/m2/year), which is about double the peak aboveground biomass of the site (Neubauer et al., 2000). NPP in TFWs was reviewed recently by Whigham (2009). Sediment deposition is an important vehicle for importing allochthonous particulate organic C into TFW soils, and it enhances organic C preservation and nutrient removal through burial. TFWs that keep pace with sea-level rise are net C sinks, at least on an areal basis, because organic matter is buried in accreting sediments (Stevenson et al., 1988; Craft, 2007; Neubauer, 2008). Sweet Hall Marsh imports 517 g C/m2/year via sediment deposition (Neubauer et al., 2002), a full one-third of all organic C inputs to the site (Figure 1). However, the magnitude of this flux varies dramatically according to the geomorphic setting of the TFW (Pasternack, 2009). For example, proximity to the turbidity maximum that forms at the freshwater-saltwater interface of tidal rivers has a profound effect on C deposition rates (Darke and Megonigal, 2003; Morse et al., 2004). The sources of allochthonous C compounds imported with TFW sediment deposits have not been well characterized, but presumably include upland soils, dissolved organic carbon (DOC) sorbed to mineral particles, and plankton. The age and chemical composition of these sources varies, affecting the extent to which they are ultimately decomposed or preserved as soil organic matter. A meta-analysis of the 13C signature of >300 bulk organic samples from salt marshes, mangroves, and seagrass beds suggested that the organic matter preserved in coastal sediments is
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dominated by allochthonous sources at sites with £10% organic soil C (Boullion and Boschker, 2006). Because allochthonous C represents 33% of all C inputs to Sweet Hall but only 10% of microbial soil respiration (54–71 g C/m2/year; Neubauer et al., 2002), it seems that allochthonous C compounds are generally more recalcitrant to decomposition than autochthonous, mainly plant-derived, compounds. Buried organic matter is also a sink for N, P, and other elements in organic tissues (see Sections 4 and 5).
2.2. Carbon outputs The collective respiration of plants, microbes, and animals at Sweet Hall Marsh is supported by total organic C inputs (GPP þ sediment-associated C) of 1,579 g C/m2/year (Figure 1). As discussed above, roughly one-third of GPP is consumed by the plants themselves in growth and maintenance respiration, processes that release CO2 directly to the atmosphere, soil atmosphere, or soil solution. Most of the remaining GPP takes the form of plant biomass that supports the heterotrophic respiration of bacteria, fungi, insects, grazing snails, and a variety of other organisms (Hines et al., 2006). An uncertain fraction of GPP may be lost from plants as root exudates. Organic C inputs from plants and sediments are eventually subjected to decomposition and microbial degradation, producing soil organic matter, dissolved inorganic carbon (DIC) and DOC, and CH4. At Sweet Hall Marsh, 15% (229 g C/m2/year) of all organic C inputs are buried by accreting soil and enter a very slowly decomposing soil organic matter pool. The remaining 85% is cycled in timeframes of hours to months via plant, animal, and microbial respiration, and other microbial degradation processes such as fermentation. Rates of organic C burial in TFWs along the Atlantic and Gulf coasts of North America, and on the Scheldt River, EU, ranged from 10 to 930 g C/m2/year (Table 1, Neubauer, 2008 and references therein). In addition to sequestering organic C, sediment accumulation adds elevation to wetlands soils at approximately the rate of sea-level rise (Morris et al., 2002). 2.2.1. Exports of CO2, DIC, DOC, and POC C compounds exported from tidal wetlands influence the chemical composition of the atmosphere and adjacent estuaries. The plant and microbial respiration that takes place in saturated or flooded soils generates DIC, which partitions into CO2, HCO–3, and CO2– 3 according to pH. At Sweet Hall Marsh, 12% of all organic C added to the marsh is exported as DIC. A portion of the DIC pool is emitted directly to the atmosphere as CO2, while the remainder is exported to the York River (Figure 1; Neubauer and Anderson, 2003). Evidence is mounting that freshwater and low-salinity tidal wetlands are dominant sources of DIC to estuarine waters. Extrapolating DIC export from Sweet Hall Marsh (197 g C/m2/year) to all tidal marshes of the York River Estuary suggests that 47% of excess water column DIC (i.e., DIC unexplained by conservative mixing of freshwater and marine end members) is imported from wetlands (Neubauer and Anderson, 2003). Several other studies suggest there is significant
Vertical accretion and nutrient burial rates in tidal freshwater and oligohaline wetlands in the United States (g C, N, or P/m2/year)
Location
Method
North River, MA Tivoli Bay, Hudson River, NY Two sites, Delaware River, NJ Otter Point Creek, MD
Modelb 210 Pb
Three sites, Choptank River, MD Eight sites, Choptank River, MD Twenty-five sitesd, Patuxent River, MD Jug Bay Marsh, Patuxent River, MD Sweet Hall Marsh, Pamunkey River, VA
210
Pb
210
Pb
210
Pb
137
Cs, Pb 210 Pb 210
Pollene 137
Cs
Accretion (mm/year)
C buriala
N burial
P burial
Reference
(1.3–3.6) 4.5 + 1.4 (3.6–6.9) (6–12)c
ND 142 + 90 (33–285) (139–204)
(3.1–8.6) 11.6 + 3.6 (8.0–15.9) (7.1–8.8)
(0.3–0.8) 2.3 + 1.0 (1.1–3.6) (3.2–6.1)
Bowden et al. (1991) Merrill (1999)
5.0 + 4.5 (2.1–10.2) 9.2 + 2.7 (6.1–10.9) 8.4 + 4.7 (3.2–21.5) 8.5 + 6.0 (1.1–21.9) (3.7–8.9)
163 + 153 (65–339) 366 + 103 (265–470) ND
(2.7–11.7)
(0.5–2.1)
(19.2–27.1)
(0.2–2.0)
358 + 258 (9–930) (70–249)
21 + 11 (7–44) 16.9 + 11.9 (0.5–36.1) (8–26)
1.7 + 1.1 (0.5–4.3) 4.3 + 3.9 (0.1–12.6) (0.6–2.3)
Merrill (1999), Merrill and Cornwell (2000) Merrill (1999), Merrill and Cornwell (2000) Malone et al. (2003)
8.5
229 + 45
18.1 + 3.1
ND
Biogeochemistry of Tidal Freshwater Wetlands
Table 1
Church et al. (2006)
Merrill (1999) Khan and Brush (1994) Neubauer et al. (2002, 2005a)
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Table 1
(Continued ) (g C, N, or P/m2/year) Accretion (mm/year)
C buriala
N burial
P burial
Reference
Cs
(3.5–4.6)
(103–122)
(7–8)
(0.4–0.8)
Craft (2007)
137
Cs
(6.5–10.6)
(153–239)
(9–16)
(0.5–1.0)
Hatton et al. (1983)
137
Cs
7.5
198
12
ND
DeLaune et al. (1986)
Location
Method
Carr’s Island, Altamaha River, GA Barataria Basin, Mississippi River, LA Barataria Basin, Mississippi River, LA
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J. Patrick Megonigal and Scott C. Neubauer
Values are presented as means + 1SD, with reported ranges in parentheses. Studies that did not contain burial rates for at least two of the three elements (C, N, P) are not shown. ND, no data. a As needed, organic matter (OM) accumulation rates were converted to C assuming % C = 0.5 (% OM). b Ranges calculated from a mechanistic model of sediment decomposition. P burial determined from soil P content in Bowden (1984). c Cores (1 per site) were dated using both 137Cs and 210Pb. At one site, accretion rates from the two methods were identical. At the other site, this table reports the midpoint for the 137Cs (14 mm/ year) and 210Pb methods (9 mm/year), as was done by Church et al. (2006). d Seventeen sites for N burial. e Range in rates estimated since 1900 for high and low marsh cores.
541
Biogeochemistry of Tidal Freshwater Wetlands
export of DIC from freshwater and low-salinity tidal wetlands to estuaries and coastal oceans (Smith and Hollibaugh, 1993; Frankignoulle and Bourges, 1996; Cai and Wang, 1998; Frankignoulle et al., 1998; Nietch, 2000). Far less DOC is exported from tidal wetlands to estuaries compared to DIC, yet DOC is arguably the most important form of C exported from these systems. Saline tidal marshes generally export DOC to estuaries (Nixon, 1980 and references therein, Chapter 16) where it influences estuarine microbial metabolism, nutrient cycling, and ultraviolet (UV) light penetration of the water column (Epp et al., 2007). There is no evidence of DOC export from Sweet Hall Marsh (Neubauer, 2000), but DOC is exported from TFWs on the Hudson River, USA (Findlay et al., 1998) and the Patuxent River, USA (Figure 2). Raymond and Bauer (2001) proposed that TFWs are 30% of all DOC sources to the comparatively pristine York River, USA. The chemical composition of DOC influences the effects it will have on receiving estuaries. Features such as the aromatic ring content affect UV radiation adsorption (Tzortziou et al., 2007) and perhaps the ability of DOC to support microbial respiration. Chromophoric dissolved organic matter (CDOM) is the light-adsorbing component of DOC and a dominant fraction of the DOC pool in estuaries. The CDOM exported from a TFW on the Patuxent River, USA has a shallow spectral slope (SCDOM) (Figure 2), indicating the marsh is a source of relatively complex, high molecular weight, and aromatic-rich DOC. Presumably the SCDOM of this TFW marsh reflects the relatively high lignin content and complexity of organic C compounds in emergent wetland plants compared to phytoplankton (Enriquez et al., 1993; Tzortziou et al., 2008).
0.0185
2.0
Absorption (440 nm)
1.8
0.0181
1.6 0.0177 1.4 0.0173 1.2 0.0169
1.0
Flooding tide
Flooding tide
Absorption spectral slope (S)
slope (S)
440 nm absorption
0.0165
0.8 0
2
4
6
8
10 12 14 16 Elapsed time (h)
18
20
22
24
Figure 2 Tidal cycle variation in CDOM optical properties as a function of elapsed time since the first measurement (h) for the tidal cycle sampled at Jug Bay, USA on 7^8 September 2005. Left axis: CDOM absorption magnitude at 440 nm, aCDOM(440), Right axis: CDOM absorption spectral slope (S). The data were collected by M. Tzortziou, unpublished. Two horizontal lines mark the period of flooding tide when waters are dominated by estuarine CDOM sources.
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J. Patrick Megonigal and Scott C. Neubauer
There are relatively few studies of particulate organic carbon (POC) export from TFWs, but the process is expected to be dominated by seasonal and episodic events such as storms (Findlay et al., 1990). Findlay et al. (2005) estimated that TFWs of the Hudson River, USA, either import POC or have a near-zero POC balance. As with DIC and DOC, the contribution of terrestrial sources to POC is highest at the freshwater end of tidal rivers (Hoffman and Bronk, 2006). However, it is difficult to separate the POC contribution of TFWs from upland vegetation and upland soils on the basis of stable C and N isotopes, C:N ratios, or other biomarkers (Hoffman and Bronk, 2006). 2.2.2. Export of methane A portion of organic matter decomposition in saturated soils yields methane (CH4) rather than CO2. In a recent review of wetland CH4 emissions in the conterminous United States, fluxes from five TFW sites were 32 + 37 g CH4/m2/year (mean + SD; Bridgham et al., 2006). This compares to 36 + 47 g CH4/m2/year from all freshwater wetlands in the review and only 10 + 22 g CH4/m2/year from saltwater wetlands, apparently due to suppression by sulfate reduction. Although wetlands are 40% of all CH4 sources globally, the global contribution of TFWs to atmospheric CH4 is negligible because of their limited area (Bridgham et al., 2006). CH4 emissions from TFWs vary widely, as expected for any broad class of wetland ecosystems (van der Nat and Middelburg, 2000; Bridgham et al., 2006). Rates of CH4 emission underestimate the contribution of methanogens to overall microbial respiration by 30–90% because they do not account for microbial CH4 oxidation to CO2 (Megonigal and Schlesinger, 2002), export of CH4 in groundwater (Kelley et al., 1995), or ebullition (Chanton et al., 1989, van der Nat and Middelburg, 1998a).
3. PROCESSES GOVERNING ORGANIC C ARBON M ETABOLISM The fate of organic matter in wetland ecosystems is regulated by complex interactions between plant processes that affect electron donor and electron acceptor availability, and microbial processes that degrade and modify organic matter (Megonigal et al., 2004). Here we review the TFW literature on C cycling processes, focusing on plant–microbe interactions. A major initial constraint on C metabolism in wetlands is the chemical composition of plant material, a topic that was recently reviewed by Findlay et al. (2009) and is not considered here. Tissue quality interacts strongly with physiochemical factors such as O2 availability, nutrient availability, temperature, and pH (Day, 1982; Benner et al., 1985). Perhaps the most important of these is O2, which severely limits decomposition in saturated soils. Aerobic respiration yields more free energy than anaerobic respiration, and is less dependent on highly constrained chemical interactions among microorganisms. Yet, there are virtually no estimates of aerobic respiration in wetlands (but see Howes et al., 1984) because there are no methods for measuring in situ O2 flux that account for O2 introduced through root
Biogeochemistry of Tidal Freshwater Wetlands
543
O2 loss (ROL). (We prefer this term to radial O2 loss because it is more intuitive and, arguably, more precise). O2 diffusion across the soil surface is much easier to quantify and is usually limited to a few millimeters in saturated soils. However, O2 can penetrate much deeper if the water table drops for an extended period of time. An advantage of studying microbial processes in TFW soils, compared to nontidal soils, is that the depth of O2 penetration is relatively stable across seasons (Megonigal and Schlesinger, 2002). At aerobic-anaerobic interfaces, reduced forms of elements are oxidized by O2, regenerating oxidized forms that then support anaerobic respiration.
3.1. Anaerobic respiration Microorganisms derive energy by transferring electrons from an external electron donor to an external electron acceptor. Most respiration in wetland soils depends directly or indirectly (in the case of H2) on organic C as the electron donor. Generally, C limits microbial respiration because demand for electron donors exceeds the supply. This is true even in organic soils because most of the organic C is recalcitrant from the perspective of anaerobic respiration. Competition for electron donors among anaerobic microorganisms favors the respiration pathway that yields the greatest free energy in the order denitrification > manganese reduction > iron reduction > sulfate reduction > methanogenesis. There is some evidence that humic substances are common terminal electron acceptors, falling perhaps between iron and sulfate reduction in this sequence (Megonigal et al., 2004). Many TFWs occur in urbanized watersheds and are exposed to high NO 3 in floodwater. However, the relative contribution of the denitrification pathway to organic C mineralization, compared to other respiration pathways, has not been quantified in TFW soils to our knowledge. In tidal freshwater river sediments from the Altamaha River, USA, denitrification supported about 10% of anaerobic C mineralization (Weston et al., 2006). The NO 3 concentrations in this study were 20 mM, which is similar to NO concentrations in the Hudson River and many 3 other TFW systems. Although it seems unlikely that denitrification is ever a dominant pathway of microbial respiration in TFW soils, it can nonetheless be an important NO 3 sink in TFW-dominated estuaries (see Section 4.2). Iron oxide minerals can be the dominant electron acceptor in anaerobic mineral soils (Roden and Wetzel, 1996; Megonigal et al., 2004). The first studies to conclusively establish that Fe(III) reduction supports microbial growth were done in tidal freshwater river sediments (Lovley and Phillips, 1986, 1987). Over a decade later, van der Nat and Middelburg (1998a) suggested that Fe(III) reduction explained up to 80% of anaerobic respiration in mesocosms constructed with TFW soils and plants. The contribution of Fe(III) reduction to anaerobic metabolism appeared to be far higher in mesocosms planted with Scirpus lacustris than Phragmites australis, suggesting that species-specific plant characteristics influence Fe(III) reduction rates. A field study in a Peltandra virginica-dominated TFW showed that Fe(III) reduction mediated 20–98% of anaerobic C metabolism (Neubauer et al., 2005b, Figure 3). The importance of Fe(III) reduction declined throughout
544
J. Patrick Megonigal and Scott C. Neubauer
μmol C/g dry soil/day
(a)
5
10
15
(c) 100
Jul
10 cm
Jug Bay
20
% of total
Jack Bay Jug Bay
0
50 cm 10 cm
80 60 40 20
50 cm
0 Jun
Jack Bay Jug Bay
(b)
0
5
10
15
Aug
20 Aug
10 cm
Jul
Fe(III) reduction SO42– reduction Methanogenesis
10 cm
CO2 + CH4 production
Figure 3 (a) and (b): Soil organic C mineralization rates in a tidal freshwater marsh ( Jug Bay) and a brackish marsh ( Jack Bay) in July and August 2002. The July data (a) provide a comparison of rates at 10 and 50 cm depth. The August data (b) provide a comparison of the total rate of anaerobic C decomposition as determined by the sum of CO2 and CH4 production (striped bars) versus the sum of C mineralization from three possible anaerobic pathways (Fe(III) reduction, SO4 2 reduction, and CH4 production). (c) Seasonal changes in the relative importance of Fe(III) reduction (open squares), SO4 2 reduction (filled squares), and methanogenesis (filled triangles) to from June to August. Error bars show –SE, n = 3^5 replicate cores. From Neubauer et al. (2005b).
the growing season in parallel to plant activity, again suggesting that plants indirectly regulate this microbial process (Figure 3). In both cases, it is likely that plants influence soil iron cycling through ROL-driven regeneration of Fe(III) oxides in the rhizosphere. Indeed, the rhizosphere is a hot spot of iron cycling because it supports vigorous Fe(II) oxidation and Fe(III) reduction (Weiss et al., 2004, 2005; Neubauer et al., 2007, 2008). Manganese-respiration has received very little attention because concentrations of Mn(III, IV) are usually far lower than Fe(III) in soils (Neubauer et al., 2005b). In theory, this limitation could be overcome by differences in Fe and Mn chemistry, such as the fact that Mn(III, IV)-reduction is favored thermodynamically over Fe(III) reduction. Indeed, solid-state Au/Hg voltammetric microelectrode profiles in a TFW suggested that Mn(III, IV) reduction is more important than Fe(III) reduction in some locations (Ma et al., 2008). Humic substances are the most recent class of terminal electron acceptor identified in anaerobic substrates (Nevin and Lovley, 2000). It is difficult to directly measure microbial humic acid reduction. However, it has been observed that the amount of CO2 and CH4 produced in root-free, anaerobic soil incubations often far exceeds the summed contributions of denitrification, metal reduction, and sulfate reduction (Neubauer et al., 2005b; Keller and Bridgham, 2007). Microbial
Biogeochemistry of Tidal Freshwater Wetlands
545
reduction of humic substances may be supporting the unexplained CO2 respiration measured in these incubations. Sulfate reduction is not expected to be an important pathway for C mineralization in TFWs because of limitation by low SO2 4 availability at concentrations <1 mM, but it has rarely been measured in TFWs (Figure 3). However, sulfate reduction rates in TFWs will increase as SO2 4 concentrations rise due to sea-level rise and saltwater intrusion into tidal freshwater rivers (Weston et al., 2006). There is evidence that increased rates of sulfate reduction will stimulate organic matter mineralization (Portnoy and Giblin, 1997; Weston et al., 2006; Craft, 2007). However, other studies reported no differences in rates of anaerobic organic matter mineralization in comparisons of brackish versus tidal freshwater marsh soils (Neubauer et al., 2005b, Figure 3) and sediments (Kelley et al., 1990). This is an important question to understand with respect to climate change because of the implications for soil C pools and nutrient turnover in former TFWs (Neubauer and Craft, 2009).
3.2. Processes regulating methane production, oxidation, and emission Plants are the largest source of organic C that ultimately supports methanogenesis in soils, so factors that influence plant productivity indirectly regulate CH4 emissions. Several lines of evidence suggest that CH4 production is closely coupled to photosynthesis and NPP (Megonigal et al., 2004). In TFWs the evidence includes plant removal experiments (van der Nat and Middelburg, 1998a), a 14CO2 tracing experiment (Megonigal et al., 1999), and relationships between CH4 emissions and photosynthetic rates (Vann and Megonigal, 2003). Nonetheless, plants simultaneously reduce potential CH4 emissions by releasing O2 from roots into the soil that regenerates competing terminal electron acceptors such as Fe(III) and supports microbial CH4 oxidation. The net effect of plants on CH4 production, oxidation, and transport favors increased CH4 emissions. Perhaps the most definitive evidence of this are numerous observations of greater CH4 emissions from planted than plant-free soils (Kelley et al., 1995; van der Nat and Middelburg, 2000) and the strong positive correlation between NPP and CH4 emissions across North America (Whiting and Chanton, 1993). Methanogenesis is suppressed when there is an adequate supply of competing electron acceptors. In mineral-rich TFW soils, the dominant competing electron acceptor is Fe(III), which plants regenerate as poorly crystalline Fe(III) oxides in the rhizosphere (Weiss et al., 2004). Suppression of methanogenesis by Fe(III) reduction can range from complete to negligible depending on several factors, including plant activity (van der Nat and Middelburg, 1998a; Neubauer et al., 2005b). Sealevel rise should suppress CH4 production in TFWs as rising SO2 4 availability increases the activity of sulfate-reducing bacteria that suppress methanogenesis (Neubauer and Craft, 2009). CH4-oxidizing bacteria are abundant in microaerobic zones of the wetland plant rhizosphere where they have the potential to respond to variations in plant
546
J. Patrick Megonigal and Scott C. Neubauer
CH4 oxidation (mg CH4 /m2/day)
40
Y = –0.623 + 0.731X, r 2 = 0.962
30
20
Upper site Lower site
10
0 0
10 20 30 Gross CH4 emissions (mg CH4 /m2/day)
40
Figure 4 Relationship between rates of gross CH4 emission and CH4 oxidation measured over a 13-month period in two tidal freshwater wetlands. From Megonigal and Schlesinger (2002).
physiology and morphology that influence ROL (van der Nat and Middelburg, 1998a,b). CH4 oxidation activity in TFW soils can be O2-limited (van der Nat and Middelburg, 1998b) or CH4-limited (Figure 4; Megonigal and Schlesinger, 2002).
4. N ITROGEN BIOGEOCHEMISTRY Historically, many tidal freshwater marsh nutrient studies have focused on understanding if and how marshes affect estuarine water quality, and were designed to quantify exchanges of dissolved inorganic nitrogen (DIN) between marshes and tidal waters (collectively, “flux studies,” e.g., Grant and Patrick, 1970; Heinle and Flemer, 1976; Simpson et al., 1978; Bowden, 1986; Chambers, 1992; Campana, 1998; Ziegler et al., 1999). This type of study provides valuable information, but is also limited by high spatio-temporal variability and difficulties in obtaining accurate hydrologic budgets to scale up the measurements. Flux studies do not provide detailed information on internal transformations that are occurring within marsh soils and sediments. Numerical simulation models (Morris and Bowden, 1986) allow process rates (e.g., organic matter production and mineralization) to be calculated from measurements of soil organic and inorganic nutrients. If robustly designed, these models can be used to explore how the system might respond to future environmental changes (e.g., level of watershed nutrient loading). Recently, isotope tracers have been used to determine both the fate of water column nitrogen
547
Biogeochemistry of Tidal Freshwater Wetlands
(N) and the processes by which the N is transformed or removed from the water column (Gribsholt et al., 2005, 2006, 2007). This approach provides an elegant means of quantifying N transformations that eliminates many of the issues associated with flux studies and process rate measurements. Isotope tracer studies can be used to quantify fluxes and transformations in the water column (Tobias et al., 2003; Gribsholt et al., 2005, 2006, 2007) and processes occurring throughout the root zone across a range of temporal scales (White and Howes, 1994; Tobias et al., 2001b). We only know of two comprehensive N models for tidal freshwater marshes that consider exchanges of N between the marsh, estuary, and atmosphere, as well as internal N transformations in soils (Figure 5a,b; Bowden et al., 1991; Neubauer (a) North River, MA (USA) Atmosphere
0?
N2 + N2O
7.4–16
0.4
0.6
0.6
Marsh
5.4
Live plants 0.3
16
16
–
Fresh litter
NO3
2.5 7.8
NH4+
20
Peat
9.1
9.1–0
17
41
5.9
Marsh 11
River
2.1
0–9.1
(units are g N/m2 marsh/year)
(b) Sweet Hall Marsh, Pamunkey River, VA (USA) 0.2
Water (high tide)
4.1
0.3
AGB
6.1
27 Benthic microalgae
Water (low tide)
BGB 14
NO3– + NO2–
25
72
1.7
4.7
Porewater NO3– + NO2–
8.2
13 Particulate N
16
36
NH4+
Porewater NH4+
1.2
142
Sediment 18 (units are g N/m2 marsh/year)
37
79
27
48
Sediment particulate N
Figure 5 Ecosystem-scale tidal freshwater marsh N models. (a) Process-oriented N budget for a 15.8 ha marsh based on field, laboratory, and modeling efforts. Figure is redrawn from Bowden et al. (1991). (b) Process-based N mass balance model for a 401 ha marsh based on measured field fluxes, with unmeasured fluxes calculated from literature values or to force the model to steady-state conditions. Figure is reproduced, with slight modifications, from Neubauer et al. (2005a). AGB = aboveground biomass; BGB = belowground biomass. Fluxes in each figure are g N/m2/year.
548
J. Patrick Megonigal and Scott C. Neubauer
et al., 2005a). Both models were built with seasonal data from different locations, but neither is seasonally or spatially explicit. The model by Bowden and colleagues (Figure 5a) describes N cycling at a North River, Massachusetts, USA, marsh that had organic-rich soils (40–63% organic matter) and a well-developed, persistent plant litter layer. Neubauer and collaborators (2002) studied Sweet Hall marsh on the Pamunkey River, USA, which had relatively little plant litter and more mineral soils (16–21% organic). Average nutrient concentrations (NHþ 4 and NO3 ) in the North River were 4 times greater than those in the Pamunkey River (Neubauer et al., 2005a). Despite the limitations of mass balance modeling and differences between these marshes in plant community type, soil type, marsh elevation, nutrient loading, and climate, several features of the N cycle were similar and may be common to TFWs generally: 1. Exchanges of NHþ 4 and NO3 between the TFW and tidal waters were small compared to rates of internal N cycling and transformations in soils. N in sediments deposited on the soil surface can be a significant source of new N to marshes, although the importance of this N source varies with flooding frequency and a suite of factors that influence sediment deposition rates (Darke and Megonigal, 2003). 2. Marsh–estuary exchanges of NO 3 are generally directed into the marsh (i.e., net uptake by the marsh) and are similar in magnitude to rates of denitrification, suggesting the two processes are coupled. 3. TFWs are efficient at recycling and retaining nutrients within the soil profile. The efficiency of nutrient recycling may be greater in older wetlands with deep soils than in younger wetlands (Morris and Bowden, 1986). 4. The generation of inorganic N via organic matter mineralization can provide more than enough N to support primary production. This suggests that plant production may be largely uncoupled from nutrient loading in the adjacent tidal waters over relatively short periods of time (ca. one to several years). Over longer periods of time, the progressive assimilation and accumulation of water column N by the TFW offsets N losses to denitrification and helps build the soil N pool, which can then be mineralized to support plant demands.
4.1. Nitrogen exchanges Exchanges of NHþ 4 between TFWs and floodwaters are controlled by the diffusive gradient between soil pore waters and tidal waters, which is influenced strongly by microbial NHþ 4 assimilation. Thus, TFWs with an extensive litter layer often show net NHþ uptake from the water column (Heinle and Flemer, 1976; Bowden, 4 1986) despite high pore water concentrations because the litter layer is acting as a sink for both pore water and water column NHþ 4 . Other wetlands are sources of NHþ to tidal waters (Campana, 1998; Ziegler et al., 1999; Neubauer et al., 2005a). 4 Although wetland–estuary exchanges of NHþ (and other N forms) may be sig4 nificant on a whole-estuary basis, the magnitude of these fluxes is generally small relative to N transformations occurring within soils (e.g., Figure 5a,b). In a pair of elegant 15NHþ 4 labeling experiments conducted in May (early growing season with
549
Biogeochemistry of Tidal Freshwater Wetlands
active plant growth) and September (late growing season with senescent/flowering plants), the fate of water column NHþ 4 was tracked over 15 days in a N-rich tidal freshwater marsh in Belgium, EU (Gribsholt et al., 2005, 2006, 2007). Despite the temporal separation between the experiments, the fates of NHþ 4 were remarkably similar between the months. In each experiment, the majority of the water column NHþ 4 was exported from the system without being transformed by the marsh (Figure 6). Approximately 4% of the NHþ 4 was sequestered by the marsh and ended up in plant biomass, litter, or the marsh soil (either via physical sorption or microbial assimilation). Overall, microbial pathways of N uptake were more important than the direct assimilation of tidal water NHþ 4 by plants (Gribsholt et al., 2006). However, plants are likely to play indirect roles in modifying water column N loads by providing both O2 and labile organic C to soil microbes. The 15 N label also was found in other N pools within the water column, indicating that active N transformations were occurring within the water column and/or in flooded marsh soils. Of these N transformations, nitrification accounted for the largest fraction of the added NHþ 4 , with smaller amounts in the suspended particulate N, N2, and N2O pools. The marsh soils appeared to be a significant site for nitrification in May (Gribsholt et al., 2005), whereas soil denitrification rates were highest in September (Gribsholt et al., 2006). TFWs are often significant sinks for water column particulate N deposited on the soil during tidal flooding. The N content of accumulated sediments ranged from 4 to 16 mg N/g sediment in several tidal freshwater marshes in Virginia (Morse et al., 2004; Neubauer et al., 2005a) and is significantly correlated with the soil N content (Morse et al., 2004). The sediment-associated N is presumably a combination of detrital material, microbial biomass, and NHþ 4 sorbed to mineral surfaces. On an annual Tielrode marsh, Scheldt River (Belgium) May 2002
September 2003
NO3– + NO2–
SPN
N2 + N2O
NO3– + NO2–
SPN
N2 + N2O
8.7
0.5
0.02
7.7
0.7
0.51
N transformations Export
69
(Unknown)
N transformations Export
15
17
NH4+
Marsh N storage 1.6 Sediment
15
NH4+
36
0.4
1.7
Litter
51
(Unknown)
Marsh N storage 0.7
0.3
Sediment
0.3
1.8
Litter
1.1
(units are % of total 15N added)
Figure 6 Transformations and uptake of NH4þ, as inferred by whole-ecosystem 15NH4þ labeling of a 3,477 m2 marsh area (Gribsholt et al., 2006). Fluxes indicate the percent of the added 15NH4þ label that was recovered in each N pool after label addition. The total input of 15N was 1.97 and 1.41 mol 15N-NH4þ in May and September, respectively; the label addition increased the total tidal water NH4þ concentration by 14% (May) to 73% (September). SPN, suspended particulate N.
550
J. Patrick Megonigal and Scott C. Neubauer
basis, inputs of allochthonous particulate N can be large with respect to the marsh N budget, contributing up to 20 g N/m2/year (Bowden et al., 1991; Morse et al., 2004; Neubauer et al., 2005a). There is significant spatial variation in deposition rates driven by marsh elevation and flooding frequency (Morse et al., 2004). Over decadal scales, the burial of N sequesters significant amounts of N in tidal freshwater marsh soils, on the order of 10–30 g N/m2/year (Table 1 and references therein). Deposition and burial rates are much higher in Sweet Hall marsh than the North River marsh (cf. Figure 5a,b), a difference reflected in the lower organic content of Sweet Hall versus North River marsh soils. Indeed, regional patterns of sedimentation may explain why soil accretion rates in TFWs of the Northeast United States are correlated only with organic accumulation, while those in the southeastern United States are correlated with both mineral and organic accumulation (Neubauer, 2008). Over each tidal cycle, a large volume of water floods and ebbs from the surface of TFWs. Because of the high rate of surface water exchange, wetland uptake of DIN from flood waters is inefficient (Hopkinson, 1992) and can meet only a small fraction of plant N demand. This leads to the “requirement” that existing nutrients are retained within the wetland. Indeed, only 1% of the total N supplied to the NHþ 4 pool from external and internal sources is lost; the remainder is recycled to other N reservoirs in the marsh (Figure 5a,b). It is likely that the slow turnover of marsh pore water (67 to >800 days to 30 cm at Sweet Hall marsh) drives this efficient N retention. Microbial immobilization of NHþ 4 into particulate matter is a primary mechanism by which N is retained in the marsh; this mechanism can retain 50% of mineralized N in both tidal freshwater (Bowden et al., 1991; Neubauer et al., 2005a) and salt marshes (Anderson et al., 1997). In contrast to the pore water NHþ 4 pool, plant biomass N in TFWs is directly exposed to the tides and less efficiently retained, with 50% exported as dissolved or particulate organic matter (Hopkinson, 1992; Neubauer et al., 2005a). In support of the link between water turnover and system closure, there was little evidence of N export from a “periodically flooded” high marsh (Bowden et al., 1991).
4.2. Nitrogen transformations In addition to burial, tidal freshwater marshes can also permanently remove DIN from riverine and estuarine waters via denitrification. Sources of NO 3 for denitrification include nitrification within the marsh and the uptake of external (water column) NO 3 . Mass balance calculations indicate high rates of N removal in upper estuaries (Howarth et al., 1996) and many characteristics of TFWs appear to favor denitrification (e.g., high active surface area, shallow depth to anaerobic zone, high organic matter availability). However, few studies have directly measured denitrification in TFWs (Groszkowski, 1995; Merrill, 1999; Merrill and Cornwell, 2000). Greene (2005) reported that median denitrification rates for a tidal freshwater marsh (120 mmol N/m2/h) were slightly larger than the median rate for a wide range of intertidal and aquatic systems (75 mmol N/m2/h). There is considerable spatial variability between and within TFWs (Merrill, 1999; Greene, 2005). The environmental controls on denitrification have been extensively reviewed by several authors (Seitzinger, 1988; Cornwell et al., 1999; Wallenstein et al., 2006) and
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will not be covered in great detail here. In tidal freshwater marshes, denitrification rates were correlated with benthic sediment O2 demand in a New York marsh, but not in a TFW in Maryland, USA (Merrill, 1999). Based on laboratory manipulations, denitrification rates increase with increases in water column NO 3 (Merrill, 1999; Greene, 2005). Similarly, the N models of Bowden et al. (1991) and Neubauer et al. (2005a) suggest that denitrification is supported primarily by water column NO 3. Because ROL and diffusion of O2 across the soil surface can support oxidation in tidal freshwater marsh soils (Neubauer et al., 2005b), denitrification is likely to be coupled to both in situ nitrification and water column NO 3 uptake. Gribsholt et al. (2005, 2006, 2007) presented evidence for coupled nitrification–denitrification in TFWs; 15 following the addition of a 15 NHþ 4 label to tidal flood waters, some of the N label appeared in the dissolved N2 and N2O pools (Figure 6). Much of this nitrification takes place in marsh soils (possibly associated with plant roots) rather than in the water column (Gribsholt et al., 2005) although the importance of the soil as a site for nitrification can vary seasonally (Neubauer et al., 2005a; Gribsholt et al., 2006). When integrated over the entire network of TFWs within an estuary, nutrient removal may be substantial since small contributions by individual marshes can have a large cumulative impact on water quality. This is especially true in systems with large areas of tidal marsh (relative to open water). For example, in the Patuxent and Choptank rivers, USA, slightly more than 30% of the total N inputs at the fall line are permanently removed by low-salinity tidal marshes via burial and denitrification (Merrill, 1999; Malone et al., 2003). In contrast, N removal by tidal freshwater marshes in larger systems such as the Hudson and Delaware rivers, USA, is less efficient, with only 2–5% of the N sequestered or denitrified (Academy, 1998; Merrill, 1999). þ Dissimilatory NO 3 reduction to NH4 (DNRA) is a mechanism by which NO3 can be retained in the marsh rather than lost to the atmosphere via denitrification. There is very little evidence from TFWs about the relative importance of DNRA and denitrification as fates for soil NO 3 . Bowden (1986) determined that DNRA rates were 5% of NO supply (i.e., nitrification) rates. In contrast, Neu3 bauer et al. (2005a) calculated that DNRA was about 40% of nitrification. Based on work in other systems, the availability of labile C relative to NO 3 (i.e., electron donor:electron acceptor ratio) is important in determining the fate of NO 3 , with DNRA dominating with high organic C availability (Fazzolari et al., 1998; Christensen et al., 2000) and denitrification increasing in importance at higher NO 3 concentrations (Nijburg et al., 1997; Tobias et al., 2001a,b). Thus, lower NO 3 in the Pamunkey River (Neubauer et al., 2005a) may explain the higher importance of DNRA in that system. Across the estuarine gradient, DNRA is generally more important (relative to denitrification) in estuarine and marine systems whereas denitrification increases in importance in freshwater systems (Tobias et al., 2001b), a pattern that may be related to sulfide inhibition of denitrification (Brunet and Garcia-Gil, 1996; An and Gardner, 2002).
4.3. Nutrient regulation of plant production There is evidence to suggest that rates of plant production in tidal freshwater marshes are largely uncoupled from allochthonous nutrient inputs. First, rates of
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N mineralization in TFWs are considerably greater than plant N demand and rates of diffusive DIN uptake from tidal waters (e.g., Figures 5a and 5b; Bowden et al., 1991; Neubauer et al., 2005a). In these studies, gross N mineralization provided almost 3 times more NHþ 4 than was needed to support annual plant N require15 ments. Second, 15 NHþ 4 label tracing showed that less than 2% of the applied N was incorporated into plant leaves and roots after 15 days (Figure 6), suggesting that the plants were not dependent on water column N over short timescales (Gribsholt et al., 2005, 2006, 2007). Third, there are many examples in which direct fertilization of TFWs with N, P, or N þ P generally did not increase in either aboveground biomass or biomass nutrient content (Whigham and Simpson, 1978; Walker, 1981; Booth, 1989; Chambers and Fourqurean, 1991; Morse et al., 2004). However, in other cases fertilization did increase plant growth (DeLaune et al., 1986; Booth, 1989; Frost et al., 2009), so it has proven difficult to unambiguously determine the nature (or existence) of nutrient limitations in TFWs (Chambers and Fourqurean, 1991). Phosphorus (P) is a limiting nutrient in many freshwater ecosystems because inorganic P precipitates with Fe, Al, Ca, and Mg minerals, and organic P is generally unavailable to plants. In contrast, marine systems are often N limited. This generality has not been adequately tested in TFWs, but P additions did not affect primary productivity in TFWs in Virginia and Georgia, USA (Morse et al., 2004; Frost et al., 2009).
5. P HOSPHORUS BIOGEOCHEMISTRY Because TFWs are located in upper estuaries where watershed-derived inputs of P are concentrated, these intertidal systems may be key sites in landscapes for P sequestration and transformation. In a pair of tidal freshwater marshes, sediments deposited on the soil surface contained 0.3–1.7 mg P/g sediment and contributed inputs of 0.6–2.3 g P/m2/year (Morse et al., 2004). Uptake of inorganic P by organisms results in a relative enrichment of organic P in surface soils (Morse et al., 2004). In a South Carolina TFW, much of the soil organic P was bound to humic acids whereas the inorganic P was primarily associated with Fe or Al (Paludan and Morris, 1999). Phosphatase (the enzyme that liberates organic-bound P) is secreted by plants, algae, and bacteria under conditions of PO3 4 limitation. Phosphatase activity is expected to be greatest where the biological demand for inorganic P is high, most soil P is in organic forms, and soil sorption limits pore water PO3 4 concentrations. The activities of three phosphatase enzymes were highly correlated with aboveground plant biomass and soil organic content in a successional sequence of TFWs (Huang and Morris, 2003). Similar positive correlations between phosphatase activity, soil organic matter, and soil organic P were observed along a salinity gradient, with the highest activity in TFWs (Huang and Morris, 2005). Furthermore, organic P made up a greater fraction of total P in the late (intertidal marsh) versus early (open water) successional stages. The interactions between P
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availability and demand may lead to a positive feedback that drives ecosystem succession, whereby high demand for inorganic P reduces available PO3 4 concentrations, leading to an increase in phosphatase activity and increased organic P mineralization, further increasing plant growth (Huang and Morris, 2003). Over years to decades, P mineralization results in a significant decrease in organic P concentrations with increasing soil depth (Paludan and Morris, 1999). Despite high phosphatase enzyme activities in TFW soils, concentrations of dissolved PO3 4 are often low due to the combined effects of biological demand and chemical sorption processes that remove free PO3 from marsh pore waters. 4 Sundareshwar and Morris (1999) showed that P sorption is higher in more freshwater systems because sediments tend to have higher surface areas and a lower content of Fe and Al minerals. The lack of significant accumulation of inorganic P in deeper soils (Paludan and Morris, 1999) implies that organic P is not simply mineralized to inorganic P but is instead removed from the soil via plant uptake or hydrological export. Although TFWs are sinks for sediment-associated particulate P, PO3 4 fluxes are highly variable and there can be net PO3 4 uptake (Simpson et al., 1978; Gilbert, 1990), seasonal variability (Simpson et al., 1983; Campana, 1998), or negligible PO3 4 fluxes (Anderson et al., 1998) between TFWs and tidal waters. One factor that may affect spatial and temporal variations in marsh-estuary tidal fluxes of PO3 4 is the interplay between P, Fe, and O2 dynamics. In soils that are regularly exposed to O2 during low tides (e.g., the marsh surface and creek bank edges), Fe(II) can oxidize to Fe(III) and lead to the formation of an “iron curtain” of iron oxyhydroxide minerals that efficiently sorb PO3 4 , causing P retention in the marsh (Chambers and Odum, 1990). In combination with diagenetic effects (discussed above), this mechanism may contribute to decreases in total soil P content with increasing depth that have been observed in some TFWs (Bowden, 1984; Chambers and Odum, 1990; Merrill, 1999; Paludan and Morris, 1999), although not in others (Simpson et al., 1983; Greiner and Hershner, 1998). The mineral (i.e., Fe) content of the soils is likely to play a role in the efficiency of such an iron curtain. The ecological implications of the iron curtain on ecosystem P dynamics are unclear because it retains a potentially limiting nutrient within the marsh, but not necessarily in a bioavailable form. Another implication is that high marshes, which are flooded less frequently and therefore exposed to air for longer periods of time, may have an extensive iron curtain that allows for more efficient P retention and recycling than low marshes, leading to increased P accumulation in high versus low marsh habitats (Khan and Brush, 1994). The storage (burial) of P in marsh soils is an important mechanism by which P can be removed and sequestered from estuarine waters. Over time scales ranging from decades to centuries, TFWs sequester significant amounts of P, with burial rates ranging from <0.5 to >4 g P/m2/year (Table 1 and references therein). These burial rates are roughly comparable to rates of sediment-associated P deposition onto the marsh surface (Morse et al., 2004). Extrapolating marsh P burial rates to a landscape scale shows that low salinity tidal marshes can effectively sequester a significant fraction (>60%) of watershed derived P in relatively small, marshdominated estuarine systems (e.g., Patuxent and Choptank Rivers, MD, Merrill,
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1999; Malone et al., 2003). In contrast, about 12% of the combined sewage and riverine-derived P to the upper Hudson River, USA, is buried in freshwater tidal and nontidal marshes (inputs from Limburg et al., 1986; Phillips and Hanchar, 1996; burial rate from Merrill, 1999). Based on average literature values for marsh P burial, only 7% of the P entering the upper Delaware River estuary, USA, is permanently buried in tidal marshes (Academy, 1998). Thus, TFWs can be longterm sinks for significant amounts of watershed-derived P, but the extent of wetlands within the estuary (relative to the size of the estuary or watershed) appears to be important in determining how efficiently TFWs perform this function.
6. S ILICON B IOGEOCHEMISTRY The weathering of terrestrial silicate (Si) minerals ultimately leads to inputs of dissolved silica (DSi) to estuaries and the coastal ocean where diatom production can be limited by low DSi availability. Recent evidence suggests that silica transformations in TFWs play a key role in transforming silica from biogenic (BSi) to DSi forms. For example, under low discharge conditions (i.e., summer) in the Scheldt Estuary, EU, the input of DSi from fluvial sources was as low as 10,000 kg/ month, an amount that can be exported from the 450 ha of TFWs in the system in six tidal cycles (Struyf et al., 2006). Because rates of DSi export increased with decreasing concentrations of DSi in estuarine waters, marsh-mediated recycling of Si may be especially important when low ambient DSi concentrations otherwise limit aquatic primary productivity (Struyf et al., 2006). Understanding the factors that regulate DSi export from TFWs requires additional research on biogeochemical transformations in TFW soils (Struyf et al., 2005a,b, 2007).
7. BIOGEOCHEMICAL E FFECTS OF SEA -LEVEL R ISE The increase in global ocean levels has the potential to affect TFWs by modifying their hydroperiod and by pushing the salt front up-estuary so that these systems are exposed to more saline waters. In general, the biogeochemical effects of changes in flooding frequency will be mediated by changes in plant productivity (Odum et al., 1995; Morris et al., 2002), decomposition (Odum and Heywood, 1978; Chambers and Odum, 1990), and deposition of allochthonous minerals and nutrients via sedimentation (Khan and Brush, 1994; Pasternack and Brush, 1998; Darke and Megonigal, 2003). Future changes in salinity in a given TFW will depend on the wetland’s position on the current salinity gradient, rates of sea-level rise, and changes in river discharge. Both sea-level rise and river discharge are influenced by global warming (Burkett et al., 2001). Changes in salinity can affect a number of nutrient cycling processes including N and P sorption, denitrification, and nitrification (Howarth et al., 1988; Caraco et al., 1989; Rysgaard et al., 1999). The intrusion of salt water and associated SO2 4 can lead to the breakdown of the “iron curtain” of Chambers and Odum (1990).
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As Fe oxides are reduced either biologically (by Fe(III)-reducing bacteria) or chemically (via H2S), sediment-bound P is released and the Fe(II) can be sequestered in Fe–S compounds (e.g., pyrite) (Caraco et al., 1989; Lamers et al., 2001). þ Elevated salinity can also lead to decreased sorption of NHþ 4 to soil particles as NH4 þ 2þ is displaced by positively charged cations such as Na and Mg . Increased NHþ 4 concentrations may suppress N2 fixation because nitrogenase is inhibited by free NHþ 4 (Howarth et al., 1988). The physiological effects of salinity on nitrifying and denitrifying microbes reduce the activity of these organisms (MacFarlane and Hebert, 1984; Furumai et al., 1988; Stehr et al., 1995, Rysgaard et al., 1999). If increased SO2 4 concentrations accelerate soil organic matter decomposition (see Section 3.1), rates of nutrient mineralization would also increase. Salt water intrusion will affect greenhouse gas emissions. Thermodynamics dictate that microbial SO2 4 reducers will outcompete methanogens for substrates such as organic C (Megonigal et al., 2004). As a result, salt water intrusion should increase SO2 4 reduction at the expense of CH4 production. Salt water intrusion could also affect the production of N2O if high H2S concentrations inhibit both nitrification and denitrification, as has been shown for unvegetated sediments (Joye and Hollibaugh, 1995; Brunet and Garcia-Gil, 1996; An and Gardner, 2002). Much of the research on the effects of rising salinity on soil and sediment biogeochemistry has focused on transient effects. Over longer time periods, saltsensitive plants, animals, and microbes will likely be replaced by salt-tolerant species (Magalha˜es et al., 2005). There is relatively little known from direct manipulations of salinity about the direction of these longer term effects.
8. C ONCLUDING C OMMENTS It is perhaps appropriate that TFW biogeochemistry has not been well studied given the fact TFWs occupy less area than many other wetland ecosystems. However, TFWs are grossly underrepresented even in the tidal wetland literature. A Web of Science search (1980–2008) shows that <5% of all tidal wetland literature concerns TFWs, whereas TFWs can represent 20% of the total tidal wetland area in a region (Stevenson et al., 1988; Dahl, 1999). There are reasons why TFWs deserve increased attention. They are species-rich ecosystems that support waterfowl, fish, and terrestrial wildlife. They influence the chemistry of adjacent estuarine waters through exchange and transformation of organic C and nutrients. Because of their location at the head of estuaries, TFWs are important sites for the deposition of nutrient-laden sediments, and they can rapidly sequester C, N, and P through burial. In short, TFWs are important features of the landscapes in which they occur. TFWs are sentinel ecosystems for monitoring the influence of global climate change on coastal ecosystems. Poised at the interface of nontidal rivers and saline estuarine waters, they are influenced by river discharge and sea-level rise. River discharge is sensitive to precipitation and evapotranspiration, which are in turn sensitive to global warming (Palmer et al., 2008). Increasingly frequent incursions
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of saline water into TFW ecosystems can be expected as sea levels rise and droughts become more common. It is uncertain what the long-term effects of such episodic events will have on element cycling and plant community composition in TFWs, but based on current distributions of plant species, even small increases in salinity will elicit dramatic changes in plant and microbial community composition, and fundamentally alter the characteristics of TFW biogeochemical cycles (Neubauer and Craft, 2009). At present, there are many hypotheses about the response of TFWs to these perturbations. For example, we could expect TFWs to be less sensitive to sea-level rise than tidal saline wetlands because they are located at the head of estuaries near riverine sediment sources. However, their location also presents barriers for TFW transgression inland due to steep upland slopes. We do not understand TFWs well enough to predict how they will respond to climate and land use change. Tidal freshwater wetlands have not received the level of biogeochemical scrutiny that has been directed toward nontidal freshwater and tidal saline wetlands. In the absence of more complete knowledge, we often assume that TFW processes adhere to generalizations drawn from better-studied ecosystems. This approach has proved fruitful, but limited. For example, there is now doubt about the assumption that TFW plants are more rapidly decomposed than saline tidal wetland plants (Craft, 2007). Based on hydrology, it seems reasonable to assume that TFWs have relatively open nutrient cycles, but that was not the case in two TFW systems that have been fully studied. Clearly, understanding biogeochemical processes in TFWs will require more direct observations of TFWs in relation to their tidal saline and nontidal analogs.
ACKNOWLEDGMENTS We thank Stuart Findlay, Melanie Vile, Eric Boschker, and Mark Brinson for improving the manuscript with their insightful reviews. Nicholas Mudd reviewed the manuscript for typographic errors and citations, which we appreciate. This is contribution #1483 from the University of South Carolina’s Belle W. Baruch Institute for Marine and Coastal Sciences.
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G EOMORPHOLOGY AND S EDIMENTOLOGY OF M ANGROVES Joanna C. Ellison
Contents 1. Introduction 2. Mangrove Environmental Settings 2.1. Terrigenous settings 2.2. Islands 2.3. Inland mangroves 3. Tidal Range and Sea-Level Control 4. Sedimentation in Mangroves 4.1. Excessive sedimentation 5. Mangroves as Sea-Level Indicators 6. Storms/Tsunamis 7. Inundation Changes Affecting Mangroves 8. Conclusions Acknowledgment References
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1. INTRODUCTION Mangrove forests occur most extensively on low-energy, sedimentary shorelines of the tropics, generally between intertidal elevations. Their global extents are limited by the 20C sea surface temperature isotherm; hence, coastal distributions in the subtropics are influenced by the origin of influencing oceanic currents (Figure 1), thus reaching higher latitudes on east coasts of continents relative to west. Within these temperature-controlled global ranges, the extent of mangrove habitats is mostly controlled by availability of suitable environmental settings, and biodiversity therein is mostly controlled by biogeographic history (Duke et al., 1998).
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Figure 1 World distributions of mangroves (bold coastlines) and Holocene sea-level curve types. Sources: Pirazzoli (1991), Duke et al. (1998).
2. MANGROVE ENVIRONMENTAL SETTINGS The major structural elements of mangrove forests are woody tree species that have evolved special physiological and morphological adaptations to grow in the intertidal region of the marine environment (Lugo and Snedaker, 1974; Chapman, 1976). Hence, while mangroves are by definition a biogenic community primarily of trees and associated fauna, they do however have strong geomorphological characteristics. This is because mangroves are facilitated in their development by sheltered geomorphic settings, and the vegetation itself influences sedimentation. Old-growth mangroves have been found to occur in areas with suitable stability in geomorphology and within-site environmental conditions (Lugo, 1997). The geomorphology and sedimentology of mangroves have been well reviewed in the past by several authors, including Thom (1982, 1984), Woodroffe (1992), and Augustinus (1995). This review expands on areas already covered in these earlier reviews.
2.1. Terrigenous settings Thom (1982, 1984) adapted the delta classifications of Wright et al. (1974) to describe five mangrove geomorphic settings on coastlines dominated by terrigenous sediment deposition. The attributes of each type are specified in Table 1.
Attributes
Type River dominated
Tide dominated
Wave dominated
River and wave dominated
Drowned bedrock valley
Geomorphologic setting
Deltaic distributaries
Estuarine with elongated islands
Distributaries and lagoons
Open estuary
Sediment Tidal range Mangrove locations
Allochthonous Low Seaward edge and distributaries
Allochthonous High Tidal creeks and islands
Barrier islands/ spits and lagoons Autochthonous Any Inside lagoons
Limited Any Tributary river mouths
Dominant process
Freshwater discharge
Tidal currents
Wave energy
Examples
Mississippi, Orinoco
Ord (WA), Fly (PNG)
El Salvador, New South Wales
Allochthonous Any Abandoned distributaries and lagoons Wave energy and river discharge Grijalva, Burdekin Qld
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Table 1 Attributes of mangrove settings on coastlines dominated by terrigenous sediment supply
Sea level Sydney harbor
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Terrigenous sediment supply dominance at each of these five settings necessitates that each is associated with river mouths, the geomorphology of which and the associated mangrove habitats is controlled by the different processes dominant. In such environments, the mangrove distributions are an excellent geo-indicator of status and change in coastal conditions of progradation and erosion (Saintilan, 2004; Souza Filho et al., 2006). Some examples are examined below from the mangrove center of biodiversity of New Guinea and Northern Australia. Southern Papua New Guinea (PNG) has the greatest diversity of mangroves in the world, owing to its location at the center of the Indo-Malayan mangrove center of diversity (Duke, 1992). 2.1.1. The Fly River, Papua New Guinea The Fly River Delta is dominated by tidal currents even though located in a microtidal region, partly due to the strength of tidal currents offshore through the restricted Torres Strait. Mangrove forests cover 87,400 ha on estuarine islands, with distribution of mangrove communities controlled by salinity (Robertson et al., 1991). Of the mangrove area, 44% is Nypa fruticans forest, in areas with high-tide salinities of 1–10. Rhizophora apiculata–Bruguiera parviflora forest covers 36% in sections where water salinities >10. Subdominant species in this community include Bruguiera gymnorrhiza, Avicennia marina, Ceriops decandra, and Heritiera littoralis. Subdominant emergent mangrove species present in this community include Xylocarpus granatum and H. littoralis. Sonneratia lanceolata occurs on accreting banks in low-salinity areas and A. marina on accreting banks in higher water salinities of >5. 2.1.2. Purari River, Papua New Guinea The Purari Delta is only a few hundred kilometres east of the Fly, but owing to its high discharge through restricted channels and overall consequent low salinity it is river-dominated. Mangrove distributions differ, with most extensive mangrove stands dominated by Bruguiera species with some R. apiculata in salinities of <10, and N. fruticans in salinities of >10 (Cragg, 1983). Fringing mangroves of accreting banks are again dominated by S. lanceolata. The Purari carries a large sediment load, delivering 88.6 million m3/year of sediment to the head of the delta, of which 6% is sand, 48% is silt, and 46% is clay (Pickup and Chewings, 1983). Bore holes indicate deposition of 1,700 m of terrigenous sediment over the last 5 million years (Thom and Wright, 1983), creating a depositional plain between 30 and 60 km wide. This sediment-rich outflow limits the Northern Great Barrier Reef from extending into the Gulf of Papua. The large open estuaries of the Pie and Kikori rivers between the Fly and the Purari have less river discharge, hence are more saline. There, the most extensive mangrove stands are dominated by Rhizophora stylosa in salinities of around 30 parts per thousand, and fringing accretion banks are colonized by A. marina var.eucalyptifolia and Sonneratia alba. Lower salinity communities are similar to those in the Purari.
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2.1.3. Queensland and Northern Territory, Australia Along with southern New Guinea, northeastern Australia has the greatest concentration of mangrove species in the world (Duke, 1992). Owing to lower rainfall, Australian rivers generally have a lower discharge and sediment yield relative to New Guinea, and far more is known about their geomorphic evolution. In spite of this lower sediment yield, Holocene mangrove stratigraphic records in tropical Australia show that fluvial sedimentation of terrestrially derived material has caused seaward relocation of mangrove swamps since sea level stabilized around 6,000 years ago. At Princess Charlotte Bay in north Queensland, Chappell and Grindrod (1984) and Grindrod (1988) found shoreline progradation of fluvial/deltaic sediment at rates up to 2.7 m/a, determining the location and development of mangroves. Crowley et al. (1990) and Crowley and Gagan (1995) reconstructed the Holocene vegetation history of the Mulgrave/Russell estuary, south of Cairns. This showed that Rhizophora-dominated mangrove swamps were widely established by 6,800 BP (years before present). Subsequently, the estuary has infilled with sediments, leading to replacement by Bruguiera–Ceriops forest, and eventually freshwater swamp forest dominated by Melaleuca and Pandanus. At Missionary Bay, Queensland, Grindrod and Rhodes (1984) demonstrated mangrove response to mid to late Holocene sea-level change, as shown in Figure 1 to have included a mid-Holocene highstand (Chappell et al., 1983). This demonstrates that during earlier more rapid sea-level rise, the mangrove fringe was about a hundred meters wide, but with sea-level stabilization then the extensive swamp was able to establish. The largest area of mangrove vegetation recorded for the mid-Holocene was in the Northern Territory, where extensive mangroves (estimated 80,000 ha) flourished between 7,000 and 5,500 years BP. This was a time of sea-level stabilization, then fluvial sedimentation at these sites destroyed the habitats of mangroves (Woodroffe, et al., 1985, 1987, 1989; Grindrod, 1988). This sequence is well illustrated by the pollen record of mangrove assemblage changes over time in the South Alligator River drill-core 40 (Woodroffe, et al., 1985). Sea level was transgressive earlier in the record, with increasing influence of marine processes and corresponding changes in mangrove communities. Between 8,000 and 6,000 year BP mangrove vegetation communities kept pace with a 12 m sea-level rise, implying rates of accretion of 6 mm/a (Woodroffe, 1990). The mid-Holocene mangrove swamp was replaced after 5,500 BP by landward plant communities, as a result of sedimentary infill. A similar sequence has been reported for the nearby Mary River (Woodroffe and Mulrennan, 1993). Here, a large mangrove area developed with sea-level stabilization around 6,500 years BP, and remained until 4,000 year BP. During this time the rate of sedimentation was up to 10 mm/a, declining after 6,000 year BP. Sedimentation caused the mangrove forest to be replaced by freshwater wetlands, under which the rate of sedimentation was no more than 0.5 mm/a (Woodroffe and Mulrennan, 1993). There has been little progradation since, the shoreline more recently being modified by episodic chenier ridge deposition, channel shifting, and infill (Mulrennan and Woodroffe, 1998). In contrast to riverine mangrove settings found in estuaries, or basin mangroves which occur in partially impounded depressions (Cintron and Schaeffer Novelli,
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1984), fringe mangrove forest types are settings exposed directly to the open sea. When the fringing mangroves are degraded by human impact, their resilience reduces to result in coastal erosion (Furukawa and Wolanski, 1996; Mazda et al., 1997; Massel et al., 1999). Coastal erosion can also occur with mangrove degradation in riverine settings (Mazda et al., 2002).
2.2. Islands Thom (1984) added the category of carbonate settings to these five terrigenous settings (Table 1), distinct from those not only by the higher proportion of organic matter in sediment but also by dominance by authochthonous accumulation, derived primarily from in situ breakdown of mangrove production and root mat growth. The result is a firm organic peat. These microtidal carbonate settings include platforms such as Florida Bay, as well as oceanic islands. There is a closer association between mangroves, seagrasses, and coral reefs in island settings relative to continental margins (Linton and Warner, 2003). These associated ecosystems exist in a dynamic equilibrium influenced by contact with suitable elevations and protection provided by geomorphologic features such as shingle ridges. Sediments and nutrients, carried by any freshwater runoff are first filtered by coastal forests, then by mangrove wetlands, and finally by seagrass beds. The existence of coral reefs is directly dependant on the buffering capacity of the shoreward ecosystems which help create the oligotrophic conditions under which coral reefs flourish, so limiting the algal growth which can threaten coral reef health (Golbuu et al., 2003). At locations where poor land management leads to high sediment discharge from rivers, this role can be critical (Victor et al., 2006). In turn, coral reefs and associated shingle ridges buffer the soft sediment ecosystems of mangroves and seagrass from wave action (Frank and Jell, 2006). A further association between mangroves and coral reefs in these settings is chromophoric dissolved organic matter (CDOM), of which mangroves are important sources (Maie et al., 2006), and is carried in outwelling tidal water. This can be transported over nearshore reefs affording them some sunscreen protection through the control of UV penetration to the reef surface (Anderson et al., 2001). Mangroves of island settings tend to be more respondent to sea-level change owing to their low sedimentation rates, as reviewed in Section 5.
2.3. Inland mangroves A further category of mangrove settings not included in earlier reviews are those that occur inland, isolated from the ocean. While they are rare, they have particular geomorphic interest as they are usually relics of a former sea level or facilitated by a karst geology. Several limestone islands in the Caribbean area have mangroves in inland saline ponds connected with the ocean by submarine caves. The many small mangrove areas of Bermuda include some in inland anachaline ponds that exchange water with the ocean through submarine caves (Thomas, 1993). These occur at sea level, and water salinity is close to sea water. Thomas et al. (1992) described biotic
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characteristics of the largest mangrove ponds, finding species variability of mangrove and root biota between ponds, caused by isolation of communities. In the Bahamas, mangroves occur 50 km inland on the limestone island of Inagua at three locations with no apparent connection with the ocean (Lugo, 1981). Inland mangrove areas are cut off from the ocean by lithified beach ridges. Evaporation from the large shallow lakes raises salinity to up to 70, which has the effect of stunting the mangrove trees. Barbuda has inland mangrove ponds 2–4 km from the ocean and separated by Pleistocene beach ridges of 2–4 m in height (Stoddart et al., 1973). These inland Rhizophora thickets are dense and productive, but tended to be of lower height than the 7 m coastal Rhizophora on Barbuda. In the eastern Pacific, inland mangroves occur on many limestone islands of the Marshall Islands (Fosberg, 1975) primarily B. gymnorrhiza. In the Northern Marshalls small inland areas of mangrove are more common than coastal settings. Inland mangroves also occur in several locations on Tuvalu, primarily R. stylosa (Woodroffe, 1987). These settings are thought to have a karstic connection with the ocean. In the Indian Ocean, inland mangroves of the genus Bruguiera occur on the east shore terrace of Christmas Island (van Steenis, 1984). The stand occurs at elevations of 24–37 m above sea level, is 0.33 ha in size and 120 m inland (Woodroffe, 1988). Corals in growth position on the terrace were dated to the last interglacial sea level highstand, suggesting that the stand of mangroves has persisted in freshwater conditions as a shoreline relic during the last 120,000 years of lower sea level. In northern West Papua the mangrove Sonneratia caseolaris occurs growing at a freshwater lake around 75 m above sea level (van Steenis, 1963). This is attributed to the geological uplifting of northern New Guinea. The same species occurs 10 km inland of estuarine mangroves at Timika, southern West Papua, close to the Mile 21 replanting station of the Freeport mine (Ellison and Simmonds, 2003). This is attributed to be a relic of coastal progradation as this coast is subsiding (Ellison, 2005). In Western Australia, a large mangrove community occurs 25 km inland, at Mandora (Beard, 1967). The location is a salt creek in limestone, which is lined by mangrove trees of up to 5 m (A. marina). There is no tidal connection with the ocean at the site. The largest known inland mangrove area in the world also occurs in W.A. at Lake MacLeod. This is a former sea embayment that was closed at the south end by development of sand dunes during mid-Holocene sea-level fall. Ocean water passes underground 18 km through the limestone barrier, to rise in sinkholes in the central west part of the lake, which has permanent waters of about 6,000 ha (Handford et al., 1984). At least 22.5 ha of mangroves fringe these nontidal ponds, of A. marina stunted by high salinity (Ellison and Simmonds, 2003). These inland settings have generally narrow mangrove margins because of lack of tides. In the other mangrove settings the tidal range of the location in combination with the topographic gradient largely controls the distribution of mangroves.
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3. TIDAL R ANGE AND SEA -LEVEL C ONTROL Tidal processes in estuaries and other coastal mangrove settings operate between high water and low water levels in all cases at least once a day, so constituting a critical part of fine sediment movement and nutrient delivery to these highly productive intertidal ecosystems. Mangrove forest structure is characterized by “zones” of tree species, in patterns that often run perpendicular to the shore, which reflects strong spatial patterns in inundation frequency, sediment and nutrient inputs, salinity, and biological processes (e.g., bioturbation and predation). These processes are dependent on the underlying geomorphology as expressed in the microtopography (Semeniuk, 1994), in relation to sea-level position and tidal range. Figure 2 shows the simplified mangrove zonation from mean sea level (MSL), with different species each favoring a microelevational range up to hightide levels where the mangroves margin landward with lowland rainforest. Mangroves occur between the high tide and MSLs, making them particularly sensitive to sea-level rise and other factors that influence hydrology of the intertidal zone. Mangroves are widely recognized in the literature to occur between these mean tide and high-tide elevations, which has been demonstrated in relatively few studies. Within this intertidal habitat of mangroves, species have different preferences for elevation, salinity, and frequency of inundation, resulting in these species zones (Duke et al., 1998). Several surveys from large mangrove systems in northern Queensland as well as Darwin Harbour have shown that the mangrove seaward edge occurs at MSL, which supports the widespread understanding of mangrove habitat occurrence in the upper half of the tidal spectrum. Mangroves of Coral Creek, northern Hinchinbrook Island, were surveyed in detail in 1978 by the Australian Survey Office. This showed the mangrove seaward edge to occur at –0.18 to þ0.07 m relative to MSL (Boto and Bunt, 1981; Wolanski et al., 1992). Similar surveys were
MHWS MSL MLWS
Avicennia Sonneratia
Rhizophora
Bruguiera
Nypa
Rain forest
Figure 2 The typical mangrove zonation from the South Papuan center of mangrove biodiversity. MHWS = Mean high water spring tide level; MLWS = Mean low water spring tide level.
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carried out by the Australian Survey Office in 1974 at Zoe Bay, SE Hinchinbrook Island, showing the mangrove seaward edge to occur at –0.18 to þ0.02 m above MSL, and in 1985 the Murray River estuary just north of Hinchinbrook, finding the mangrove seaward margin to be at approximately 0.0 Australian Height Datum, which is at sea level (Priest, 1990). Accurate survey of mangrove substrate elevations in microtidal Bermuda demonstrated elevations of –0.2 m below MSL at the seaward edge of mangroves, and 0.3 m above MSL at the landward edge (Ellison, 1993). In extensive continental mangroves of southern West Papua, the elevation of the seaward edge of mangroves was surveyed accurately at the mouth of two adjacent estuaries, and at species zone transitions from here to the landward mangrove margin with lowland rainforest (Figure 2). At both estuaries scattered trees occurred to seaward at –0.35 or 0.40 m below MSL (Ellison, 2005), and the seaward transition from these scattered pioneer trees to dense forest occurred at 0.15 m above MSL, and. The transition from Rhizophora to landward Bruguiera forest occurred at þ1.1 + 0.1 m. The landward edge of mangroves with freshwater forest was found to occur at þ1.6 + 0.2 m above MSL, within a tidal range of 3.3–3.6 m. Offshore in the Great Barrier Reef, three marine-dominated low island mangrove systems on the were accurately surveyed by the author at Low Isles (16230 S, 145340 E), Three Isles (15060 S, 145.250 E), and Pipon (14070 S, 145.120 E). The mean elevation of the mangrove/lagoon margins was found to be 0.36 m below MSL, and using ANOVA the differences in means were not significant between islands. These elevation studies indicate the sensitivity of mangrove zones to sea-level position, hence inferring disruption and relocation with small amounts of sea-level change. The lower elevation than river sites could be caused by the efficient aeration of substrate sediment at low tide, owing to the low field capacity of coral sand. The position of mangroves, salt marsh, and salt flats relative to sea level emphasizes the sensitivity of these ecosystems to sea-level rise, and highlights the potential for disruption and relocation of vegetation zones and structure of the vegetation with global climate change. These issues are reviewed below by reference to paleoenvironmental sedimentological records.
4. SEDIMENTATION IN M ANGROVES Mangroves exist in active coastal settings, influenced by tidal movements, wave action, catchment runoff including floods, and storm action. These are all direct influences on the sediment balance of the mangrove area, as well as indirect influences such as sediment supply either upstream, along shore, or offshore. This sediment budget is summarized in Figure 3. Major sources include input from rivers in river mouth situations, longshore transport, gains from offshore which occur mostly during high-magnitude storm/tsunami events, autochthonous input from mangrove productivity, and relative sea-level fall. Major losses include river sediment not trapped in the mangrove cell, longshore transport downcoast, mangrove litter outwelling, and relative sea-level rise. The sediment budget is a balance of
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Longshore drift
Litter fall
Sediment store Marine inputs (storms/tsunamis)
Figure 3
Litter outwelling
River-transported sediment
Sediment budget of a mangrove swamp.
volumes of sediment entering or leaving the mangrove environment, influencing whether the environment is erosional or accretionary of a number of timescales. Sedimentation is also influenced at the smaller scale by tidal currents, which in mangrove tidal creeks are generally asymmetrical with peak ebb tidal currents in creeks of large areas of mangrove recorded to be 20–50% higher than peak flood tidal currents (Wolanski et al., 1992). This tends to maintain a deep main channel, but high friction within the densely vegetated forest tends there to reduce tidal velocities (Spenceley, 1977). This vegetational friction on water movement combined with flocculation of clays contributes to substrate accretion (Furukawa and Wolanski, 1996; Furukawa et al., 1997). The density of mangrove vegetation further exerts a drag coefficient on tidal waters to protect the sediment from erosion (Mazda et al., 1995). Kraus et al. (2003) found that while Rhizophora prop roots assist in the settling of suspended sediment from estuarine waters, Sonneratia pneumatophores are more successful in maintaining sediment elevation over the longer term. Avicennia pneumatophores have also shown to promote sediment accretion (Young and Harvey, 1996) demonstrating the usefulness of these seaward margin species in claiming new mangrove habitat. Substrate accretion or substrate elevation change is the net expression of the site sediment budget (Figure 3), consequential from both sources and losses. Accretion is also influenced by in situ processes such as decomposition of organic matter, compaction of the sediment column, and root mat growth. Recent sedimentation rates in mangrove swamps (<100 years) have been measured by a variety of techniques, such as the laying of an exotic layer on the mangrove mud surface and later measuring its depth of burial. Techniques used include brickdust (Bird, 1971), and feldspar clay (Cahoon and Lynch,
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1997). Sediment accretion in the last 50–100 years has been analyzed from mangroves using 137Cs and 210Pb analyses (Lynch, et al., 1989; Cahoon and Lynch, 1997), and sedimentation stakes (Bird and Barson, 1977; Spenceley, 1977, 1982). These short-term records of mangrove sedimentation show that the majority accreted at rates of <5 mm/a, with a maximum 10 mm/a. Within mangroves of the Great Barrier Reef sedimentation in mangroves has been observed to vary between –11 mm/a (erosion) to 10 mm/a (Furukawa and Wolanski, 1996; Bird and Barson, 1977; Spenceley, 1977, 1982; Bryce et al., 2003). In a study of sedimentation in southern Australia, sedimentation was higher in mangroves compared to salt marsh [approximately 5 mm/a in mangroves and 2.5 mm/a in salt marsh (Rogers et al., 2005)]. However, Rogers et al. (2006) found that where salt marshes were being encroached by mangroves, this increased the autocompaction so reducing the vertical accretion. Using a Surface Elevation Table (Cahoon et al., 2002) in combination with a marker horizon enables accretion to be distinguished from both shallow subsidence and elevation change. Across a variation of sites, Cahoon (2006) found that vertical accretion averaged 5 mm/a, elevation change 1 mm/a, and shallow subsidence 4 mm/a. These results over the last few years correspond well with radiocarbon dated long-term rates of sediment accretion in mangroves (Ellison and Stoddart, 1991; Ellison, 2005). In conditions of sedimentation surplus, mangroves colonize seaward either into bays especially offshore of river mouths or over reef flats. North of Cairns in Queensland, the Barron River delivers sediment from a fairly disturbed catchment to a growing mangrove estuary. Here Bird and Barson (1977) measured accretion under mangroves of 5 mm/a under the Avicennia seaward margin, and up to 3 mm/a within the Rhizophora zone. Panapitukkul et al. (1998) demonstrated mangrove progradation of approximately 38 m/a at Pak Phanang Bay in SE Thailand with high rates of river delivery of sediment. Mangroves have been shown to promote coastal stability in conditions of erosion and high human impact also from this region (Mazda et al., 2002; Thampanya et al., 2006).
4.1. Excessive sedimentation While these studies show that mangroves deal well with accretion rates of generally <5 mm/a, allowing keep up with relative sea-level rise and extension of the mangrove habitat seaward, mangroves are however known to suffer mortality in conditions of excessive sedimentation. In deltaic areas that receive massive amounts of sediment, mangroves opportunistically colonise suitable habitats but may get disrupted by changes in sediment supply and microtopography. West (1956) described deposition of sand in stands of Rhizophora at the Atrato Delta in Colombia causing the killing of mature trees. Thom (1967) described the distributions of mangroves in the Tabasco Delta, Mexico, as being subject to the changing sedimentation patterns of the river. Atmadja and Soerojo (1994) observed that some Indonesian mangroves are subject to excessive sedimentation from river floods, causing death of some mangrove
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species, especially Avicennia and Sonneratia. Lugo and Cintron (1975, p. 840) described excess sedimentation in Puerto Rican mangroves as a result of floods, leading to interruption of root and soil gas exchange, and eventual death of the trees. Similar problems have been demonstrated experimentally by Terrados et al. (1997) showing that sediment burial of 8 cm and above retarded growth and increased mortality of R. apiculata seedlings. They suggested that rapid sediment accretion altered oxygen supply to the hypocotyl root system. There are numerous accounts of sedimentation as a result of human disturbance causing mortality in mangroves. Hutchings and Saenger (1987) described that if dredge spoil is discharged over the roots of mangroves, it leads to their death. In NW Australia, at King Bay, 2 ha of mangroves were buried and deteriorated when an adjacent dredge stockpile eroded (Gordon, 1988). At Point Samson, harbor dredge spoil smothered pneumatophores and killed fringing A. marina (Gordon, 1988). Allingham and Neil (1995) described mangrove death from smothering at Mud Island, off Brisbane airport in Queensland. This was from deposition of dredge spoil on the island as shingle ridges, and mangroves died from smothering adjacent to these. In Singapore, Lee et al. (1996) described at a coastal construction site mature individuals of Avicennia spp. and S. alba dying or dead due to fast accretion of sediments along the embankments, burying or covering the pneumatophores. In the Federated States of Micronesia, Devoe (1992) reported from Yap of sedimentation in mangroves adjacent to road construction, causing destruction of mangroves. In the Solomon Islands, Kwanairara (1992) reported degradation of mangroves with siltation coming in from adjacent development. In New Caledonia, large-scale nickel mining results in massive soil erosion (Wodzicki, 1981), and siltation in mangrove swamps of mine waste is a problem. Clay deposits in mangrove swamps, particularly after storms, caused localized areas of mangrove decline. At Bowen (N. Queensland) up to 0.5 ha of A. marina was killed by about 12 cm of sediment in-wash from the adjacent Bowen Coke Works in the late 1980s (Ellison, 1999). In Saudi Arabia, large areas of mangroves (principally A. marina) were destroyed or degraded by coastal infilling (Price et al., 1987; Aleem, 1990). Degradation of mangroves from sand and silt sedimentation is reported from Guinea-Bissau by Simao (1993). Deposition of sand onto mangroves is described as a stress by Echevarria and Sarabia (1993) from the Tumbes River in Peru. In Colombia, Alvarez-Leon (1993) described mining at Cartagena Bay as causing sedimentation affecting adjacent mangroves, along with soil erosion from catchment deforestation. In Cuba, Padron et al. (1993) described sand accumulation over prop roots and pneumatophores as affecting mangroves. Excess input of sediment to mangroves can cause impacts ranging from reduced vigour to death, depending on the amount and type of sedimentation, and the species involved. Impacts start to occur with sudden deposition of at least 5 cm of sediment, even if the aerial roots are taller. The cause seems to be the smothering of lenticel sites on aerial roots, as these tend to be concentrated lower on the aerial root to reduce the gas diffusion gradient (Scholander et al., 1955).
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5. MANGROVES AS SEA-LEVEL I NDICATORS Given that mangroves generally grow between MSL and mean high water, their sedimentary records have been used as precise sea-level indicators. Other indicators such as coral can grow at a variety of depths, while wave cut notches can give an accurate height but provide little to date (Ellison, 1989). Such research has recently gained significance in providing reconstruction of how mangroves responded to past sea-level changes, with growing indication that global eustatic sea levels are no longer stable. Hence there is increasing evidence that mangroves are affected by coastal environmental change, particularly hydrological variations and sea-level rise (Ellison and Farnsworth, 1997). The response of mangroves to such impacts tends to be gradual and, particularly in undisturbed systems, manifested typically as a change in their extent, structure, and species composition and hence their species zonation (Figure 4). As mangroves are sensitive to even minor transitions in coastal conditions (e.g., altered drainage patterns, saltwater intrusion, accretion or erosion in response to sea-level variations), changes in the zonation of these ecosystems are often indicative of broader scale changes and associated impacts in coastal regions (Blasco et al., 1996; Ellison and Farnsworth, 1997). Comparing present trends in species and communities with paleoecological records of past extents provides excellent information on how they may respond to climate change (Hansen et al., 2001; Hansen, 2003). Mangrove response to sealevel rise has been investigated by the author by reconstruction of Holocene analogues in the Cayman Islands, Tonga, and southern New Guinea as well as Bermuda. In summary radiocarbon dating of stratigraphy determined a sediment accretion rate of 1 mm/a for the low island locations, and up to 1.5 mm/a in two estuaries of southern New Guinea. 5
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Figure 4 Holocene stratigraphy of Missionary Bay mangrove swamp (adapted from Grindrod and Rhodes, 1984 with the Townsville sea-level curve of Chappell et al., 1983 redrawn to the same vertical scale). MHW = Mean high water; MLW = Mean low water.
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Based on an assessment of stratigraphic records of island mangroves during sealevel changes of the Holocene period, Ellison and Stoddart (1991) generalize that low island mangroves, which generally accumulate vegetative detritus for substrate buildup and lack a large source of inorganic sediment, can keep up with a 1.2 mm/a relative sea-level rise rate. Mangroves of high islands and continental coastlines, which have relatively large supplies of terrigenous inorganic and organic sediment from rivers and longshore drift, can keep pace with a 4.5 mm/a relative sea-level rise rate (Ellison and Stoddart, 1991). This rate is of significance because as established in Sections 3 and 4, within the intertidal habitat of mangroves, species have different preferences for elevation, salinity, and frequency of inundation, resulting in species zones. Elevation of the ground surface under mangroves can be raised, by accumulation of vegetative detritus or inorganic matter brought in by tides or rivers. If the sedimentation rate keeps pace with rising sea level, then the salinity and frequency of inundation preferences of mangrove species zones will remain largely unaffected. If the rate of sea-level rise exceeds the rate of sedimentation, then the mangrove species zones will migrate inland to their preferred elevation, and seaward margins will die back. The accumulation of sediment under mangroves will help to compensate for rising sea levels. In Bermuda, tide gauge measured rates of sea-level rise over the last century have been within the rates projected for this century. This allows observation of mangrove response to rising sea level at a location that has been experiencing sea-level rise in the past, at a rate within the ranges projected for this century elsewhere in the world. The largest Bermudan mangrove area at Hungry Bay had existed for the last 2,000 years during a period of sea-level stability (Figure 1), and during the last century lost 26% of its area due to retreat of its seaward edge (Figure 5). Survey showed that swamp elevations were lower in the tidal spectrum than normal, and mangroves at the seaward margin were under inundation stress (Ellison, 1993, 1996). This Bermuda site demonstrated that mangrove sediment is subject to erosion by rising sea levels, with removal of mangrove substrate (above MSL) and with some deposition subtidally offshore of the mangroves (Ellison, 1993). Similar erosion patterns to Bermuda, with reversed succession as elevation declines, have been described by Semeniuk (1980) in NW Australia. The effect of sheet erosion on mangrove zonation is migration of pioneer/seaward mangroves into more landward zones. The effect of cliffing on mangrove zonation is loss of the seaward zone, leading to truncated zonation and narrow fringes. The effect of tidal creek erosion is slumping of banks and loss of trees. In the Northern Territory of Australia, Woodroffe and Mulrennan (1993) have documented dramatic recent changes to the Lower Mary River floodplain, with saltwater intrusion and upstream expansion of the tidal creek network. This has resulted in the death of freshwater wetland communities with loss of 60 km2 of Melaleuca forest and upstream invasion of mangroves. There are a number of possible reasons for these events, including relative sea-level rise (Woodroffe, 1995). Similar, though less spectacular, extension of creeks has occurred on other river systems, such as the Alligator River (Woodroffe, 1995). Loss of freshwater
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Figure 5 (a) Location of Hungry Bay mangrove swamp, Bermuda. (b) Extent of mangrove dieback at Hungry Bay. (c) Stratigraphy of transect from sea to land across Hungry Bay mangrove swamp (adapted from Ellison, 1993).
wetlands with saline intrusion is documented in the Florida Keys (Ross et al., 1994), where longer tide records have enabled researchers to attribute the cause to relative sea-level rise. The Cayman Islands experienced a slowly rising sea level to present levels during the Holocene (Figure 1), and at 4–5 m below present sea level 20 km2 of
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Figure 6 (a) Location of North Sound and adjacent mangrove swamps, Grand Cayman. (b) Extent of mangrove dieback at Little Sound. (c) Stratigraphy of transect from sea to land across North Sound.
mangroves existed under the present North Sound lagoon until between 4,080 and 3,230 years BP (Figure 6). This was found by coring beneath the current lagoon to find mangrove peat buried beneath seagrass beds. This mangrove recession event and replacement by lagoon environments is what occurs during relatively slow sea-level rise, and mangroves recolonize at a higher elevation. In Tonga, mangrove response to sea-level rise was investigated at the largest pocket of mangroves in Tonga, in the western Fanga’uta Lagoon (Figure 7). Figure 7c shows mangrove peat occurring between about 2.5 and 1.5 m below present sea level, indicating a large mangrove swamp that persisted between 7,000 and 5,500 year BP during sea-level rise reconstructed to be at a rate of 1.2 mm/a. Around 5,500 year BP this mangrove area died back to become a lagoon, persisting
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Figure 7 (a) Location of mangrove swamps, Tongatapu. (b) Location of the stratgraphy transect across the Folaha mangrove swamp. (c) Stratigraphy of transect from sea to land across Folaha mangrove swamp (adapted from Ellison, 1989).
above this level as just a narrow fringe. The mangrove swamp later reestablished over the whole site following sea-level fall in the later Holocene, following the regional sea-level curve (Figure 1). Gradual retreat of mangrove zones with slowly rising sea level has also been demonstrated from the extensive coastal swamps of southern New Guinea (West Papua) (Ellison, 2005). Figure 8 is a pollen stratigraphic diagram showing a Bruguiera zone present at the core site for most of the Holocene, replaced around 3,000 years ago by a Rhizophora zone. This sequence of events was replicated at four other core sites throughout this and the adjacent Ajkwa Estuary. Landward
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on a m s an a em a a le um hu ia ati us ia st ari pu a er lia och eni stic nn ner nth orn pto ec car tier nitz a l e a l ic on ca sb am xco ylo eri um yp av ten sp cro v C A O A L A S A X H E D S N
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Figure 8 Mangrove pollen diagram from theTipoeka Estuary, southern New Guinea (adapted from Ellison, 2005).
Bruguiera being replaced by seaward Rhizophora indicates landward retreat of the mangroves, at a rate of sea-level rise determined to be only 0.67 mm/a. These paleoenvironmental records of mangroves from a number of locations with different Holocene sea-level curves (Figure 1) all shown sensitivity to sea-level rise, including dieback and massive mortality events Sediment supply determines mangrove ability to keep up with sea-level rise. Mangroves of low relief islands in carbonate settings that lack rivers are likely to be the most sensitive to sea-level rise, owing to their sediment-deficit environments. However, as demonstrated from southern New Guinea, continental mangroves (such as mainland Australia) will also demonstrate significant mortality and attempt relocation inland.
6. S TORMS/TSUNAMIS Mangroves have an important role in stabilising coasts during storm and tsunami events, both by frictional reduction of energy of waves, and promoting sedimentary resilience to erosion through the root mat (Smith et al., 1994; Massel et al., 1999; Mazda et al., 2002; Dahdouh-Guebas et al., 2005; Danielsen et al., 2005; Katharesan and Rajendran, 2005; Vermaat and Thampanya, 2006). Storm impacts geomorphologically can deposit marine sediments into mangrove seaward margins to either cause mortality or build shore parallel sand ridges or chenier ridges.
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Damage in mangroves following cyclones is usually a narrow zone of coastal wave damage, and complete defoliation over the narrow area of cyclone paths (Jaffrar, 1992). In Florida, Craighead and Gilbert (1962) described the effects of Hurricane Donna on mangroves of southern Florida, particularly Rhizophora mangle, Avicennia germinans, Laguncularia racemosa, and Conocarpus erectus. Widespread death occurred, averaging 25–75% death over large areas and locally reaching 90%. This was not due to mechanical damage and defoliation, since some trees put forth new leaves, but was due to deposition of up to 13 cm of sediment that caused root damage and oxygen deficiency. Chappell and Grindrod (1984) described recent burial of Rhizophora mangroves by 70 cm of shell deposited adjacent to a storm chenier ridge. Unlike the temperate equivalent environment to mangroves of salt marshes, there is little evidence in mangrove sedimentology of paleotsunamis. These are marine sourced shallow (<20 cm) coarse sand layers in fine peaty marsh lithology that drape or truncate the preexisting marsh surface, tend to thin landward laterally, sand textures fine landward, contain marine macrofossils, may show fine/coarse laminae indicative of sequential waves, and are contemporaneously datable from adjacent sites (Williams et al., 2005). In the Caribbean, Scheffers et al. (2005) did not find evidence of paleotsunamis in mangrove areas, rather as boulder deposits incorporated into tephra depositions. However, mangrove and back pond sedimentary environments in Puerto Rico have revealed a record of past storm surges caused by cyclones (Donnelly, 2005). Large-scale flooding is frequently associated with tropical cyclone impacts, and there are several accounts of sustained inundation after storms causing mangrove mortality. In Northern Natal, mangroves occur as scattered individuals in the St Lucia Lake system, about 15 km above the estuary mouth. Following a cyclone, the lake level rose some 2.5 m above mean summer level, dropping to 1.7–1.5 m 2 weeks later, and returning to typical summer levels by 2 months later (Forbes and Cyrus, 1992). Mortalities of 36.7% to B. gymnorrhiza were recorded in the lake, and in the channel between the lake and the estuary 33% of Bruguiera over 1 m tall and 16% of A. marina suffered mortality. Steinke and Ward (1989) noted rapid mortality of mangroves up to 1.1 m tall, while taller individuals up to 3.5 m died 3–4 months later. Inundation changes affecting mangroves as caused by other factors are reviewed below.
7. INUNDATION C HANGES AFFECTING M ANGROVES As reviewed in Section 3, many mangrove species have restricted preferences for inundation frequency as controlled by the level of tidal waters relative to their substrate. However, mangroves frequently occur in mobile sedimentary environments where their inundation patterns can be altered by tidal restriction, particularly in associated with coastal development. Sustained inundation following blockage of mangrove draining channels is recorded from several sites. In Brunei-Daressalam (Borneo) a 50 ha Avicennia mangrove area was flooded following enclosure by a natural sand bar (Choy and Booth,
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1994). Water depths of 0.5 m above Mean high water spring tide level (MHWS) occurred for 8 weeks before the floodwater was released, though salinities did not vary from normal brackish levels. A. marina suffered substantial losses of 1.5–5.0 m tall plants, while most trees >5 m survived. Choy and Booth (1994) record necrosis (death of tissue) to bark and surface wood tissues on mangroves stems below flood level, and death of pneumatophores of trees <5 m tall. They also noted side shoots developing from higher up on Avicennia stems, similar to Rhizophora prop roots, and stimulation of flowering and fruiting by surviving plants. In Natal, 120 km north of St. Lucia, mass mortality of mangroves occurred at Kosi Estuary in 1965 (Breen and Hill, 1969). Following closure of the river mouth, water levels in the mangrove estuary rose by about 0.3 m over the succeeding 140 days. Mass mortality occurred primarily of A. marina, with the bark of all trees rotted around 0.2 m above mud level. Mangroves on higher ground, chiefly Lumnitzera racemosa survived. Following river flooding in 1956, 42 ha of Bruguiera in Tjilatjap, Central Java were killed, attributed to inundation by river water of low salinity, along with sediment deposition (Soerianegara, 1968). A. germinans areas at Clam Bay, Florida died following blockage of tidal channels such that they were inudated by accumulated rainwater (Turner and Lewis, 1997). In NE Puerto Rico mortality of 15 ha of A. germinans and Laguncularia racemosa occurred when dredging caused abrupt change to the mangrove hydrological regime, resulting in permanent flooding (Jimenez et al., 1985). The cause of death was attributed to reduced soil oxygen. In Colombia, 30,000 ha of mangrove mortality occurred when human-induced changes to the hydrological system caused hypersalinization of soils (Elster et al., 1999). At Clarence Estuary in NSW, structural flood mitigation works caused loss of tidal fluctuations in mangrove areas (Pollard and Hannan, 1994). This caused near total loss of mangroves above the flood gates, from reduced tidal inundation. In North West Australia, road construction dammed tidal creeks at Pope’s Nose Creek to cause death of 6 ha of A. marina and Ceriops tagal (Gordon, 1988). Construction of the Dampier Salt evaporator pond caused permanent ponding and mass mortality to about 11 km2 of mangroves (Gordon, 1988). Impoundment of mangroves for mosquito control in Florida provides several examples of prolonged inundation causing mangrove mortality (Brockmeyer et al., 1997). Extensive death of A. germinans and R. mangle occurred at India River, East Florida following 4 months flooding of 30–45 cm depth (Harrington and Harrington, 1982). Naidoo (1983) found that prolonged flooding resulted in lower ability of leaves to conduct water, an increase in stomatal closing, and degeneration of chloroplasts in B. gymnorrhiza, leading to reduced rates of photosynthesis. When lenticels of aerial roots become inundated, oxygen concentrations in the plant fall dramatically (Scholander, et al., 1955). If inundation is sustained, low oxygen conditions occur and mortalities follow. This is thought to have been the reason for mortality of 15 ha of A. germinans in NE Puerto Rico (Jimenez et al., 1985), following permanent flooding caused by adjacent dredging. Studies of inundation effects on mangrove growth have also been carried out in simulation of projected sea-level rise. Ellison and Farnsworth (1997) grew R. mangle seedlings with a 16 cm increase in water level. They maintained normal tidal fluctuation
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around this raised mean, and found that growth slowed dramatically at the sapling stage, until after 2.5 years, plants were 10–20% smaller than control plants. Mangrove seedlings are thought to be more susceptible to death from inundation than adult trees. Elster et al. (1999) found that during sustained flooding of about 30 cm, all seedlings of A. germinans and Laguncularia racemosa died within 1 month. Hence, while mangroves are tolerant of tidal inundation, massive mangrove tree mortality occurs when mangroves are flooded for prolonged periods. This is due to the interruption of gas exchange in the root systems (Jimenez et al., 1985). While natural blockage of tidal creeks occurs is more common, however, that human-induced hydrological changes resulting in sustained inundation causes mangrove mortality.
8. C ONCLUSIONS Mangroves as coastal biogenic sedimentary systems stabilise massive amounts of sediment to benefit nearshore coral reefs. However, mangroves are relatively sensitive to changes in sea level and inundation (Blasco et al., 1996), recording this sensitivity in the consistent microelevational ranges of the different mangrove species zones (Ellison, 2005). Massive mortality in mangrove forests occurs where all size classes are affected, in response to rapid environmental change, compared to normal mortality of older trees (Jimenez and Lugo, 1985). Occurrence of mass mortality of mangroves has been demonstrated to be caused by a number of causes related to geomorphic or sedimentological factors. These include excess sedimentation, sedimentation rates not keeping pace with sea-level rise, hydrological blockage causing either sustained or restricted inundation (Hatton and Couto, 1992), or sediment erosion. Other major causes of mangrove mortality include oil spills (Lewis, 1983; Duke et al., 1997), not including a growing list of countries where 50–80% of mangroves have been cleared in the last 15 years (Diop, 2003). In these areas, reduced mangrove area and health will increase the threat to human safety and shoreline development from coastal hazards such as tsunamis, erosion, flooding, and storm waves and surges (Danielson et al., 2005).
ACKNOWLEDGMENT I thank Peter Manning for drawing all the figures in this chapter.
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Spenceley, A.P., 1982. Sedimentation patterns in a mangal on Magnetic Island, nr. Townsville, North Queensland, Australia. Singa p. J. Trop. Geogr. 3, 100–107. Steinke, T.D., Ward, C.J., 1989. Some effects of the cyclones Domoina and Imboa on mangrove communities in the St. Lucia estuary. S. Afr. J. Bot. 55, 340–348. Stoddart, D.R., Bryan, G.W., Gibbs, P.E., 1973. Inland mangroves and water chemistry, Barbuda, West Indies. J. Nat. Hist. 7, 33–46. Thampanya, U., Vermaat, J.E., Sinsakul, S., Panapitukkul, N., 2006. Coastal erosion and mangrove progradation of Southern Thailand. Estuar. Coast. Shelf Sci. 68, 75–85. Terrados, J., Tampahnya, U., Srichai, N., Kheowvongsri, P., Geertz-Hanzen, O., Borromthanarath, S., Panapitukkul, N., Duarte, C.M., 1997. The effect of increased sediment accretion on the survival and growth of Rhizophora apiculata seedlings. Estuar. Coast. Shelf Sci. 45, 697–701. Thom, B.G., 1967. Mangrove ecology and deltaic geomorphology, Tabasco, Mexico. J. Ecol. 55, 301–343. Thom, B.G., 1982. Coastal landforms and geomorphic processes. In: Snedaker, S.C., Snedaker, J.G. (Eds.), The Mangrove Ecosystem: Research Methods. UNESCO, Paris, pp. 3–15. Thom, B.G., 1984. Mangrove ecology – a geomorphological perspective. In: Clough, B.F. (Ed.), Mangrove Ecosystems in Australia: Structure, Function and Management. Australian Institute of Marine Science, Townsville, Australia, pp. 3–18. Thom, B.G., Wright, L.D., 1983. Geomorphology of the Purari delta. In: Petri, T. (Ed.), The Purari – Tropical Environment of a High Rainfall River Basin. Mongraphs in Biology, vol. 51. Dr. W. Junk, The Hague, pp. 47–65. Thomas, M.L.H., 1993. Mangrove swamps in Bermuda. Atoll Res. Bull. 386, 1–17. Thomas, M.L.H, Logan, A., Eakins, K.E., Mathers, S.M., 1992. Biotic characteristics of the anchialine ponds of Bermuda. Bull. Mar. Sci. 50, 133–157. Turner, R.E., Lewis III, R.R., 1997. Hydrologic restoration of coastal wetlands. Wetlands Ecol. Manage. 4, 65–72. Vermaat, J.E., Thampanya, U., 2006. Mangroves mitigate tsunami damage: a further response. Estuar. Coast. Shelf Sci. 69, 1–3. van Steenis, C.G.G.J., 1963. Miscellaneous notes on New Guinea Plants VII. Nova Guinea Bot. 12, 189. van Steenis, C.G.G.J., 1984. Three more mangrove trees growing locally in nature in freshwater. Blumea 29, 395–397. Victor, S., Neth, L., Golbuu, Y., Wolanski, E, Richmond, R.H., 2006. Sedimentation in mangroves and coral reefs in a wet tropical island, Pohnpei, Micronesia. Estuar. Coast. Shelf Sci. 66, 409–416. West, R.C., 1956. Mangrove swamps of the Pacific coast of Colombia. Ann. Am. Assoc. Geogr. 46, 98–121. Williams, H.F.L., Hutchinson, I., Nelson, A.R., 2005. Multiple sources for late-Holocene tsunamis at Discovery Bay, Washington State, USA. The Holocene 15, 60–73. Wodzicki, K., 1981. Some nature conservation problems in the South Pacific. Biol. Conserv. 21, 5–18. Wolanski, E., Mazda, Y., Ridd, P., 1992. Mangrove hydrodynamics. In: Robertson, A.I., Alongi, D.M. (Eds.), Tropical Mangrove Ecosystems. American Geophysical Union, Washington DC, pp. 43–62. Woodroffe, C.D., 1987 Pacific island mangroves: distributions and environmental settings. Pac. Sci. 41, 166–185. Woodroffe, C.D., 1988. Relict mangrove stand on last Interglacial terrace, Christmas Island, Indian Ocean. J. Trop. Ecol. 4, 1–17. Woodroffe, C.D., 1990. The impact of sea-level rise on mangrove shorelines. Prog. Phys. Geogr. 14, 483–520. Woodroffe, C.D., 1992. Mangrove sediments and geomorphology. In: Robertson, A.I., Alongi, D.M. (Eds.), Tropical Mangrove Ecosystems. American Geophysical Union, Washington DC, pp. 7–42. Woodroffe, C.D., 1995. Response of tide dominated mangrove shorelines in Northern Australia to anticipated sea-level rise. Earth Surf. Proc. Land. 20, 65–85. Woodroffe, C.D., Chappell, J., Thom, B.G., Wallensky, E., 1989. Depositional model of a macrotidal estuary and floodplain, South Alligator River, Northern Australia. Sedimentology 36, 737–756.
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Woodroffe, C.D., Mulrennan, M.E., 1993. Geomorphology of the Lower Mary River Plains, Northern Territory. North Australia Research Unit, Darwin, 152pp. Woodroffe, C.D., Thom, B.G., Chappell, J., 1985. Development of widespread mangrove swamps in mid-Holocene times in northern Australia. Nature 317, 711–713. Woodroffe, C.D., Thom, B.G., Chappell, J., Head, J., 1987. Relative sea level in the South Alligator River Region, North Australia during the Holocene. Search 18, 198–200. Wright, L.D., Coleman, J.D., Erickson, M.W., 1974. Analysis of Major River Systems and their Deltas: Morphologic and Process Comparisons. Baton Rouge, Coastal Studies Institute, Technical Report No. 156, Louisiana State University Press, Baton Rouge, LA, 114pp. Young, B.M., Harvey, L.E., 1996. A spatial analysis of the relationship between mangrove (Avicennia marina var. australasica) physiognomy and sediment accretion in Hauraki Plains, New Zealand. Estuar. Coast. Shelf Sci. 42, 231–246.
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C H A P T E R
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G EOMORPHOLOGY AND S EDIMENTOLOGY OF M ANGROVES AND S ALT M ARSHES : T HE F ORMATION OF G EOBOTANICAL U NITS Rube´n J. Lara, Claudio F. Szlafsztein, Marcelo C.L. Cohen, Julian Oxmann, Bettina B. Schmitt, and Pedro W.M. Souza Filho Contents 1. Driving Forces Determining Main Morphology and Vegetation Types in the Coastal Zone 2. Depositional Environment and Substrate Formation for the Development of Mangroves and Salt Marshes 3. Influence of Sea Level and Climate Oscillations on Local Geobotanical Features 4. Major Factors Leading to the Development of Salt Marshes and Mangroves at the Amazon Coastal Region: An Integrated Analysis 5. Influence of Geomorphology and Inundation Regime of Geobotanical Units on their Sediment Biogeochemistry 5.1. Regularly inundated wetlands 5.2. Rarely inundated wetlands 5.3. Waterlogged wetlands 5.4. Salt marshes and mangroves: Nutrient sources or sinks? References
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1. D RIVING FORCES D ETERMINING MAIN MORPHOLOGY AND V EGETATION T YPES IN THE C OASTAL Z ONE Geomorphology, tectonics, and sea level have a major influence on the global geographical distribution, regional characteristics, and local structure of salt marshes and mangroves ecosystems. The tectonic setting is responsible for formation of large-scale features, which control the coastal geomorphological evolution. Geomorphology is concerned with phenomena over a vast range of temporal and spatial scales. Considering the last aspect, it is possible to categorize the landforms from megascales (>103 km linear and area >106 km2) to microscales (<0.5 km linear and area <0.25 km2) (Summerfield, 1991). The landform scale is also related with the major controlling factors (endogenic and exogenic) influencing landform genesis, Coastal Wetlands: An Integrated Ecosystem Approach
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Table 1 Main types of coastal morphology and wetland vegetation of active and passive margins (adapted from Short, 1999)
Tectonics Landforms Weathering Drainage Sediment quantity Sediment grain size Coastal morphology Wave attenuation Tide range Occurrence of salt marshes or mangroves
Convergent margin
Passive margin
Active, earthquakes Mountains Narrow continental shelf Deep-sea trough Physical, mass movement Short, steep streams Low Fine to coarse Rocky, few beaches Low Minimal amplification Scarce
Quiescent, stable Coastal aggradation plains Wide, low continental shelf Continental slope Chemical, fluvial Long, meandering rivers High Fine Extensive barriers and deltas Moderate to high Enhanced Frequent
from continental motion and very long-term climatic change to an individual earthquake and meteorological events. This kind of hierarchy is also observed in the shelf and coastal zone (Table 1). The largest scale is principally controlled by plate tectonics, while processes such as erosion, transport, and deposition further shape the mid- to small-scale features of the coastal zone. The type of sediment can determine specific characteristics within the shore zone influencing the vegetation type. Taking into account the major control scales and their related processes, the coastal type at the large scale is determined by the position of the area relative to plate margins, leading to a division of the world’s coastline into active margin, marginal sea, and passive margin types (Summerfield, 1991; Short, 1999). Active margin coasts, also referred to as collision or convergent coasts, lie along convergent plate boundaries, such as the west coast of South America and are characterized by the delivery of relatively coarse sediments from mountainous catchments to a coastal zone typically consisting of a narrow continental shelf and deep water not far offshore. They often support only rocky shorelines, with poor beach, reef, salt marsh, or mangrove development. At marginal sea coasts and at passive margin coasts, wide continental shelves can occur, favoring extensive growth of marshes and mangroves along the coastline. Marginal sea coasts are related to collision margins, the coastlines of which developed in the intraplate domain bounded by the continental ridge. They are frequently modified by fluvial plains and deltas; their hinterland may vary considerably in relief and their adjoining shelves vary much in width. Passive margin or trailing edge coasts are located along passive continental margins. They are very stable (in particular, the shield coasts of Africa, Australia, and parts of eastern South America) and fed by large rivers often draining enormous basins, thus contributing large volumes of fine sediments to low-angle continental shelves. This in turn can lead to the formation of extended tidal flats, which can build an adequate substrate for the development of large plains of mangroves and salt marshes.
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2. DEPOSITIONAL ENVIRONMENT AND SUBSTRATE F ORMATION FOR THE D EVELOPMENT OF M ANGROVES AND S ALT M ARSHES There is an association between tectonic setting and the main depositional characteristics of the coastal zone (Table 1). Wave erosion is important on convergent margin coasts. Deposition, however, is predominant on passive margin coasts, which are supplied with abundant sediment from distal orogenic belts (e.g., from the Himalaya to the mangroves and marshes in the Sundarban); these belts may even be situated on the other side of the continent (e.g., the Amazonian mangroves receiving material from the Andes belt). Also, extensive barrier island formation with back-barrier wetland development can be associated with trailing edge coasts (e.g., Florida, USA). An extensive sedimentary coastal plain development can be also be found at marginal sea coasts, such as in the region of the delta of the Mekong River, which has a significant input of sediment from the distant Tibet highlands. The simple classification of coasts according to the predominant sediment type is a useful way to differentiate between features and processes characteristic of specific coastal environments. Muddy sedimentary coastal plains are a type of landform that is particularly relevant for the development of mangroves and salt marshes. Most of them form in low-lying, relatively sheltered areas with abundant sediment supply, such as behind barrier islands, around estuaries, bays, and deltaic coasts. In general, they are characterized by low wave energy climates and dominated by tidal process (Viles and Spencer, 1995). Muddy coasts can be subdivided into (1) open muddy coasts, (2) muddy coasts protected by barriers, and (3) muddy coasts along estuary and bay margins (Pethick, 1993). A prerequisite for their formation is the presence of a major river system discharging large amounts of suspended silts and clays. Enormous muddy coasts have been formed by the sediment discharge of the Ganges–Brahmaputra (1,670 106 t/year) and Amazon (1,200 106 t/year) river systems, which are two of the most sediment-rich rivers in the world (Milliman and Meade, 1983) and have given rise to vast coastal wetland systems. Although quite different area estimations exist, the Sundarban and the Amazonian mangroves are the two largest unitary mangrove systems worldwide, with about 10,000 km2 (Vanucci, 1999, 2002). High tidal range and a low wave energy environment also promote the formation of extensive mud deposits along low gradient coasts, such as in estuarine tidal flats (Perillo and Piccolo, 1999). The sediments carried to the coast by rivers tend to flocculate into larger aggregates when they encounter salt water. These particles will tend to settle in quiet coastal waters such as lagoons and sheltered estuaries. This mud is brought in by the incoming tide and is deposited before the tide reverses. It is possible to distinguish three main zones in mudflats that are relevant for the establishment and distribution of coastal vegetation: (1) the subtidal zone, which is submerged even at low tide, (2) the intertidal slope, a more steeply inclined though still gently sloping zone, and (3) the high-tide flat, which has a very gently sloping surface, some of which is partially submerged at high tide. In the
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subtidal zone of muddy coasts, there is usually no vegetation development. The intertidal zone is dissected by channels, and its slope is usually subjected to successive episodes of erosion and deposition that provide a substrate that can be insufficiently stable for large vegetation growth to occur. The high-tide flat and the sectors between channels or creeks in the intertidal zone encompass a continuum in sediment deposition, vertical accretion, and inundation frequencies, which enables a variety of halophytic plants to colonize the mud surface. Depending to a great extent on temperature, substrate elevation, and porewater salinity, a salt marsh and/ or a mangrove swamp may develop. Where salt marshes and mangroves occur, the latter are characteristically concentrated on the offshore edge of the high-tide flat where a rim with a slightly higher elevation is sometimes developed (Pethick, 1993). Mangroves trees are widely distributed throughout the tropical climates of world. Their branching root systems and twisted trunks and branches provide the same resistance to tidal currents as the smaller salt marsh plants do (Pethick, 1993). The mangrove root system acts as a highly efficient sediment trapping mechanism. It dissipates surface wave energy, reduces wave heights, and decelerates water flow, thus accelerating the process of land formation (Furukawa and Wolanski, 1996; Cahoon et al., 2002). Continued vertical accretion is limited by the reach of the highest tides, a nearly level equilibrium surface developing where the addition of organic debris is more or less balanced by compaction and decay of organic material. Mangroves attain full development on fine-grained, soft organic mud deposited in a number of different environmental settings, characterized by specific geomorphologic and hydrological features (Woodroffe, 1992). Nevertheless, a large number of species will grow in a variety of sedimentary environments. River-dominated mangrove forests develop in the deltas of large tropical rivers, for example, in the Bay of Bengal in the deltas of the Ganges and Brahmaputra rivers (Coleman, 1969). Tide-dominated mangroves are characteristic of areas of high tidal range where there is an extensive, low-gradient intertidal zone available for colonization (e.g., Amazonia). In wave-dominated settings, where there is an abundant supply of sand, mangrove-colonized coast may consist of shore-parallel sandy ridges, often barrier islands enclosing a series of elongate lagoons, as in southeastern Australia (Roy, 1984). Drowned bedrock valley or ria settings described from many large coastal embayments drowned by the postglacial rise in sea level may provide sheltered environments for mangrove development on muddy substrates, as in northwestern Australia (Semeniuk, 1985). Mangroves have developed also on sediments deposited in carbonate environments such as Florida Bay (Davis, 1940). Salt marshes are vegetated mudflats located higher, relative to mean tide level, than other parts of mudflats. Consequently, they are flooded much less frequently, and the tidal currents flowing over the marsh surfaces are much weaker. Closed salt marshes are often enclosed by spits and bays which, besides providing sheltered conditions for deposition, also restrict their horizontal growth. Open marshes develop horizontally and vertically on the banks of estuaries or on the shores of larger coastal embayments (Davies, 1980). The most distinctive features of any salt marsh are (1) its high-density creek system with branching tributaries and meandering channels. These systems are not primarily
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drainage channels because the water which floods and ebbs the marsh surface flows over the seaward edge of the marsh rather than through the creeks. (2) The small shallow pools called “salt pans” that cover many marsh surfaces. The primary salt pans originate on the initial marsh surface as vegetation begins to spread. Some small areas are left unvegetated, and the seawater in these embryo pools evaporates, producing highly saline conditions in which no plant life can survive. Channel salt pans are long sinuous pools that seem to originate from salt marsh creek abandonment. The old creek fills with sediments but leaves a surface depression that holds saline water, thus preventing vegetation from colonizing (Pethick, 1993).
3. INFLUENCE OF SEA LEVEL AND C LIMATE OSCILLATIONS ON L OCAL G EOBOTANICAL F EATURES In tropical coastal regions, there is frequently a continuum of topography, inundation regime, and water residence time that, according to geoforms and salt tolerance of the different plant species, can result in the formation of salt marshes or mangroves. Under a determined geomorphological and hydrodynamical setting, the key parameters in determining the existence of herbaceous or woody vegetation will be temperature and sea level, which will define the inundation frequency and the sediment porewater salinity (Lara and Cohen, 2006). In consequence, the current global distribution and the characteristics of salt marshes and mangroves reflect the evolution of environmental conditions, basically sedimentological and hydrological (fluvial and sea level), that followed the last glacial period. Worldwide, during the Last Glacial Maximum, about 18,000 years BP, the sea level was around 120 m lower than today (Shackleton and Opdyke, 1973). The eustatic sea level rise stabilized around 6,000 years BP, and during the Late Holocene the relative sea level trends have been governed by climate and isostatic and/or by tectonic adjustments. In spite of hydrological and morphological differences, a brief comparison among various estuaries and their stratigraphy reveals the strong influence of present and former relative sea levels on the fixing and expansion of mangroves and salt marshes. On the southwest Florida coast, mangroves occur on low-relief islands. This region has been subjected to a relative rise in Holocenic sea level, which has continuously decelerated to its present rate. The Holocene sediment package in this area consists of two sequences. The lower sequence is transgressive and was generated as coastal salt marsh and/or terrestrial environments were submerged and replaced by a shallow coastal marine setting. The upper sediment sequence consists primarily of (1) biogenic layers shallowing upward or (2) thickened mangrove peat layers, reflecting island emergence and shoreline stabilization, respectively. Based on coastal stratigraphy and 14C dates, the formation of this transgressive/regressive sediment sequence is directly related to changing rates of Holocene sea level rise, reported to have occurred between 3,500 and 3,200 years BP (Parkinson, 1989). The mangroves in the deltas of the Ganges–Brahmaputra and the Mekong River systems expanded with delta plain progradation during the late Holocene.
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The sediments were deposited in receiving basins of different tidal ranges of 4 m in the Meghna Estuary (India), 2.5–3.8 m in the South China Sea, and 0.5–1.0 m in the Gulf of Thailand. In the Mekong River Delta region, the maximum postglacial marine transgression appears to have occurred around 6,000–5,000 years BP, with relative sea level approximately 4.5 m above the present level (apsl). During the highstand and regressions of relative sea level over the last 4,550 years, delta progradation has produced a large plain of 62,520 km2. This extremely fast process could have been due to (1) very high sediment supply, (2) neotectonic movements and relative sea level changes, and (3) widespread mangrove forests playing an important role in enhancing sediment accumulation (Nguyen et al., 2000). The braided river channels of the Mekong River system have contributed to the development of a coastal plain deposit with elevations of 2.0–2.5 m apsl; mangrove and salt marsh occupy the lowlands of 0.5–1.0 m apsl. Similar processes occur in the Ganges–Brahmaputra Delta, where on descending from the Himalayan plateau to an upper delta plain, the rivers experience rapid lateral migration, producing a patchwork of flood plains of various ages. West of the river mouths, the lower delta plain is covered by a mangrove forest (Sunderban), drained by a network of river distributaries and secondary tidal channels, and formed in an earlier phase of Holocene delta progradation (Allison, 1998). Climate can also influence the stratigraphic structure of salt marshes and mangroves in different ways depending on the occurrence or absence of tropical storms. For example, in the Amazonian coastal wetlands, the absence of major storms results in a lack of clastic material in the sediment column. Besides that, salt marshes are relatively distant from the coastline and behind the mangroves. Peat is mostly absent in these sediments, but most probably due to a combination of species zonation, tidal export, and intense crab consumption of litter (Schories et al., 2003). In the case of the Mekong River Delta, salt marshes occur near the present coastline and behind the mangroves and receive substantial volumes of clastic sediment from the sea during storms (Nguyen et al., 2000). They have commonly originated as tidal flats that are sites of relatively quiet water deposition at spring high tide (Davis, 1996). These salt marshes are well drained and have a firm substrate of laminated clays, silts, fine sands, and scattered organic matter (OM), but peats are absent. Mangroves can produce peats which are commonly incorporated in the sedimentary deposits of the coastal area. The appearance of these peats is of great stratigraphic importance because it indicates the approximate position of relative sea level (Kamaludin, 1993). As stated before, given a relatively smooth coastal morphology and tropical temperatures, it will be mainly the relative sea level and to lesser extent the rainfall regime that determine whether mangroves or salt marshes are the dominant vegetation forms across the topographical gradient. These factors also determine their temporal evolution. An impressive example of this can be found in the extensive wetland system formed along a passive margin in the “Large Marine Ecosystem” (LME, 2002) extending from the Caribbean Sea boundary, just off of Venezuela, to the Paraiba River Estuary in Brazil. In north Brazil, the sector of the Amazonian coastal wetlands between Bele´m (Para´) and Sa˜o Luis (Maranha˜o) contains the world’s largest unitary mangrove system, with about 8,900 km2
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(Kjerfve and Lacerda, 1993). Therefore, the evolution and temporal alternation of mangroves and salt marshes in this region will be subsequently presented in more detail, embracing the main factors discussed in the first three sections of this chapter.
4. M AJOR FACTORS L EADING TO THE D EVELOPMENT OF SALT M ARSHES AND MANGROVES AT THE AMAZON C OASTAL R EGION : AN INTEGRATED ANALYSIS The Amazonian coastline is primarily dominated by areas of massive freshwater and sediments inputs, mainly from the Amazon and Tocantins rivers. This region is characterized by a wide shelf and experiences a macrotidal range (4–10 m) and upwellings along the shelf edge. The coast is part of a subsidence region and is therefore especially susceptible to sea-level oscillations. Tectonic control has affected the coastal geomorphology in terms of long-term and large-scale landscape development, and the tectonic framework of the Equatorial Atlantic margin controls the regional geomorphic characteristics of the Quaternary deposits (Souza Filho, 2000). Further, the Amazonian wetland development has been influenced by neotectonic movements, which favored basin formation in which continental muddy sediments in an estuarine environment accumulated. This caused the development of analogous geobotanical units along the coast, characterized by mangrove peninsulas, which include salt marshes and freshwater swamps. The coastal plain is mostly limited by 1–2 m high dead cliffs bounding a coastal plateau of Tertiary sediments and is limited to the north by a dune-beach ridge barrier subject to modern marine processes (Souza Filho and El-Robrini, 1996). The salt marshes in the higher central part of the peninsulas are flooded only by the highest spring tides and constitute a hypersaline habitat, at least during the dry season; with porewater salinities reaching about 90, they are dominated by halophytes such as Sporobolus virginicus and Sesuvium portulacastrum. In the rainy season, when porewater salinity is much lower, the Cyperaceae Eleocharis geniculata and Fimbristylis spadicea can dominate. This marsh area merges smoothly with a wide mangrove-colonized mudflat dissected by deep creeks, which extend for 3–6 km down to the mid-tide mark. The upper mudflats, covered by monospecific stands of Avicennia germinans, are inundated only during normal spring tides and the sediment has porewater salinities between 90 and 50. At about mid-tide level, a slope break occurs and an area of relatively steep, frequently flooded (233 days/year) mudflats with porewater salinity around 36 and mixed mangrove forest leads down to Rhizophora mangle-dominated sectors fringing the low-tide mark. Considering the past sea-level changes and the sedimentological conditions determining the current distribution of the different vegetation units, Cohen et al. (2005a) elaborated a model for the development and transition between mangrove and salt marshes in the region during the Holocene. The mangrove initiated its colonization around 6,500 years BP (Vedel et al., 2006) when the pre-existing fluvial valleys were flooded by a rapidly rising postglacial sea level.
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Its stabilization around 5,100 years BP probably triggered mangrove expansion in the region, overlapping basal sand with marine shell fragments. This stabilization agrees with the records of high sea level along the Brazilian, Suriname, and Venezuela coasts at 5,100 years BP (Suguio et al., 1985; Tomazelli, 1990; Angulo and Lessa, 1997; Rull et al., 1999) and around 6,130 years BP in Australia (Woodroffe and Grime, 1999). The rising sea level during the mid-Holocene was most likely close to the modern level, but with mangrove forests restricted to the topographically highest central part of the peninsulas, currently occupied by salt marshes and to landward channels, which probably developed mangrove fringes on Tertiary sediments. Mangroves were present probably without interruptions from 5,100 until 1,000 years BP in the current central marshes under relatively constant RSL (Behling et al., 2001). From 1,800 until 420 years BP, the mud progressively filled the estuarine basins of the Amazonian coastal region and the mangroves expanded, with a relative sea level probably not lower than 1 m below the current one. Around 420 years BP, a transition occurred at the highest sectors from an Avicennia-dominated mangrove forest to a Poaceae-dominated marsh and after 200 14C years BP to Cyperaceae under constantly low inundation frequency. Based on the habitat salinity zone, this vegetation change suggests a gradual fall of relative sea level or sediment accretion that gradually eliminated the mangroves in the higher sites. Besides the currently relatively well understood influence of relative sea level, the manifestations of past global climate events on coastal wetland development in the Amazon region is less known. However, there is preliminary evidence indicating that during the last 1,000 years, the mangroves and salt marshes in north Brazil registered two periods of low inundation frequency temporally correlated with the so-called Little Ice Age period: (1) between AD 1130 and 1510 and (2) between AD 1560 and the end of the 19th century (Cohen et al., 2005b). These two events may reflect a sea level regression and/or drier conditions with less rainfall and involve shifts in the boundaries of salt marshes and mangrove, favoring the development of hypersaline tidal flats and marshes. During the last decades, the Amazonian mangroves have been migrating to higher elevation zones occupied by marshes, suggesting a relative sea-level rise associated with the global tendency toward a eustatic sea-level rise, due to the increase in temperature and glaciers melting around the world during the last 150 years.
5. INFLUENCE OF G EOMORPHOLOGY AND I NUNDATION R EGIME OF G EOBOTANICAL U NITS ON THEIR S EDIMENT BIOGEOCHEMISTRY In coastal wetlands, biogeochemical processes and physicochemical parameters of the sedimentary substrate, such as porewater salinity, redox potential, pH, sulfide levels, and nutrient availability, are highly dependent on the frequency and duration of inundation and waterlogging. Under a given hydrographical
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setting, these features are mainly determined by the topography and geomorphology of the system, which, influencing the mineralogic and granulometric features of the sediment, further contribute to vegetation zonation (Woodroffe, 1992). In turn, the plants themselves affect biogeochemical processes in sediments through, for instance, aeration via roots, selective ion uptake, exclusion of salts or nutrients, leaching, or underground decomposition of roots. Salt marshes characteristically occur at higher elevations than mangroves and, being inundated by fewer tides (Saintilan and Williams, 1999), their substrate is in general drier and subject to a greater range of salinities and redox conditions (Clarke and Hannon, 1967). However, salt marshes in topographic depressions may be subject to lower inundation frequencies but can stay waterlogged for longer time periods as drainage is impeded. Coastal wetlands act as sediment traps, retaining part of the incoming load from tides and rivers. Sediment properties such as grain size and clay content influence the biogeochemical cycles through nutrient and ion exchange capacity. Lu et al. (2005) reported for marine surface sediments that nitrogen (N) content increased with the proportion of fine-grained substrate and also observed a shorter releasing time of exchangeable N in fine sediments as opposed to coarser substrates. In salt marsh sediments, total organic carbon, total nitrogen (TN), and total sulfur increased with increasing clay content (Zhou et al., 2007). To illustrate how geomorphology and the topographic and hydrological setting influence the nutrient dynamics of mangroves and salt marshes, we will discuss different situations for N and phosphorus (P) in (1) regularly inundated, (2) rarely inundated, and (3) waterlogged wetlands.
5.1. Regularly inundated wetlands Regular tidal flushing aerates the wetland substrate and removes stressors such as sulfides and salts (King et al., 1982; Howes et al., 1986; Feller et al., 2003a) from the surface layer. Increased water movement can also reduce the sulfide concentration by, for instance, adding iron (Fe) to the system which in turn can precipitate sulfides (King et al., 1982). Regularly inundated wetlands can be temporarily saturated with water, and exhibit anoxic conditions, but only where drainage at low tide is restricted, for example, in topographic depressions, do we find long lasting hypoxic conditions as described below (Section 5.3). 5.1.1. Nitrogen In most flooded wetland soils, the primary form of mineralized N is ammonium (Mitsch and Gosselink, 1993). Its extent of sorption onto sediment particles is associated with OM and clay content of the sediment. A positive correlation between OM content and ammonium adsorption capacity has been observed in several studies (Boatman and Murray, 1982; Tam and Wong, 1995; Raaphorst and Malschaert, 1996). However, most wetland sediments have a low ammonium absorption capacity. In Florida, this phenomenon is related to the low cation exchange capacity and low clay content of the carbonate soils predominant in the
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region (Rosenfeld, 1979). For sites with high clay and OM content, the decreased adsorption capacity for ammonium could be due to the blocking of ion exchange sites through humic substances or through competitive cations such as Fe (Boatman and Murray, 1982; Holmboe and Kristensen, 2002). Most pathways in the N cycle are closely linked to oxygen availability (Fisher and Acreman, 2004) and thus to sediment aeration. Thus, the geomorphological and hydrological setting, which condition the extent to which the sediments stay waterlogged or are flushed by the tides, will largely determine N transformation in the sediment. In this context, the most relevant process is nitrogen mineralization, that is, the transformation of organic N to ammonium nitrogen during OM degradation, which occurs under both aerobic and anaerobic conditions. However, the oxidation of ammonium to nitrate or nitrification is restricted to the thin oxidized layers on the sediment surface, in the rhizosphere of the plants, or the lining of crab burrows. Denitrification (the reduction of nitrate to N2O or N2) is a significant pathway through which N is lost from salt marshes and mangroves (Alongi et al., 1992; Mitsch and Gosselink, 1993). It only occurs under anaerobic conditions, but is limited through the aerobic nitrification process. On the other hand, N2 fixation, the conversion of N2 gas to organic N, has been shown to be a significant N source to wetlands (Holguin et al., 1992; White and Howes, 1994; Pelegri and Twilley, 1998; Burke et al., 2003; Ravikumar et al., 2004). Some studies report a reduction of N2 fixation under hypoxic conditions (Ogan, 1990; Kreibich and Kern, 2003), but others report that plants adapt quickly to waterlogged conditions in terms of N2 fixation of the associated bacterial assemblages (Pugh et al., 1995). Salinity however has been reported in several studies to reduce N2 fixation through inhibition of nodulation and/or inhibition of nitrogenase activity (Bekki et al., 1987; Soussi et al., 1999; Serraz et al., 2001). 5.1.2. Phosphorus Reducing soil conditions of regularly flooded areas promote Fe3þ reduction, dissolution of ferric phosphates (e.g., strengite), release of Fe2þ, reduction of sulfate to potentially Fe-binding sulfides, and desorption of P bound to ferric oxyhydroxides (Mortimer, 1971; Lindsay and Vlek, 1977). Increases of available P and decreases of total P (TP) in sediment are generally accompanied by these effects. Similar changes in Fe3þ/Fe2þ ratio and available P concentration occur with decreasing redox potential (Eh) along inundation gradients and increasing sediment depth, producing three-dimensional distribution patterns of biogeochemical parameters. Accordingly, Eh decreased along a salinity gradient from freshwater marsh to brackish and salt marsh in surface sediments from the Cooper River, South Carolina (Sundareshwar and Morris, 1999), while Fe3þ declined in surface sediments of these marshes from freshwater marsh to brackish marsh and was below detection limit in the salt marsh (Paludan and Morris, 1999). These data were consistent with the highest relative degree of free sorption sites on metal particles into the fresh water marsh, decreasing P sorption, decreasing TP, increasing porewater-dissolved reactive P, and increasing Ca-bound P from freshwater marsh to salt marsh (Sundareshwar and Morris, 1999).
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While sorption of P by Fe depends mainly on Eh; pH and salinity are significant factors for changes in P sorption by Fe, Al, and Ca in flooded regions, besides the impact of the metal concentration by itself (Paludan and Morris, 1999; AndrieuxLoyer and Aminot, 2001), the Fe:P ratio as a measure of free sorption sites for phosphate on FeO(OH) (Jensen et al., 1992), and in the case of Fe, the concentration of reactive Fe that can interact with P (Raiswell and Canfield, 1998). Salinity and pH affect also the P sorption by OM (Koch et al., 2001), humic acids, and metal-humic acid complexes (Gerke and Hermann, 1992; Paludan and Morris, 1999; Morse et al., 2004). Gooch (1968) hypothesized that a seasonal cycle of sorption and release of inorganic P in a natural salt marsh in Delaware is controlled by the adsorption of P on ferric hydroxide and precipitation of ferric phosphate during the winter, and release of P in spring, mediated by the changes in sulfide concentration and pH, which therefore may be responsible for the summer phosphorus eutrophication in this region. P sorption in carbonate systems is less affected by Eh, since relatively more inorganic P is bound to calcium carbonate than to redox-sensitive compounds of metals like Fe and Mn or is fixed in minerals like apatite and octacalciumphosphate (Moore and Reddy, 1994; Feller et al., 2003a). Although waterlogged sediments are known to release P into overlying water (Mortimer, 1971) and the amount of available P typically increases with decreasing Eh, Feller et al. (2003a) and McKee et al. (2002) reported low porewater-soluble reactive P concentrations, irrespective of the Eh and waterlogging. Data from Hinchinbrook Island in Australia (Boto and Wellington, 1984) showed a strong biomass-Eh correlation, may be as a result of oxygen translocation by the plants to the root zone. As it is common in mangrove areas, aboveground biomass was highest at the low elevated fringe sites near channels, where TP and bioavailable P were high. However, the redox potential at the fringe sites was also high, may be additionally though aeration by tidal flushing, and decreased with increasing elevation. Thus, probably a major reason for the higher bioavailable P values at the lower sites is the higher degree of tidally influenced sediment exchange, compared to higher dwarf forests. Reduced flow in areas of dense vegetation results in excess of sediment deposited from incoming tidal water. High pH values due to flooding with seawater, increasing pH caused by reduction, and reductive dissolution of ferric Fe compounds generally enhance desorption and dissolution of P in flooded soils. Therefore, sediment exchange may represent a major source of P enrichment in topographically low areas with dense vegetation (Boto and Wellington, 1983). Accordingly, Florida Bay is supposed to be the source of dissolved inorganic nutrients for southern Everglades mangroves, while freshwater inputs from the Everglades marshes are an important source of dissolved organic matter (DOM) in these wetlands (Davis et al., 2003; Chambers and Pederson, 2006). The Gulf of Mexico seems to be the major source of the P enrichment in Florida Bay (Fourqurean et al., 1992a; Fourqurean et al., 1992b), and the TP concentration decreases in mangrove forests of South Florida with increasing distance from the Gulf of Mexico (Chen and Twilley, 1999). Amounts of bioavailable P and TP in surface sediments of two tidal freshwater marshes in Virginia appeared to be directly linked to sediment accumulation with higher P contents at the topographically lower marsh, where sediment inputs were highest (Morse et al., 2004).
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5.2. Rarely inundated wetlands In wetlands, the frequently inundated and/or waterlogged sectors usually exhibit lower salinity, whereas salt stress increases in the topographically higher, rarely inundated, and thus less anoxic regions. In these areas (usually characterized by salt marshes or dwarf mangroves), the main stress for plants and influence on sedimentary processes is the increasing salinity. 5.2.1. Nitrogen Nitrate reductase (NR), the first enzyme in the nitrate assimilation pathway, is a limiting factor of plant growth and development (Solomonson and Barber, 1990) and is influenced by a variety of environmental factors (Crawford, 1995). In many plants, NR activity decreases under salt stress (Parida and Das, 2004), specifically due to the presence of chloride salts in the external medium. Chloride ion seems to produce a reduction in nitrate uptake and consequently a lower nitrate concentration in the leaves, although a direct effect of chloride itself on the activity of the enzyme cannot be discarded (Flores et al., 2000). Additionally, salt can affect growth of salt marsh and mangrove plants through the action of sodium as a competitive inhibitor of ammonium uptake (Odum, 1988; Bradley and Morris, 1991; Feller et al., 2003a). TN levels in surface sediments can be elevated in inland marshes as compared to the seaward mangrove (Boon and Cain, 1988; McKee et al., 2002; Schmitt, 2006) or of the upland ecotone (Traut, 2005). This could be due to higher N inputs through N2-Fixation or mineralization to lower losses (e.g., less litter export or less denitrification), and/or to less plant uptake due to low requirements or slow growth. Clarke (1985) found a difference in TN allocation in wetland ecosystems, indicating that the plant component of the TN pool is large (55%) in the mangrove zone, whereas in the salt marsh the plant pools are small (15%) in comparison with the sediment substrate where they grow. It is, however, not always the total amount of the nutrient that is important, but the effects that hydrological conditions have on pathways such as primary production, availability, uptake, and decomposition. 5.2.2. Phosphorus In contrast to the observed P enrichment by likely sediment deposition at the mangrove fringe, P pools as well as accumulation rates seem to be higher in soils of irregularly flooded marshes compared to regularly flooded marshes (Craft et al., 1988; Sundareshwar and Morris, 1999). At low pH values, Fe and Al hydroxides carry a net positive charge. Hence, low soil pH values promote P sorption, while at higher pH values the metal hydroxides are negatively charged (Stumm and Morgan, 1981) and therefore lower the adsorption capacity in marsh environments with higher salinity. Additionally, higher concentrations of anions like chloride and sulfate reduce the isoelectric point of metal hydroxide particles (Stumm and Morgan, 1981) and compete with phosphate for available sorption sites. The geomorphologically related changes in P sorption may result in a switch from P limitation in higher elevated areas to N limitation in topographically lower,
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temporary inundated regions. Accordingly, primary production was found to be N limited in salt marshes (Valiela and Teal, 1979) and could be probably P limited in fresh water marshes (Paludan and Morris, 1999; Sundareshwar and Morris, 1999). Although marshes were generally considered to be N-limited by other authors (Valiela et al., 1973; Patrick and Delaune, 1976; Bowden, 1984; Kiehl et al., 1997; Bedford et al., 1999), irrespective of the elevation, these findings contributed to the vision of a shift from P limitation in freshwater environments to N limitation in coastal and estuarine ecosystems. Fertilization experiments have shown that N limitation at fringe sites may shift to P limitation in dwarf forests (Boto and Wellington, 1983; Feller et al., 2003a). While other studies carried out in mangrove areas agreed (Feller et al., 2003b) or disagreed (Feller, 1995; Koch and Snedaker, 1997; Naidoo, 2006) with the generally accepted paradigm of N limitation in costal and estuarine environments, the question of P or N limitation in mangrove regions remains unclear. McKee et al. (2002) hypothesized that a switching in nutrient limitation observed in Belize reflected the spatial changes of external nutrient supply and environmental stress factors, while the latter caused changes of internal nutrient demand. R. mangle-dominated dwarf forests in Belize had strongly reduced soil and were found to be P limited, while A. germinans-dominated dwarf forest in Florida were hypersaline and N-limited. Flooding-related stress may increase plant demand for P, whereas salinity stress may increase demand for N (Feller et al., 2003b). The different effects of nutrient enrichment, observed in fertilization studies, are probably a result of diverse biotic and abiotic interactions in geomorphologically and sedimentologically dissimilar environments. Additional stressors may be responsible for the heterogenous results of fertilization experiments in mangrove areas in comparison to studies in freshwater wetland ecosystems. A literature survey of 40 fertilization studies (Koerselman and Meuleman, 1996) has investigated the elemental N:P ratio of the vegetation as a reliable prediction tool for assessing the nature of community nutrient limitation in a variety of European freshwater wetlands. The authors concluded that N:P ratios <14 predicted N limitation, N:P ratios >16 predicted P limitation, and wetlands with N:P ratios between 14 and 16 were colimited by N and P. Lockaby and Walbridge (1998) found a maximum litterfall production at N:P ratios in litterfall of ca.12 in forested wetlands of the southeastern USA. A review and analysis of data concerning the nature of nutrient limitation in temperate wetland types in North America described the N:P ratios of surface soils in marshes and swamps to be lower than in bogs and fens (Bedford et al., 1999). As N:P ratios in sediments from mangrove forests of the Saigon River Delta were highly negatively correlated with pH and N:P ratios of leaves reflected this relation, it is reasonable to assume that a shifting in limitation could be partly attributed to a shift in sediment pH (Oxmann, 2007). The major part of studies concerning P dynamics in wetlands attributed enrichment, availability, limitation, and turnover rates of phosphorus to adsorption effects, reflecting their importance in P distribution and composition, especially in wetland ecosystems (see also Hesse, 1962; Alongi et al., 1992). Analysis of data from 57 wetlands has shown that the binding capacity of Fe and Al was a major factor in the retention of P (Fisher and Acreman, 2004). It is widely confirmed that
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adsorption dominates P retention at low concentrations, whereas phosphate minerals control solubility of P at high concentrations (Lindsay, 1979). While the optimum pH values for P sorption and mineralization generally differ and the sediment properties vary from sulfate acidic to calcareous, it is likely that the ratio between adsorbed P and mineral P is mainly controlled by the pH in tidal swamps and marshes. However, there is also a number of potential interactions between precipitation/dissolution and adsorption/desorption reactions.
5.3. Waterlogged wetlands Hypoxia (i.e., low redox potentials) is the prevailing condition influencing plant life and nutrient dynamics in waterlogged soils. Low redox potentials have a negative effect on nitrogen uptake in higher plants (Lindsay and Vlek, 1977; King et al., 1982; Bradley and Morris, 1991; Saenger, 2002). 5.3.1. Nitrogen In coastal wetlands with abundant sulfate from marine water, hypoxia leads to sulfide accumulation through bacterial sulfate reduction. Sulfide can have direct toxic effects on plants but also hampers N uptake, which decreases with increasing sulfide concentration and is reflected, for instance, in a decreased leaf elongation, slow growth, or even plant die-back (Mendelssohn et al., 1981; Mendelssohn and McKee, 1988; Bradley and Morris, 1990; Koch et al., 1990). Puiatti and Sodek (1999) found a negative impact of waterlogging on N transport in the xylem sap of plants. Sulfide concentrations can be reduced by Fe precipitation to Fe sulfides such as pyrite and mackinawite (King et al., 1982). This reaction in turn is dependent on the available Fe in the soils and/or new Fe input through water movement, which is often restricted in these systems. As mentioned above, many processes in the nitrogen cycle are closely linked to oxygen availability. Nitrification is restricted to oxidized layers, whereas denitrification occurs under anaerobic conditions only but is limited through the aerobic nitrification process. Concerning N2 fixation, some studies report a reduction under hypoxic conditions (Ogan, 1990; Kreibich and Kern, 2003) but others state that plants adapt quickly to waterlogged conditions in terms of the N2 fixation of the associated bacterial assemblages (Pugh et al., 1995). Binkley et al. (2003) reported that N2 fixation is often limited by P availability (see below). 5.3.2. Phosphorus As pointed out above, P cycles in flooded soils are affected by several changes in physicochemical conditions. Under microbially mediated reductive conditions, the dissolution mechanisms concerning P were summarized by Hutchison and Hesterberg (2004) as (1) reductive dissolution of Fe(III) minerals with associated phosphate, (2) competitive adsorption of DOM and phosphate by ligand exchange on mineral surfaces, (3) DOM-enhanced dissolution of surface Fe or Al with concomitant release of phosphate (PO4), (4) formation of aqueous ternary DOM–Fe–PO4 or DOM–Al–PO4 complexes, and (5) decreased phosphate
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sorption with increasing pH. Additionally, organic anions from OM biodegradation are released during flooding (Appelt et al., 1975; Sposito and Page, 1985; Bauld, 1986) and compete with phosphate ions for free sorption sites or reduce P sorption by complexation with Fe and Al. Oxalate was often found to be the organic acid with predominant P mobilization efficiency (Stro¨m et al., 2005). In relation to the high production of calcium oxalate in various wetland plants, chelation by oxalate could alter phosphate sorption, whereas the concentrations in sediments seem to be linked to carbon content and pH (Oxmann, 2007). Wright et al. (2001) hypothesized that under anaerobic conditions, several factors could be important for the release of labile P, such as the lysis of aerobic microorganisms, the hydrolysis of stored polyphosphates by facultative anaerobes, and subsequent P release, as well as the decrease in biological P demand. Especially for waterlogged environments, the low crystallinity of Fe compounds seems to be the reason for increased P sorption capacities following submergence, while precipitation of Fe(II) compounds (e.g., vivianite) may occur at high Fe(II) concentrations. Long-term flooding may increase the proportion of amorphous Fe oxides (Darke and Walbridge, 2000) or mixed Fe(II)Fe(III) hydroxy compounds (Ponnamperuma et al., 1967; Khalid et al., 1977; Cornell and Schwertmann, 1996) and readsorb previously released P. Several studies found increased P sorption after flooding, related to amorphous and poorly crystalline oxides and hydroxides of Fe (Krairapanond et al., 1993; Zhang et al., 2003). While the process of transformations between Fe compounds during reduction is not explored in detail, it can be suggested that amorphous oxides with higher P adsorption capacity will form after dissolution of crystalline Fe. The increase in available P could be seen as a consequence of these transformations. However, a number of publications also report a decrease in available P caused by submergence (Kuo and Mikkelsen, 1979; Sah and Mikkelsen, 1986; Sah et al., 1989a,b; Zhang et al., 1993). Permanently flooded dwarf forests have shown lower redox potentials than those in taller fringing forests (McKee et al., 2002; Feller et al., 2003a), but the authors found low porewater concentrations of soluble reactive P with little variation across the transect, and the ratio of available N : P increased from fringe to dwarf in controls without P or N fertilization.
5.4. Salt marshes and mangroves: Nutrient sources or sinks? The question of whether coastal wetlands are sources or sinks of nutrients has been discussed extensively in the literature, but no agreement has been reached so far (Mitsch and Gosselink, 1993; Lee, 1995). Whereas North American studies in the 1960s postulated an “outwelling theory” for salt marshes on the Atlantic coast of North America, studies in European temperate marshes could not always confirm this theory (for a review see Boorman, 1999). It may not be possible to answer this question generally for all wetlands. A wetland can be a sink for one nutrient and a source of another, or even a sink for the inorganic form of a nutrient and source for the organic form of the same nutrient (Mitsch and Gosselink, 1993; Lee, 1995). The direction and strength of nutrient flows might change throughout the year with the seasons in temperate zones where distinct growing seasons occur
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(Mitsch and Gosselink, 1993) or in the wet tropics where the monsoonal rains have a deep impact on the ecosystems (Alongi et al., 1992). The disagreement on this question of sources and sinks is probably also a consequence of the different settings of the studied ecosystems. Salt marshes and mangroves will have varying influence on the adjacent systems depending on their degree of isolation of the neighboring water body caused by geomorphological barriers and the subsequent hydrological settings and internal biogeochemical processes. Material export is partly a function of tidal range, that is, elevation and inundation regime influence the proportion of particulate and DOM that is exported from the coastal wetlands (Twilley, 1985; Boorman, 1999; Osgood, 2000). The spatial configuration and the relative age of a tidal marsh are further factors influencing material exchange processes. Older marshes that are not extending are expected to export organic and inorganic nutrients, whereas younger, actively extending systems are net importers of sediments and nutrients (Odum, 1969; Childers, 1994; Boorman, 1999). Import and export, that is, gain and loss, can also be achieved by other routes than the tides, such as N2 fixation (gain) or denitrification (loss) as examples of processes within the nitrogen cycle. However, as previously discussed, these processes are also dependent on sedimentological and hydrological parameters. Another factor influencing the import/export dynamics of coastal wetlands is the benthic fauna such as crabs. Several studies suggested that crabs could remove and process up to 80% of leaf litter in managroves, either by direct consumption or by storage, before it is exported by the next incoming tide (Robertson, 1986; Robertson and Daniel, 1989; Schories et al., 2003; Nordhaus et al., 2006). The distribution of these crabs is often related to sediment softness and presence of channels (Wessels, 1999) and is consequently linked to the geomorphology, topography, and hydrology of the ecosystems. Thus, depending on these settings, mangroves and salt marshes can be net exporters of OM as has been shown, for example, for a mangrove ecosystem in north Brazil (Dittmar and Lara, 2001; Dittmar et al., 2001) or net consumers of N and P as reported by Wosten et al. (2003) for mangroves in north Vietnam.
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C H A P T E R
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P ARADIGM S HIFTS IN M ANGROVE B IOLOGY Daniel M. Alongi
Contents 1. Introduction 2. Shifts in Established Paradigms 2.1. Rates of mangrove net primary productivity rival those of other tropical forests 2.2. Mangrove forests appear to be architecturally simple, but factors regulating succession and zonation are complex 2.3. Mangrove tree growth is not constant but related to climate patterns 2.4. Tree diversity is low, but faunal and microbial diversity can be high 2.5. Arboreal communities are important in food webs, exhibiting predatory, symbiotic, and mutualistic relations 2.6. Plant–Microbe–Soil Relations are tightly linked and help conserve scarce nutrients 2.7. Crabs are keystone species influencing function and structure in many, but not all, mangrove forests 2.8. Algae, not just detritus, are a significant food resource 2.9. Mangroves are an important link to fisheries 2.10. Mangroves are chemically diverse and a good source of natural products 3. Conclusions References
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1. INTRODUCTION Mangroves live at the interface between land and sea in subtropical and tropical latitudes. These woody plants develop as forests and grow best where low wave energy and shelter enable the deposition of fine sediments and subsequent establishment of mangrove propagules. The persistence of warm temperatures is of paramount importance for the existence of mangroves, with roughly 70 species currently occupying a total estimated area of 150,000 km2 in low latitudes (down from a global area of 198,000 km2 in 1980; Wilkie and Fortuna, 2003). Mangrove forests can attain immense biomass and height, rivaling the size of tropical rainforests; the standing crop of mangroves is ordinarily greater than other aquatic ecosystems, with equatorial mangrove forests often reaching a dry weight biomass on the order of 300–500 metric tons per hectare (Alongi, 2002). Coastal Wetlands: An Integrated Ecosystem Approach
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A number of morphological and ecophysiological adaptations of mangrove trees make them unique, including viviparous embryos, aerial roots, and physiological mechanisms to cope with salt and to maintain water and carbon balance. Living between land and sea, it is not surprising that both terrestrial and aquatic species colonize and live in mangrove trees, soils, and waters nor it is surprising that these tidal forests are a valuable economic and ecological resource. Mangroves are important nursery grounds and breeding sites for fish, birds, mammals, crustaceans, reptiles, and shellfish and are a renewable source of wood and food for many indigenous settlements. They are also accumulation sites for sediment, nutrients, and other elements, including contaminants, and offer some protection against coastal erosion. The objective of this chapter is to critically examine recent research trends in mangrove biology. The focus here is on some key topics that deviate from established paradigms of the biology and ecology of mangrove forests and their associated food webs. I take this approach simply because the biology of such complex ecosystems cannot be reviewed adequately here, and there exist several recent fine books and comprehensive reviews of mangrove biology and ecology (Ellison and Farnsworth, 2001; Kathiresan and Bingham, 2001; Lacerda, 2002; Saenger, 2002; Kathiresan and Qasim, 2005; Hogarth, 2007).
2. SHIFTS IN ESTABLISHED P ARADIGMS 2.1. Rates of mangrove net primary productivity rival those of other tropical forests Most initial productivity studies (Golley et al., 1962; see review of Clough, 1992) indicated rates of net mangrove primary production equivalent to salt marshes but less than tropical terrestrial forests and, at best, equal to carbon fixation rates measured in other marine ecosystems (Bunt, 1975). Improvements in technology and methodology, especially over the past decade, have allowed better estimates of tree photosynthesis and respiration and thus more accurate estimates of net primary production of mangrove forests. Not all mangrove forests are productive, as often exemplified by stunted trees at the landward edge or in saline arid regions (Cheeseman, 1994). Estimates of forest net primary production have been problematical, with order-of-magnitude differences in rates based on different methods. Nevertheless, these more recent data indicate that rates of mangrove net primary productivity are greater than previously believed, often rivaling rates of net primary production of tropical terrestrial forests (Komiyama et al., 2008). This is arguably the most critical paradigm shift in mangrove ecology considering that carbon fixed by the trees forms the basis for associated food webs and energy flow. Five methods are most commonly used to measure net primary production of mangrove forests: 1. Litterfall and incremental growth 2. Harvest
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3. Gas exchange 4. Light attenuation/gas exchange 5. Demographic/allometric. Litterfall is by far the most common method because it is inexpensive and easy to measure. However, it only measures leaf production and not growth of the remainder of the tree. Harvesting is labor intensive and slow, and such data are usually available only from plantation harvests and like litterfall measurements, account for only aboveground production; most leaf production is unaccounted for. Gas exchange is precise and rapid although subject to the problem (as all methods are) of extrapolating from a small area (usually a few individual trees) to an entire stand, injecting the problem of error. Moreover, relying solely on gas exchange measurements overestimates net production as it does not account for most tree respiration. Combining measurements offers the best hope of accounting for production of all, or most, tree parts. Measuring litterfall and incremental growth of the trunk accounts for all aboveground production, but not belowground production. Arguably, one of the best methods currently available is to measure light attenuation. The early efforts (Bunt et al., 1979) provided rapid and relatively easy estimates of potential net primary production but suffered from lack of actual leaf photosynthesis measurements and a number of untested assumptions based on light attenuation models from temperate forests and of the conversion of light absorbed to increases in mass over time. The method relies on relating the amount of light absorbed by the mangrove canopy to the total canopy chlorophyll content. The light attenuation method was subsequently modified, combining measurement of light attenuation with a more robust method of calculation of photon flux density at the bottom of the canopy and empirical measurements of leaf photosynthesis (Gong et al., 1991, 1992; Clough, 1997; Clough et al., 1997). This modified method still relies on measurements of light absorption by the forest canopy to estimate leaf area index, which is the amount of leaf area relative to the amount of ground area. The leaf area index (L) is then used to estimate net canopy photosynthesis (PN) using the formula, PN = A d L, where d is the daylength (h) and A is the average rate of photosynthesis per unit leaf area, which is obtained by measurement of CO2 exchange. Using this modification, Clough et al. (1997) compared their more robust estimates with the earlier light attenuation method. The comparison (Table 1) shows that the original method underestimates the true net production by a factor of 12. This suggests that many measurements using the Bunt et al. (1979) method need to be corrected by this factor, clearly increasing our estimate of mangrove net primary production. Comparing the gas exchange, litterfall, and increment growth plus litterfall, the original light attenuation and the modified light attenuation methods illustrate the methodological differences (Table 2) and the difficulty in our ability to settle on an accurate range of net primary production values for mangrove forests. It is clear that litterfall underestimates, and gas exchange measurements alone overestimate, net primary production. But from the remaining data, the modified light attenuation method currently appears to give the most reasonable estimate of total production while litterfall plus incremental growth give reasonable estimates of aboveground
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Table 1 Comparison of net primary production (t DW/ha/year) derived from the method described in Bunt et al. (1979) and the modified procedure described in Clough (1997) and Clough et al. (1997) Measurement set 1 2 3 4 Mean
Light attenuation method
Modified procedure
11.0 13.0 13.7 14.3 13.0 + 1.4
135 161 165 157 155 + 13
Measurements were made in a 22-year-old R. apiculata forest in peninsular Malaysia.
Table 2 Estimates of net primary production (t DW/ha/year) of Rhizophora apiculata forests of various age in peninsular Malaysia using five of the most used procedures Age
5 10 20 85
Gas exchange 132 122 240
Litterfall
7 10 10 8
Litterfall þ incremental growth
Light attenuation
19 34 30
14 19 16 21
Modified light attenuation 37 65 102
Source: Data from Gong et al. (1984, 1992), Ong et al. (1995), Clough et al. (1997) and Alongi et al. (2004).
production. The modified light attenuation method is most reasonable because it measures total net fixed carbon production and offers the most robust assumptions based on tree physiology and carbon balance. A number of recent studies have attempted to measure aboveground net primary production using allometry coupled with litterfall or leaf turnover (Coulter et al., 2001; Ross et al., 2001; Sherman et al., 2003). The method employed by Ross et al. (2001) is an adaptation of methods used for grasslands, incorporating detailed allometric measurements of individual trees coupled with observations of leaf demography to measure leaf turnover. Although they were not able to compare their method directly with other procedures, their net production values were at the upper end of the range for similarly sized forests. Coulter et al. (2001) similarly employed analysis of leaf nodes to produce an estimate of new leaf production. Combining leaf production with estimates of the number of inflorescence scars produced by the shedding of leaves and reproductive structures, aboveground production of Kandelia candel in Vietnam was estimated at rates comparable to, or greater than, previous values. Although caution must be applied when considering net primary productivity estimates based on a variety of methods used in disparate settings in forests of different age and living under different environmental conditions, the available data (Table 3) suggest that rates of net primary production are generally rapid compared to other marine primary producers (Gattuso et al., 1998).
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Table 3 Estimates of aboveground net primary production (NPP = t DW/ha/year) of mangrove forests in various parts of the world based on different methods Species
Location
NPP
Method
Rhizophora mangle, Avicennia germinans, Laguncularia racemosa R. mangle, A. germinans, L. racemosa A. germinans R. mangle R. mangle, A. germinans, L. racemosa R. mangle, A. germinans, L. racemosa R. apiculata Ceriops decandra R. apiculata R. apiculata (70 years) R. apiculata (18 years) R. apiculata (5 years) R. mangle (5 years) A. germinans L. racemosa Sonneratia apetala S. caseolaris A. officinalis A. marina A. alba B. gymnorrhiza B. sexangula E. agallocha X. moluccensis Mixed species R. apiculata, B. gymnorrhiza R. apiculata R. apiculata R. apiculata R. apiculata R. apiculata R. apiculata R. racemosa A. africana R. racemosa R. mucronata R. apiculata A. marina Bruguiera sexangula Kandelia candel Kandelia candel R. mucronata R. apiculata
USA
46.0
Gas exchange
USA
Demographic/allometric
USA USA USA
26.1 (fringe) 8.1 (dwarf) 20.5 16.9 22.5
Gas exchange Gas exchange Gas exchange
Puerto Rico
58.4
Gas exchange
Thailand Thailand Malaysia Malaysia Malaysia Malaysia Cuba Cuba Cuba Bangladesh Bangladesh Bangladesh Bangladesh Bangladesh Bangladesh Bangladesh Bangladesh Bangladesh Micronesia Malaysia
63.7a (13.1) 48.7a (9.7) 112.1a 102.2a (24.6) 65.7a (14.7) 36.5a (12.8) 1.6b 5.9b 5.4b 12.5b 26.4b 7.6b 4.4b 2.1b 0.6b 0.1b 4.7b 0.5b 4.2b 8.7b
Light attenuation Light attenuation Light attenuation Light attenuation Light attenuation Light attenuation Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental Harvest/incremental
Vietnam Vietnam Thailand Thailand Vietnam Vietnam Gambia Gambia Gambia India India India China China China Indonesia Thailand
4.9b 19.0 15.7 10.6 9.4 18.7 18.8 11.6 10.4 14.6 13.6 6.2 11.0 13.3 24.4 23.4 13.5
Harvest/incremental growth Incremental growth Incremental growth Incremental growth Litterfall Litterfall Litterfall Litterfall Litterfall Litterfall Litterfall Litterfall Litterfall Litterfall Litterfall/allometric Litterfall/incremental growth Light attenuation
growth growth growth growth growth growth growth growth growth growth growth growth growth growth
(Continued )
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Daniel M. Alongi
(Continued )
Species
Location
NPP
Method
Aegiceras corniculatum K. candel K. candel R. stylosa A. marina Mixed R.mangle, A. germinans, L. racemosa Mixed R.mangle, A. germinans, L. racemosa R. mangle Mixed Rhizophora spp. R. mucronata/A. marina R. apiculata, B. parviflora
China Vietnam Vietnam Australia Australia Dominican Republic
11.3 5.3 13.4 40.5a (9.6) 30.6a (6.4) 19.7c
Litterfall Demographic/allometric Demographic/allometric Light attenuation Light attenuation Demographic/allometric
Guadeloupe
21.2 (fringe) 6.2 (dwarf)
Litterfall/incremental growth
Hawaii Australia Sri Lanka Papua New Guinea Papua New Guinea Papua New Guinea Indonesia Indonesia Indonesia Indonesia Indonesia Indonesia Indonesia Okinawa
29.1 29.2 11.0 30.5a (9.7)
Litterfall/incremental growth Light attentuation Litterfall/incremental growth Light attenuation
30.1a (9.9)
Light attenuation
24.4a (6.8)
Light attenuation
104.6 96.9 103.2 106.1 109.4 63.7 74.3 36.2
Light attenuation Light attenuation Light attenuation Light attenuation Light attenuation Light attenuation Light attenuation Gas exchange
Nypa fruticans A. marina, Sonneratia lanceolata R. apiculata, A. marina R. apiculata, A. marina A. officinalis, A. marina C. tagal, R. apiculata C. tagal, R. apiculata R. stylosa, S. alba R. apiculata, K. candel Kandelia candel a
Estimate using the modified light interception method or original data recalculated using the modified method (see text). Estimates based on the original light interception method are in parentheses. Assumes an average density of 0.9332 t m3 (Saenger, 2002). c Sherman et al. (2003). d A correction factor of 4.8 was applied based on the data in Table 2.7 and all data calculated with both light inception methods (all those in above table asteriskeda). Of n = 11 forests, the original method gave a mean NPP estimate of 11.85 t DW/ha/yr and the modified method gave a mean NPP value of 57.08 t DW/ha/year, for an average difference of 4.8. All C values were converted to DW assuming that mangrove wood is 47%C (Saenger, 2002). Source: Data from Golley et al. (1962), Miller (1972), Hicks and Burns (1975), Lugo et al. (1975), Christensen (1978), Twilley (1985), Putz and Chan (1986), Aksornkoae et al. (1989), Lee (1990), Atmadja and Soerojo (1991), Gong et al. (1991, 1992), Robertson et al. (1991), Amarasinghe and Balasubramaniam (1992), Sukardjo and Yamada (1992), Sukardjo (1995), Day et al. (1996), Ong et al. (1995), Clough et al. (1997, 1999), Clough (1998), Cox and Allen (1999), Alongi and Dixon (2000), Alongi et al. (2000, 2004), Kathiresan (2000), Coulter et al. (2001), Ross et al. (2001), Sherman et al. (2003), Suwa et al. (2006). b
If we accept the data obtained using the modified light attenuation method as the best estimate of net primary productivity of mangroves, the average rate of net primary production is 64 t DW/ha/year. In comparison, the estimates based on incremental growth plus litterfall average to 11 t DW/ha/year. There is considerable range between and within both sets of values, but they both suggest that mangroves are more significant carbon fixers in the tropics than previously thought. Moreover, plotting the data versus latitude (Figure 1) gives a significant negative relationship, indicating that mangrove net primary production declines away from
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Net primary production (t DW/ha/year)
120 NPP = 100.022 – 4.145x R 2 = 0.624, P < 0.001, n = 16
100 80 60 40 20 0 0
5
10
15
20
25
Latitude (degrees N and S)
Figure 1 Latitudinal changes in net mangrove primary production measured using a modified light interception method. Data from Gong et al. (1991, 1992), Atmadja and Soerojo (1991), Robertson et al. (1991), Sukardjo (1995), Clough et al. (1997), Clough (1998), Alongi and Dixon (2000), and Alongi et al. (2000, 2004).
Tropical terrestrial (n = 70)
Mangroves (n = 50)
0
5
20 25 10 15 CO2 leaf assimilation rate (mmol/m2/s)
30
Figure 2 Comparison in CO2 leaf assimilation rates between various species of mature tropical mangrove and terrestrial trees. Mangrove data from sources listed inTable 3 (plus older references within). Tropical terrestrial tree data from references cited in Figure 2.29 in Turner (2001) plus data in Marenco et al. (2001), Leakey et al. (2003), and Kenzo et al. (2004).
the equator, mirroring the latitudinal decline in mangrove biomass and litterfall (Saenger and Snedaker, 1993). How do these data compare with productivity data for tropical rain forests? First, we must compare data obtained using identical or very similar methods. A comparison in CO2 leaf assimilation rates between mangroves and tropical terrestrial trees indicates
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Mangroves
Terrestrial forest
0 5 10 15 20 25 30 Aboveground net primary production (t DW/ha/year)
35
Figure 3 Comparison of above-ground net primary production in mangrove and tropical terrestrial forests based on measurements of biomass increments and litterfall.Vertical line in box denotes median and the boxes encompass the 25 and 75th percentiles and the lower bars denote the 5 and 95th percentiles, respectively. Data from Table 3 and from Clark et al. (2001) and Scurlock and Olson (2002) for terrestrial forests.
great overlap and thus close similarity in rates of leaf photosynthetic rates (Figure 2). Second, the most comprehensive database for both mangroves and tropical terrestrial forests involves measurement of only aboveground biomass accumulation plus litterfall. Comparing the data in Table 3 and the data analyzed by Clark et al. (2001) and Scurlock and Olsen (2002), we find a similar range of productivity estimates (Figure 3). For mangroves, the mean rate of aboveground net primary production is 11.1 t DW/ha/year, with a median value of 8.1 and 25th and 75th percentiles of 4.6 and 19.2, respectively. For terrestrial forests, the mean rate of aboveground NPP is 11.93 t DW/ ha/year, with a median value of 11.4 and a 25th percentile of 8.8 and a 75th percentile of 14.4. This similarity suggests that rates of net primary production are equivalent between mangrove and tropical terrestrial forests underscoring similarities in physiological and ecological factors regulating tree production. Rates of belowground production are sorely lacking for all forests. Respiration of roots and woody parts are also badly needed to adjust the photosynthetic rates for an estimate of net carbon fixation. Like other forests, mangrove stands vary in size and age over time, and therefore vary in rates of production and in the balance between production and respiration. Long-term patterns are important to discern as they reflect a balance between factors promoting and limiting forest growth, but only a few studies have examined the growth dynamics of mangrove forests over time or of stands of known age (Ong et al., 1985; Day et al., 1996; Fromard et al., 1998; Clough et al., 2000; Alongi et al., 2004).
2.2. Mangrove forests appear to be architecturally simple, but factors regulating succession and zonation are complex Early workers observed conspicuous changes in mangrove forest composition from the water’s edge to the highest point of tidal inundation and proposed several
Paradigm Shifts in Mangrove Biology
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hypotheses to explain these zonation patterns (Watson, 1928; Macnae, 1968). Even into the 1970s and 1980s, attempts were made either to deduce the relative importance of various physical and biological regulatory factors or to categorize forests by type (Lugo and Snedaker, 1974). Although mangrove stands ordinarily have little or no understory of scrubs and ferns and are usually less rich in tree species than other tropical forests, we now understand that their distribution and the factors regulating their floral composition and structure are highly complex (Smith, 1992; Ellison, 2002). On the local scale (e.g., a tidal height gradient or upestuary gradient), mangrove distribution is regulated by a complex panoply of biotic and environmental factors, including salinity, soil type, degree of anoxia, nutrient availability, physiological tolerances, predation, and competition (Smith, 1992; Ellison, 2002). The interplay of these factors is so complex that it is reasonable to state that generalizations about the mechanisms governing zonation are prohibitive. The reality is that each stand is different, and at the local scale, different sets of factors come into play over different temporal and spatial scales (Bunt, 1996; Ball, 1998). Salinity, for example, may be a crucial regulatory factor in one estuary (or part thereof) but less so in an adjacent estuary. Processes observed at the local scale can obscure other drivers that are important in structuring mangrove forests over longer scales of space and time. Like other forests, mangrove communities follow a natural succession of stages over time, from an initial pioneering stage of early rapid growth and development through to later maturity, senescence, and death (Fromard et al., 1998; Ward et al., 2006). However, it has been observed that mangrove forests do not always fit neatly into established ecological concepts such as the old-growth or late-successional forest (Lugo, 1997). This apparent paradox is explainable if the role of natural and anthropogenic disturbance is considered, in which the recovery of a forest from a disturbance (such as a hurricane) can “reset the clock” regarding successional stage (Piou et al., 2006). Further, a mosaic of successional stages can coexist within the same stand; tree growth and development can be disrupted by such disparate events as lightning and harvesting. A recent study of the effects of a hurricane on mangrove structural dynamics in south Florida (Ward et al., 2006) showed that forest turnover rates were greater in plots inside the primary hurricane wind field than outside. They showed that mangroves adhere to the same organizing factors as terrestrial forests and to the Twilley et al. (1998) model which predicts that initial differences in stand structure and subsequent seedling and sapling establishment rates would largely control the development of forest structure post-disturbance. The recruitment of seedlings in mangroves, as in other forests, is controlled by gaps in the canopy created by disturbance (Smith, 1992; Clarke, 2004). Compared with the total pool of species available, the number of species actually colonizing a canopy gap (usually 10–100 m2) is small. Why? As explained by Clarke (2004), recruitment limitation does not appear to facilitate coexistence; the seeds of local canopy members are much more likely to be successful in colonizing available space than propagules dispersed with distance. This explanation has been supported by more recent evidence of gap dynamics in mangroves of southeast Asia and Africa (Bosire et al., 2006; Iman et al., 2006). In reforested mangrove stands in Kenya, Bosire et al. (2006) found that recruitment into monospecific stands of adult
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Sonneratia alba of other species was unsuccessful (100% mortality) suggesting competitive exclusion. Therefore, seed predation by crabs and subsequent competitive exclusion by conspecific canopy dominants play a major role in regulating recruitment in gaps and subsequent forest succession. To summarize, we can state that stand composition and structure of mangrove forests are the result of a complex interplay of physiological tolerances and competitive interactions leading to a mosaic of interrupted or arrested succession sequences in response to physical/ chemical gradients and changes in geomorphology (Alongi, 2008).
2.3. Mangrove tree growth is not constant but related to climate patterns For decades, growth rings found in mangrove tree species were thought to be anomalous, reflecting variations in the density of wall deposits rather than the result of variations in cambial activity as related to seasonal changes in climate (Chapman, 1975; Tomlinson, 1986). Further research using improved technology has recently discovered genuine growth rings in Rhizophora apiculata (Yu et al., 2004), Rhizophora mangle (Menezies et al., 2003), and Xylocarpus mekongensis (Hancock et al., 2006), in regions with distinct seasons. Tree ring patterns in R. apiculata (Yu et al., 2004) in subtropical China were consistent with trends in annual sea level, salinity, and sea surface temperatures, indicating that the measurement of changes in alpha cellulose d 13C in mangrove rings can be used as a potential indicator of past changes in sea level. In Brazil, R. mangle forms annual rings, with the slowest growing trees (1.2 mm/year) showing a close relationship between the number of months with rainfall <50 mm and ring width. The distinctiveness of rings was greater in trees from saline soils compared with those in brackish soils (Menezies et al., 2003). Using a different set of isotopes (228Ra/226Ra), Hancock et al. (2006) were able to measure 3 rings/ year in X. mekongensis, matching climate patterns in northern Australia. Like other mangroves, Rhizophora mucronata was also thought to lack distinct growth rings, but using high-resolution stable carbon and oxygen isotope ratios, Verheyden et al. (2004, 2005) found an annual cyclicity in the isotope ratios of stem wood that appears to relate best to mean relative humidity and rainfall. Moreover, they found a clear discontinuity in the pattern that may reflect unusually low rainfall during El Nin˜o. Mangrove dendrochronology is still in its infancy, so further studies should shed more light on long-term patterns of tree growth in response to climate change.
2.4. Tree diversity is low, but faunal and microbial diversity can be high Whether considered from either the generic or species level, the diversity of mangrove tree species within a given area is low compared with other tropical forests (Saenger, 2002). Highest diversity is found in the Indo-Malaysia area of the Indo-west Pacific and the lowest diversity occurs in western Africa (Table 4). While up to approximately 50 species are found in the most diverse
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Paradigm Shifts in Mangrove Biology
Table 4 Number of genera and species of mangroves among the six biogeographical provinces (Adapted from Alongi, 2002) Province West America East America West Africa East Africa Indo-Malaysia Australasia
Genera
Species
8 7 6 11 22 21
12 11 8 11 51 47
mangrove areas, tropical rainforests may contain up to several hundred tree species (Clark et al., 2001). And while some invertebrate phyla are similarly species poor (e.g., nematodes and other benthic meiofauna; Alongi, 1989), recent evidence suggests much greater species richness among fungi, bacteria, protists, viruses, and other phyla (Kathiresan and Bingham, 2001). An adequate inventory of microbes in mangrove forests has not yet been developed, but new species are continually being found throughout the world, especially as microbial techniques and methods improve. With technological advances, it is reasonable to expect the true number of bacterial species to eventually number in the thousands. Even on a small piece of leaf litter, a rich variety of microbes exist. For instance, 12 species of flagellates, 2 species of sarcodines, 17 species of ciliates, and 2 species of suctorids were discovered on a single mangrove leaf in India, in addition to several species of nematodes, diatoms, and copepod nauplii (Padma et al., 2003). Like the bacteria, it is likely that the true number of species of other microbiota is far higher. Similarly, fungal diversity is high, especially on rotting vegetation. On Rhizophora leaves in a Panamanian forest, over 183 different morphological types of fungi were discovered, with over 60 types on Avicennia leaves and 106 types on Laguncularia leaves (Gilbert et al., 2002). As of 2003, a total of 625 species of terrestrial, freshwater, and marine fungi have been reported from mangroves worldwide, including 278 ascomycetes, 277 mitosporic fungi, 30 basidiomycetes, and 14 oomycetes (Schmit and Shearer, 2003). Different fungal species are often found in different parts of the forest. Diversity of fungi was higher in woody litter than on leaves, with 78 taxa found on the floor of two Indian mangrove forests (Ananda and Sridhar, 2004). As in other ecosystems, species diversity declines as individual body size increases. Most aquatic invertebrate groups consist of a few to <50 species within a given forest area (Alongi and Sasekumar, 1992) with highest diversity most often found among the crustaceans (Kathiresan and Bingham, 2001). Insects and birds, although most are only temporary visitors, are highly diverse with species numbers often exceeding 300 within a single mangrove estuary. Fish are the most diverse among vertebrate phyla with species numbers usually ranging from 100 to 250 per estuary (Robertson and Blaber, 1992). In a southeast Asian mangrove estuary, a maximum of 260 fish species was recorded (Hong and San, 1993). Such wide ranges of species numbers are a reflection of variable environmental conditions. There are also biogeographical differences. For instance, east Africa has a reduced
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mangrove crab richness (about 35 species) compared with southeast Asia (>100 species; Gillikin and Schubart, 2004), mirroring the diversity differences between the regions in mangrove flora. At the local scale, metazoan diversity is, on average, higher on the tree (encrusting or epibiont assemblages) or on the forest floor surface and in tidal waters than within the forest floor (Alongi, 1989).
2.5. Arboreal communities are important in food webs, exhibiting predatory, symbiotic, and mutualistic relations Like other forests, the fauna and flora inhabiting the mangrove canopy are important in structuring food webs and in influencing the species composition of mangroves, but this was not recognized until quite recently (Ellison and Farnsworth, 2001; Kathiresan and Bingham, 2001). Insects ordinarily consume mangrove material equivalent to only approximately 5% of net primary production (Robertson, 1991), but recent findings point to the importance of insects in affecting the establishment and growth of seedlings (Minchinton and Dalby-Ball, 2001; Burrows, 2003; Sousa et al., 2003) and as pollinators (Ellison and Farnsworth, 2001). In a series of field and shadehouse experiments, Sousa et al. (2003) found that the rate of insect attack varied as a function of intraspecific variation in propagule size and tree species, with Avicennia suffering 90% damage compared with 33% for Laguncularia and 20.5% for Rhizophora, respectively. These species differences were partly explained by differences in chemical composition of propagules. Major defoliation has been reported in mature forests from Hong Kong, Colombia, Costa Rica, Ecuador, Indonesia, India, Bangladesh, and Thailand (see references in Saenger, 2002) and, although rare, illustrates the potential for insect herbivory to greatly affect tree mortality. Birds and mammals either temporarily or permanently reside in mangrove forests, using the forest as shelter and to find food. The work of Lefebvre (Lefebvre et al., 1992, 1994; Lefebvre and Poulin, 1996, 1997) has established the importance of mangroves as a home for many species of birds, with some forests containing up to 315 species and feeding extensively on invertebrates on the trees, on the forest floor, and in tidal water. In Brazilian mangroves, Martinez (2004) found that sympatric Scarlet Ibis (Eudocimus ruber) and Yellow-Crowned Night Heron (Nyctanassa violacea) fed mainly on ocypodid crabs in significantly greater proportion relative to their abundance. Prey segregation plays a role in sustaining such numerous species of birds in mangroves, with prey size usually being the prime factor. Birds may exhibit comparatively restricted feeding niches given the high availability of prey within the forest. However, it is more likely that mangroves have a more significant impact on the life cycles of many bird species than vice versa; mangrove habitat destruction and fragmentation has been shown to reduce population of mangrove-dependent birds (Grant and Grant, 1997). Predation has been somewhat over-emphasized at the expense of other important trophic relationships. For instance, epiphytes growing on mangrove trees and tree parts also have a role to play in the biology of mangrove forests (Ellison and
Paradigm Shifts in Mangrove Biology
627
Farnsworth, 2001). Epiphytic orchids, bromeliads, mistletoes, and ferns are less common and of lower diversity than in upland forests but can affect the physiology of mangroves (Orozco et al., 1990) and exhibit mutualism with insects, especially ants, which provide nutrients to the epiphytes in return for shelter (Rico-Gray et al., 1989). Many epibionts do not interact directly with the trees but many fouling communities interact both indirectly and directly. In Belize, Ellison and her colleagues (Ellison and Farnsworth, 1990; Ellison et al., 1996) found that sponges living on roots precluded colonization of and damage to the roots by isopods and facilitated nitrogen uptake by the roots. When present, sponges induced the formation of fine roots that absorbed ammonium produced by the sponge fouling communities. Such mutualistic relationships likely evolved as a result of mangrove forests and their associated biota inhabiting an oligotrophic environment.
2.6. Plant–Microbe–Soil Relations are tightly linked and help conserve scarce nutrients It is now recognized that the high rates of photosynthesis and primary productivity of many mangrove forests depend on not only unique and highly evolved physiological mechanisms (Ball, 1988) but also highly evolved and energetically efficient interrelationships among soil nutrient pools, microbes, and trees (Alongi, 2005). Such close linkages are necessary in tropical habitats, as available nutrient pools (e.g., nitrate) are small and microbial growth is rapid, in the face of coping with a harsh, waterlogged environment. However, owing to the physical nature of soil, there is more inferential data than direct evidence of the complexity and nature of these interrelationships. As for other vascular plants, bacteria, fungi, and other microflora alter the microenvironment around mangrove roots via their metabolic activities, transforming and releasing nutrients, and modifying soil chemistry. These microbes depend on the leakage of nutrients from roots as a source of energy. Thus, the relationship is mutualistic, as both trees and microbes share the need for limiting nutrients. Indeed, both bacterial transformation of nutrients and subsequent tree growth are often rapid, as reflected in nutrient use efficiencies of mangroves equal to or higher than those of other tropical trees (Alongi et al., 2005). Mangrove–microbe relations have been most closely observed within the rhizosphere where highly specialized groups of bacteria and fungi coexist within the root matrix (Sengupta and Chaudhuri, 2004; Ravikumar et al., 2004; Kothamasi et al., 2006). In mangroves lining the lower reaches of the Ganges, rates of nitrogenfixing activity and numbers of bacterial colonies are related to forest age, with richer and more active colonies in the rhizosphere of early successional mangroves and lower microbial activity in mature mangrove stands (Sengupta and Chaudhuri, 2004). When the roots of seedlings and saplings of 16 mangrove species were inoculated with cultured nitrogen-fixing bacteria, root biomass nearly doubled, leaf area increased by 270%, and shoot biomass increased by nearly 30% compared with control plants (Ravikumar et al., 2004). Also in Indian mangroves,
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Kothamasi et al. (2006) found arbuscular mycorrhizal fungi in the aerenchymatous cortex of several mangrove species, suggesting that the plants may be providing the fungus with oxygen; phosphate-solubilizing bacteria were also abundant suggesting that these bacteria mobilize insoluble phosphates for the plants. Other workers have also recently found that bacterial colonies reside within the aerenchyma, including methogens (Purvaja et al., 2004). Other nutrient elements may also be made available to mangroves via bacterial activities, as indicated by observations of active bacterial iron and manganese reduction in close association with mangrove roots (Kristensen et al., 2000; Alongi et al., 2005; Kristensen and Alongi, 2006). Mangrove trees alter the soil environment, and this affects the growth and survival of individual functional types of aerobic and anaerobic bacteria. It has been known for decades that translocation of oxygen to the roots serves as a means of oxidizing potentially toxic metabolites, such as sulfides. However, it has only recently been shown that these activities shift the competitive balance for substrates from favoring sulfate reducers to favoring iron- and manganese-reducing bacteria, thus increasing availability of soluble Fe and Mn required for plant growth. Highly evolved and energetically efficient plant–soil–microbe relations are a major factor in explaining why mangroves can be highly productive in a typically harsh environment.
2.7. Crabs are keystone species influencing function and structure in many, but not all, mangrove forests Recognition of the importance of crabs in mangroves emerged from research conducted in Australia in the 1980s, resulting in a paradigm shift as to how mangroves link to adjacent coastal habitats. The established view based on research conducted earlier in the Caribbean was that, like their salt marsh counterparts, mangroves export large quantities of detritus to the adjacent coastal zone, subsidizing coastal food webs and the flow of energy and nutrients (Odum and Heald, 1975; Lee, 1999; Alongi, 2002). Studies in tropical Australia (and confirmed in other tropical mangroves worldwide) found that grapsid crabs, especially sesarmids, consume a significant fraction of leaf litter lying on the forest floor (Robertson, 1986), thus reducing the detritus subsidy to the adjacent coastal zone. Grapsid and ocypodid crabs are the most important organisms influencing the structure and function of many tropical mangrove forests, after bacteria and the trees (Lee, 1998, 1999; Kristensen, 2008). Through their life activities, they exert extraordinary influence on a wide variety of mangrove processes. Through their consumption of mangrove leaf litter, they significantly reduce the amount of detritus available for export, thus enhancing retention and recycling of nutrients and organic matter internally; their wastes can support coprophagous organisms further ensuring conservation of materials within the forest, and their selective consumption of mangrove propagules affects forest structure by reducing the recruitment and relative abundance of tree species whose propagules are preferentially consumed. Bioturbation by crabs also results in changes in soil texture and chemistry, surface topography, degree of anoxia, and abundance of meiofauna while stimulating microbial
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production. The presence of crab burrows enhances the flow of tidal water through the forest floor, speeding up the flow of water and associated dissolved and particulate material between forest and adjacent waterway (Ridd, 1996). Crabs, however, are not keystone species in all mangrove ecosystems. In temperate Australian mangrove-salt marsh systems and in some Caribbean mangroves, crabs play only a minor role in litter decomposition and in structuring forests and, in fact, often avoid eating mangrove leaves and seeds (Smith et al., 1989; Saintilan et al., 2000; Guest et al., 2004, 2006). Why crabs are important in some forests but not others is unclear, but it may be due to different combinations of crab and mangrove species between locations. Recent work has focused on clarifying the trophic role of crabs, especially positive feedback loops and interactions with trees and other flora and fauna in relation to food availability (Ashton, 2002; Kristensen and Alongi, 2006), and their reproductive and life history strategies in relation to tree composition and environmental factors (Lee and Kwok, 2002; Koch et al., 2005; Moser et al., 2005). In mesocosm experiments, Kristensen and Alongi (2006) found that the presence of the fiddler crab, Uca vocans, stimulated the growth and development of Avicennia marina saplings but depressed the abundance and productivity of microalgal mats at the soil surface. The association between the saplings and the crabs also greatly influenced the pathways of microbial decomposition (Table 5), with sulfate reduction being more important than iron reduction in the presence of crabs and saplings. Fiddler crabs and tree roots thus appear to have complementary effects on sediment microbial processes. Sesarmid crab biology and productivity may be related to forest type, and forest type may be linked to crab activities. In field experiments excluding crabs, Smith et al. (1991) found that the absence of crabs increased the concentration of ammonium and sulfide in soils, but reduced plant stipule and propagule production. The presence of crabs therefore facilitates plant growth by aerating the soil to limit the buildup of toxic metabolites. The presence of tree species may also influence crab productivity. In experiments in the Mai Po marshes in Hong Kong, Lee and Kwok (2002) found faster reproductive rates and higher sesarmid crab productivity in a K. candel stand as opposed to an adjacent A. marina forest. These differences Table 5 Rates of sulfate and iron (Fe III) reduction in microcosm sediments in various treatments with and without fiddler crabs and Avicennia marina saplings Depth integrated rates
CþP
CP
þCþP
þCP
Total C production Sulfate reduction
80 + 22 50 + 18 (63%)
Iron reduction
19 + 4 (23%)
90 + 12 63 + 28 (70%) 16 + 6 (18%)
56 + 18 21 + 7 (37%) 24 + 5 (44%)
38 + 7 24 + 5 (62%) 14 + 2 (36%)
Rates are given as mmol C/m2/day (+SD). Values in parentheses indicate the percentage contribution of sulfate and iron reduction to total carbon production. C, crabs absent; þC, crabs present; P, saplings absent; þP, saplings present (Adapted from Kristensen and Alongi, 2006).
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may be due to other factors, such as differences in tidal height, but modification of the soil and/or differences in food availability may play an important role. Regardless of the mechanisms involved, positive interactions between trees, crabs, and microbes make ecological sense in that the overall stability of mangrove ecosystems is enhanced.
2.8. Algae, not just detritus, are a significant food resource The pioneering work of Odum and Heald (1975) in south Florida in the late 1960s established the paradigm that detritus-based food chains were the primary mode by which mangroves support nearshore secondary production. This initial paradigm was modified in the early 1980s to include alternative energy and carbon sources for consumers (Odum et al., 1982), but much of the idea that mangroves are linked to coastal secondary production and energy flow rested, until very recently, on carbon fixed by mangroves. Recent work indicates that various types of algae are a preferred food for many organisms, including those ordinarily classified as detritivores and that herbivory and omnivory are more common than previously believed (Ellison and Farnsworth, 2001; Bouillon et al., 2008). Grapsid crabs are often voracious consumers of mangrove detritus and leaves, but the nutritional rationale for ingesting such material has been a mystery, as mangrove material is high in tannin and C : N content. Even aged leaves that have undergone microbial enrichment do not possess enough sustenance, especially labile nitrogenous compounds, to maintain crab nutrition. How do crabs sustain themselves? The answer is that grapsid crabs consume other foods in order to sustain a balanced diet. In field experiments, Skov and Hartnoll (2002) observed that crabs fed mostly on sediment, with <10% of the time spent on eating leaves. These observations suggest that crabs forage on bacteria, protists, fungi, and other small organisms from the sediment surface and within their burrows. Indeed, ocypodid crabs have high assimilation efficiencies for bacteria and are efficient consumers of benthic microalgae (France, 1998; Kristensen and Alongi, 2006). Grapsid crabs may behave similarly. Recent evidence using fatty acid markers suggests that grapsid crabs consume fungi and bacteria in addition to mangrove litter, whereas ocypodid crabs consume mostly benthic microalgae (Meziane et al., 2006); both groups supplement their diets with other foods. The use of stable isotopes established that zooplankton (Bouillon et al., 2000; Werry and Lee, 2005; Schwamborn et al., 2006), polychaetes (Hsieh et al., 2002), fish (Melville and Connolly, 2004), and a variety of other macroinvertebrates (Bouillon et al., 2002) consume and often utilize various algal foods in order to sustain their nutrition. Studies conducted in southeast India have found that benthic invertebrates rely almost entirely on benthic microalgae (Bouillon et al., 2000, 2002, 2004, 2008). However, southeast Indian mangrove estuaries are the exemplar of coastal habitats subjected to a wide variety of pollutants, especially organic wastes from heavily populated catchments (Sudhakar and Venkateswarlu, 1989) and so may be atypical of other mangroves. With such high organic loading, Indian estuaries have very high rates of primary production (Monbet, 1992), thus the
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diets of the animals in these habitats reflect the high availability of algal foods, such as benthic microalgae and phytoplankton. On the trees, the arboreal fauna is known to readily feed on various organisms growing epiphytically on stems, branches, fallen timber, and pneumatophores (Kathiresan and Bingham, 2001). Terrestrial and semiterrestrial gastropods, mites, crustaceans and insects, as well as nematodes, copepods, and other estuarine biota colonize vertical wood structures in mangrove forests and use both sessile and planktonic algae for food (Alongi and Sasekumar, 1992; Nagelkerken et al., 2008). Even pneumatophores are sites of intense biological and trophic activity involving highly diverse and abundant metazoan and algal communities (Proches and Marshall, 2002). Regardless of physical niche, it makes nutritional sense for secondary consumers to feed on a variety of foods of higher nutritional quality in addition to mangrove detritus.
2.9. Mangroves are an important link to fisheries Mangroves have long been considered important nursery grounds for finfish and shellfish (Macnae, 1968). This paradigm, however, is still controversial because it has rarely, if ever, been empirically tested. There is little direct evidence of a positive relationship between mangrove area and fisheries yield (Faunce and Serafy, 2006). One reason for this lack of empiricism is the fact that most species utilize mangroves for only part of their life cycle and for only part of the time in synchrony with the tides, using other habitats, such as adjacent seagrass beds and coral reefs, when mangroves are inaccessible (Sheaves, 2005). These biological connections between mangroves, seagrass beds, intertidal flats, and coral reefs have been considered by Sheaves (2005) as an “interconnected habitat mosaic,” in which the connectivity of these different habitats must be considered when establishing the nature of mangrove–fisheries linkages. This idea is a more realistic and pragmatic consideration of the problem than the simplistic views (and studies) offered in the past and is supported by more recent analyses of the available data for shrimp (Lee, 2004) and for all fisheries, including finfish (Saintilan, 2004; Manson et al., 2005). The analysis of shrimp yield and mangrove data by Lee (2004) involved the use of principal components analysis rather than use of regression analysis as various independent variables are usually autocorrelated. Using data on mangrove area, tidal amplitude, rainfall, temperature, human population density, length of coastline, relative mangrove abundance, and marine shrimp catch, the analysis showed that shrimp yield is strongly correlated with tidal amplitude, suggesting that shrimp catch is influenced by the amount of intertidal area available and not just the area of mangroves. And while the length of coastline was a strong correlate, it is likely that this relationship reflects the simple fact that a longer (and larger) coastline would result in a larger shrimp catch. It does not necessarily mean that the presence of mangroves per se is the cause of greater shrimp yields. There was no significant relationship between shrimp catch and relative mangrove area. This analysis points to the idea offered by Sheaves (2005) for fish that the connectivity of adjacent habitats needs to be considered as shrimp do not utilize only mangroves as habitat
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throughout their life cycle. Dietary evidence supports this view as stable isotope studies have shown that the proportion of mangrove matter used as food for shrimp declines further offshore. While shrimp proximal to mangroves are fairly mangrove food-dependent, offshore shrimp utilize mainly phytoplankton and, to a lesser extent, benthic microalgae (Chong et al., 2001). So, mangroves are important for a part of the life cycle of shrimp, but shrimp are not wholly dependent on mangrove forests and their waterways. Mangroves appear to be linked to fisheries catch in three principle ways: as a refuge from predators, as a source for food, and as shelter from environmental disturbance (Robertson and Blaber, 1992; Manson et al., 2005). Several studies have identified correlations between mangroves and fisheries, but as Manson et al. (2005) point out, correlative information does not shed light on causal relationships. Part of the problem from a statistical viewpoint is that some species are estuarine residents, spending their entire lives within the mangroves, whereas others spend only crucial juvenile stages in the estuaries, while some species are stragglers, moving into and out of the estuaries and having no clear dependence on mangroves. These various life styles and consequent relationships of fish and shrimp with mangroves are too subtle to be captured statistically. A diverse array of fish and invertebrate species clearly use mangrove habitats for a variety of reasons, for different periods of time, and at different life history stages. At present, it is difficult to separate the value of mangrove attributes from estuarine attributes more generally, as the exact nature of fisheries use of mangroves is not clearly understood (Manson et al., 2005). Empirical evidence of the importance of mangroves to fisheries species lies in studies that have considered the effect of changes in mangrove habitat on faunal abundance and diversity. Manson et al. (2005) cited studies that found (1) a loss of densities and diversity of fauna commensurate with the loss of mangrove forest, (2) a decline in faunal richness and abundance with physical disturbance, and (3) a net gain in faunal numbers and species diversity with an increase in mangrove forest. Thus, the global decline in mangrove forest area (Wilkie and Fortuna, 2003) has likely resulted in a loss of faunal (and floral) biodiversity and yield of fisheries species.
2.10. Mangroves are chemically diverse and a good source of natural products Mangroves, like other tropical vascular plants, possess a number of chemical defense mechanisms against potential herbivores. Chemical metabolites derived from mangroves have been used traditionally by indigenous people for centuries but, until recently, the chemistry of these natural products has remained poorly defined (Bandaranayake, 2002). We now know that most mangrove species produce a plethora of natural products, such as antioxidants and terpenoids, to modulate physiological activities, including membrane permeability to salt and other solutes (Oku et al., 2003; Parida et al., 2004; Vijayavel et al., 2006) in addition to providing chemical defense. Aliphatic alcohols and acids, amino acids, alkaloids, carbohydrates, lignins, polysaccharides, carotenoids, hydrocarbons, fatty acids, lipids, pheromones,
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Table 6
Individual natural products discovered from mangroves and their uses
Mangrove species
Chemical
Real and Potential Use
Acanthus ilicifolius
Stigmasterol 2-Benzoxazoline
Bruguiera sexangula B. conjugata
Brugine Brugierol
Heritiera littoralis Aegiceras corniculatum
Vallapianin Embelin, 5-O-methy embelin Excoecarin Rhizophorine
Hypercholestrolemic properties Nerve depressant, muscle relaxant Antitumor activity Antibacterial, insecticidal activity Ichthyotoxin Ichthyotoxin, antifungal
Excoecaria agallocha Rhizophora mucronata Sonneratia acida Xylocarpus moluccenis, X. granatum
2-nitro-4-(20 -nitroethenyl) phenol Xyloccensin 1 and 2
Ichthyotoxin Insecticidal, control diabetes, astringent Antihemorrhaging agent Antimalarial
Source: Bandaranayake (2002).
phorbol esters, phenolics, steroids, troterpenes, tannins and other terpenes, and related compounds have been isolated from various parts of mangroves. Extracts from mangroves and associated biota have proven activity against human, animal, and plant pathogens although the number of studies has, until very recently, been limited. Nearly all species of mangroves possess some chemical agents that have some human use (Bandaranayake, 2002; Table 6). Mangroves are a rich source of toxins, such as rotenoids, alkaloids, and terpenoids, which can be developed as repellants or agrochemicals. These toxins, if properly developed and purified, may prove useful for the treatment of diseases. A good example is the drug sodium stiboggluconate, derived from Acanthus ilicifolius, and used to treat infections of Leishmania donovani (Kapil et al., 1994). Useful metabolites have also been discovered from mangrove-associated biota, such as fungi (Chen et al., 2003; Wu et al., 2005). In addition, it has been proposed that hydrocarbon-degrading bacteria should be batch cultured and used to biologically degrade oil spills (Brito et al., 2006). Clearly, the chemistry and potential benefits of natural products derived from mangrove tissues has just begun to be adequately explored.
3. C ONCLUSIONS Mangrove forests have often been viewed as tropical counterparts to salt marshes, with analogous roles in the ecology of coastal food webs and energy flow. Often underappreciated, even by scientists, a picture is gradually emerging of a highly dynamic and complex ecosystem that (while similar in some respects to salt
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marshes) is unique. Mangroves are not simple systems whose main function is to support and supplement coastal food webs but are highly productive forests with complex physiological and growth processes (especially in relation to climate patterns), high species diversity of biota, highly evolved plant–soil and arboreal animal–plant relations, and high chemical diversity. They are indeed important to fisheries, including species important to adjacent tidal flats, seagrass meadows, and coral reefs, making these tidal forests of considerable ecological and economic importance, disproportionate to their dwindling area worldwide.
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Lee, S.Y., 2004. Relationship between mangrove abundance and tropical prawn production: a re-evaluation. Mar. Biol. 145, 943–949. Lee, S.Y., Kwok, P.W., 2002. The importance of mangrove species association to the population biology of the sesarmine crabs Parasesarma affinis and Perisesarma bidens. Wetl. Ecol. Manage. 10, 215–226. Lefebvre, G., Poulin, B., 1996. Seasonal abundance of migrant birds and food resources in Panamanian mangrove forests. Wilson Bull. 108, 748–759. Lefebvre, G., Poulin, B., 1997. Bird communities in Panamanian black mangroves: potential effects of physical and biotic factors. J. Trop. Ecol. 13, 405–415. Lefebvre, G., Poulin, B., McNeil, R. 1992. Settlement period and function of long-term territory in tropical mangrove passerines. Condor 94, 83–92. Lefebvre, G., Poulin, B., McNeil, R. 1994. Temporal dynamics of mangrove bird communities in Venezuela with special reference to migrant warblers. Auk 111, 511–514. Lugo, A.E., 1997. Old-growth mangrove forests in the United States. Conserv. Biol. 11, 11–20. Lugo, A.E., Evink, G., Brinson, M.M., Broce, A., Snedaker, S.C., 1975. Diurnal rates of photosynthesis, respiration, and transpiration in mangrove forests of South Florida. In: Golley, F.B., Medina, E. (Eds.), Tropical Ecological Systems: Trends in Terrestrial and Aquatic Research. Springer-Verlag, Berlin, pp. 335–350. Lugo, A.E., Snedaker, S.C., 1974. The ecology of mangroves. Annu. Rev. Ecol. Syst. 5, 39–64. Macnae, W., 1968. A general account of the flora and fauna of mangrove swamps in the Indo-Pacific region. Adv. Mar. Biol. 6, 73–270. Manson, F.J., Loneragan, N.R., Skilleter, G.A., Phinn, S.R., 2005. An evaluation of the evidence for linkages between mangroves and fisheries: a synthesis of the literature and identification of research directions. Oceanogr. Mar. Biol.: Annu. Rev. 43, 483–513. Marenco, R.A., Gonc¸alves, J.F. de C., Vieira, G., 2001. Leaf gas exchange and carbohydrates in tropical trees differing in successional status in two light environments in central Amazonia. Tree Physiol. 21, 1311–1318. Martinez, C., 2004. Food and niche overlap of the Scarlet Ibis and the Yellow-crowned Night Heron in a tropical mangrove swamp. Waterbirds 27, 1–8. Melville, A.J., Connolly, R.M., 2004. Spatial analysis of stable isotope data to determine primary sources of nutrition for fish. Oecologia 136, 499–507. Menezies, M., Berger, U., Worbes, M., 2003. Annual growth rings and long-term growth patterns of mangrove trees from the Braganca peninsula, North Brazil. Wetl. Ecol. Manage. 11, 233–242. Meziane, T., d’Agata, F., Lee, S.Y., 2006. Fate of mangrove organic matter along a subtropical estuary: small-scale exportation and contribution to the food of crab communities. Mar. Ecol. Prog. Ser. 312, 15–27. Miller, P.C. (1972). Bioclimate, leaf temperature, and primary production in red mangrove canopies in south Florida. Ecology, 53, 22–45. Minchinton, T.E., Dalby-Ball, M., 2001. Frugivory by insects on mangrove propagules: effects on the early life history of Avicennia marina. Oecologia 129, 243–252. Monbet, Y., 1992. Control of phytoplankton biomass in estuaries: a comparative analysis of microtidal and macrotidal estuaries. Estuaries 15, 563–571. Moser, S., Macintosh, D., Laoprasert, S., Tongdee, N., 2005. Population ecology of the mud crab Scylla olivacea: a study in the Ranong mangrove ecosystem, Thailand, with emphasis on juvenile mortality. Fish. Res. 71, 27–41. Nagelkerken, I., Blaber, S.J.M., Bouillon, S., Green, P., Haywood, M., Kirton, L.G., Meynecke, J.-O., Pawlik, J., Penrose, H.M., Sesekumar, A., Somerfield, P.J., 2008. The habitat function of mangroves for terrestrial and marine fauna. Aquat. Bot. 89, 155–185. Odum, W.E., Heald, E.J., 1975. The detritus-based food web of an estuarine mangrove community. In: Cronin, L.E. (Ed.), Estuarine Research. Academic Press, New York, pp. 265–286. Odum, W.E., McIvor, C.C., Smith III., T.J., 1982. The ecology of the mangroves of south Florida: a community profile. U.S. Fish and Wildlife Service. Office of Biological Services, Washington DC, 144pp. Oku, H., Baba, S., Koga, H., Takara K., Iwasaki, H., 2003. Lipid composition of mangrove and its relevance to salt tolerance. J. Plant Res. 116, 37–45. Ong, J.E., Gong, W.K., Clough, B.F., 1995. Structure and productivity of a 20-year-old stand of Rhizophora apiculata Bl. mangrove forest. J. Biogeogr. 22, 417–424.
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Sheaves, M., 2005. Nature and consequences of biological connectivity in mangrove systems. Mar. Ecol. Prog. Ser. 302, 293–305. Sherman, R.E., Fahey, T.J., Martinez, P., 2003. Spatial patterns if biomass and aboveground net primary productivity in a mangrove ecosystem in the Dominican Republic. Ecosystems 6, 384–398. Skov, M.W., Hartnoll, R.G., 2002. Paradoxical selective feeding on a low-nutrient diet: why do mangrove crabs eat leaves? Oecologia 131, 1–7. Smith III., T.J., 1992. Forest structure. In: Robertson, A.I., Alongi, D.M. (Eds.), Tropical Mangrove Ecosystems. American Geophysical Union, Washington DC, pp. 101–136. Smith III., T.J., Boto, K.G., Frusher, S.D., Giddens, R.L., 1991. Keystone species and mangrove forest dynamics: the influence of burrowing by crabs on soil nutrient status and forest productivity. Estuar. Coast. Shelf Sci. 33, 419–432. Smith III., T.J., Chan, H.T., McIvor, C.C., Robblee, M.B., 1989. Comparisons of seed predation in tropical tidal forests from three continents. Ecology 70, 146–151. Sousa, W.P., Kennedy, P.G., Mitchell, B.J., 2003. Propagule size and predispersal damage by insects affect establishment and early growth of mangrove seedlings. Oecologia 135, 564–575. Sudhakar, G., Venkateswarlu, V., 1989. Ecological imbalances in the rivers of Andhra Pradesh (India): effect of paper mill effluents on the rivers Tungabhadra and Godavari. Int. J. Environ. Stud. A B 34, 89–97. Sukardjo, S., 1995. Structure, litterfall and net primary production in the mangrove forests in East Kalimantan. In: Box, E.O., Fujiwara, T. (Eds.), Vegetation Science in Forestry. Kluwer Academic Publishers, Dordrecht, The Netherlands, pp. 585–611. Sukardjo, S., Yamada, I., 1992. Biomass and productivity of a Rhizophora mucronata Lamarck plantation in Tritih, Central Java, Indonesia. For. Ecol. Manage. 49, 195–209. Suwa, R., Khan, M.N., Hagihara, A., 2006. Canopy photosynthesis, canopy respiration and surplus production in a subtropical mangrove Kandelia candel forest, Okinawa Island, Japan. Mar. Ecol. Prog. Ser. 320, 131–139. Tomlinson, P.B., 1986. The Botany of Mangroves. Cambridge University Press, Cambridge, 419pp. Turner, I.M., 2001. The Ecology of Trees in the Tropical Rain Forest. Cambridge University Press, Cambridge, 298pp. Twilley, R.R., 1985. An analysis of mangrove forests along the Gambia River Estuary: implications for the management of estuarine resources. Great Lakes and Marine Waters Center International Programs Report No. 6. University of Michigan, pp. 1–75. Twilley, R.R., Rivera-Monroy, V.H., Chen, R., Botero, L., 1998. Adapting an ecological mangrove model to simulate trajectories in restoration ecology. Mar. Pollut. Bull. 37, 404–419. Verheyden, A., De Ridder, F., Schmitz, N., Beeckman, H., Koedam, N., 2005. High-resolution time series of vessel density in Kenyan mangrove trees reveal a link with climate. New Phytol. 167, 425–437. Verheyden, A., Helle, G., Schleser, G.H., Dehairs, F., Beeckman, H., Koedam, N., 2004. Annual cyclicity in high-resolution stable carbon and oxygen isotope ratios in the wood of the mangrove tree Rhizophora mucronata. Plant Cell Environ. 27, 1525–1536. Vijayavel, K., Anbuselvam, C., Balasubramanian, M.P., 2006. Free radical scavenging activity of the marine mangrove Rhizophora apiculata bark extract with reference to naphthalene induced mitochondrial dysfunction. Chem. Biol. Interact. 165, 170–175. Ward, G.A., Smith III., T.J., Whelan, K.R.T., Doyle, T.W., 2006. Regional processes in mangrove ecosystems: spatial scaling relationships, biomass, and turnover rates following catastrophic disturbance. Hydrobiologia 569, 517–527. Watson, J.G., 1928. Mangrove forests of the Malay Peninsula. Malay. For. Rec. 6, 1–275. Werry, J., Lee, S.Y., 2005. Grapsid crabs mediate link between mangrove litter production and estuarine planktonic food chains. Mar. Ecol. Prog. Ser. 293, 165–176. Wilkie, M.L., Fortuna, S., 2003. Status and Trends in Mangrove Area Extent Worldwide. Forest Resources Assessment Working Paper No. 63. Forest Resources Division, FAO, Rome 20pp þ Appendices. http://www.fao.org/docrep/007/j1533e/J1533E00.HTM Wu, X., Liu, X., Jiang, G., Lin, Y., Chan, W., Vrijmoed, L.L.P., 2005. Xyloketal G, a novel metabolite from the mangrove fungus Xylaria sp. 2508. Chem. Nat. Comp. 41, 27–29. Yu, K.-F., Zhao, J.-X., Liu, T.-S., Wang, P.-X., Qian, J.-L., Chen, T.-G., 2004. a-cellulose d13C variation in mangrove tree rings correlates well with annual sea level trend between 1982 and 1999. Geophys. Res. Lett. 31, 749–756.
C H A P T E R
2 3
E COGEOMORPHIC M ODELS OF N UTRIENT B IOGEOCHEMISTRY FOR M ANGROVE W ETLANDS Robert R. Twilley and Victor H. Rivera-Monroy
Contents 1. 2. 3. 4.
Introduction Ecogeomorphology of Mangroves (Model 1) A Multigradient Model (Model 2) Geochemical Model (Model 3) 4.1. Redox zones in mangrove soils 4.2. Transition from reduced to oxidized zones 4.3. Hydroperiod effects on transition zones 4.4. Lower oxidation zone 4.5. Linkages in multigradient and geochemical models 5. Soil Biogeochemistry Model (NUMAN, Model 4) 6. Mass Balance Exchange (Model 5) 6.1. CO2 efflux from mangrove sediments and tidal waters 6.2. The balance of N fixation and denitrification 6.3. Tidal exchange 7. Contrasting Coastal Settings and Biogeochemical Models Acknowledgments References
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1. INTRODUCTION Mangroves cover 240 103 km2 of sheltered subtropical and tropical coastlines (Lugo, 1990; Twilley et al., 1992), which forms a small portion of the world’s total coastal ocean (260 105 km2) and even less percentage of the forested landscape. Yet this landscape features 350,000 km of shorelines, including continental and oceanic islands, that are considered important to the productivity of tropical estuaries (Boto et al., 1985; Lugo, 1990; Twilley et al., 1996) and significant to global biogeochemical processes (Twilley et al., 1992; Bouillon et al., 2008a). Estuaries and coastal ecosystems represent significant contributions to the secondary Coastal Wetlands: An Integrated Ecosystem Approach
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productivity and biogeochemical processes of the coastal region that are disproportionate to their land cover. This may be particularly true where mangrove vegetation dominates the intertidal and supratidal zones of a variety of tropical and subtropical coastal settings including river deltas, muddy coasts, estuaries from rias to lagoons, and carbonate settings of continents and oceanic islands. There is a diversity of geomorphological settings that can be subdivided into a continuum of landforms based on the relative processes of river input, tides, and waves (Woodroffe, 2002). There is some indication that these diverse geomorphological habitats, each with different vegetation types, results in specific characteristics of coastal ecosystem structure and function as proposed by Thom (1967, 1982) and described by Twilley (1988, 1995, 1997) and Woodroffe (2002). This correlation between coastal landform and ecological function has particularly been documented relative to the net primary productivity (NPP) and detritus exchange across a variety of mangrove locations. This review will utilize several conceptual models to test how regional and local patterns of geomorphological settings are related to patterns of nutrient biogeochemistry among different ecological types of mangrove wetlands (Twilley, 1997). One of the major factors defining the different coastal settings in this review is the source of sediment input associated with gradients in river and tidal processes among mangrove wetlands (Figure 1; Woodroffe, 2002). One group of mangroves can be characterized by sediment patterns that are associated with delivery of clastic sediments varying in dominance by river discharge. Such terrigenous sediment systems can be classified in two broad subgroups in this
Interior Sink
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Figure 1 The classification of mangroves byWoodroffe (2002).
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review. The first group is muddy coasts that represent those continental margins influenced by river inputs characteristic of deltaic coasts (Woodroffe, 2002). Such settings are in the immediate region of major river basins; or may be downstream where massive fluid muds are transported by nearshore currents and accumulate far-field to form muddy shorelines (Woodroffe, 2002; Marchand et al., 2003). This group of river-dominated deltaic coasts can be subdivided into prograding and transgressive types with gradient in river, tidal, and wave processes forming a continuum of geomorphological settings. The second group of clastic systems is represented by estuarine coasts that have a wide range of freshwater and marine sediment inputs with varying salinities from rias to lagoons (tide-dominated in Figure 1). These coastal systems do not have the muddy shorelines of deltaic systems, but mangroves in such settings have a significant component of their soils that include clastic sediments either on continental or island shorelines. High islands in the Pacific Ocean in wet precipitation zones (>3 m/year) for example have mangrove soils that are dominated by sediments eroded from upland watersheds. In contrast to these coastal systems dominated by clastic sediments is the third group of sedimentation patterns that result from in situ processes, or biogenic formation, of sediments such as carbonate formation along reef coasts (Gattuso et al., 1998). This group may be considered an extreme of tide-dominated mangroves with no river input in Figure 1. Mangroves in these types of continental or island landforms are dominated by peaty and calcareous soils that are a combination of carbonate formation and mangrove productivity, both of which originate largely through carbon fixation. We will use these three groups, muddy coasts, estuarine coasts, and carbonate coasts, to link geomorphological settings and geophysical processes (Figure 1) to patterns of nutrient biogeochemistry. We will review five models (conceptual and numerical) to compare and contrast the biogeochemistry of mangroves across muddy, estuarine, and carbonate coasts of tropical and subtropical coastal settings. The first model is based on hypotheses concerning community structure and ecosystem processes of mangroves based on a hierarchical classification scheme of geomorphological settings and coastal processes (Figure 2). This conceptual model is dependent on gradients of geophysical and climate processes among coastal settings (Twilley et al., 1999; Woodroffe, 2002). A second conceptual model describes how multigradient patterns of resources, regulators, and hydroperiod control mangrove community structure and ecosystem function (Figure 3; Twilley and Rivera-Monroy, 2005). As suggested in Figure 2, these three gradients can describe many of the patterns in mangrove productivity, and we will investigate in this review how these three gradients may also be linked to patterns of nutrient biogeochemistry. In fact, patterns of biogeochemistry along hydroperiod gradients determine both the levels of some regulators; and how these regulators control nutrient uptake. The mechanisms as to how these three gradients interact are described in our third model, which is a geochemical model of redox zones in mangrove soils (Figure 4; Clark et al., 1998). This model describes details of redox
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Global distribution: 25°N Temperature 25°S Delta
Oceanic islands
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Figure 2 Hierarchical classification system to describe patterns of mangrove structure and function based on global, geomorphological (regional), and ecological (local) factors that control the concentration of nutrients resources and regulators in soil along gradients from fringe to more interior locations from shore (modified fromTwilley et al., 1999 and Twilley and Rivera-Monroy, 2005).
reactions of organic matter respiration and associated chemical reactions with metals across estuarine interfaces. This model clearly links the seasonal effects of climate, tidal processes, and topography (fringe vs. interior) to soil redox zones of mangroves. The fourth is a numerical model (NUMAN, Chen and Twilley, 1999a) describing interactions among organic matter production, decomposition, and sedimentation that control the accumulation of carbon, nitrogen (N), and phosphorus (P) in mangrove soils (Figure 5). This model is the only numerical model of the set; but we will use model formulation to explain that different processes across geomorphological settings can result in different levels of nutrient sinks in mangrove soils. And fifth is a mass balance model (Figure 6) that describes the role of mangroves in the exchange of carbon and N in estuarine landscapes.
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(High nutrients)
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Figure 3 Factorial interactions controlling the productivity of coastal wetlands including regulators gradients, resource gradients, and hydroperiod. (a) The production envelope associated with levels of each factor interact to demonstrate responding levels of net primary productivity. (b) The definition of stress associated with how gradients in each factor control growth of wetland vegetation (fromTwilley and Rivera-Monroy, 2005).
N and carbon exchange with tidal waters and atmosphere is one of the most complex processes to understanding the ecosystem function of mangrove forests. The five models summarize many of the biogeochemical pathways in mangrove ecosystems by starting with a simple conceptual framework of river and tides across inundation gradients and climate that result in many of the functional characteristics of mangroves as sources or sinks of nutrients in coastal landscape. This hierarchical approach of models 1 and 2 together with the mechanisms described in models 3–5 should help explain the diverse patterns of nutrient biogeochemistry in mangrove ecosystems across geomorphological settings and geophysical processes.
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O2
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Figure 4 The geochemical model of mangrove oxidation and reduction zones (Clark et al., 1998).
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Oxygen leakage from roots
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Figure 5 Conceptual model of variables included in the NUMAN model (Chen and Twilley, 1999a).
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Figure 6 Mass balance exchanges of nutrients across atmosphere and tidal boundaries (using N as example) in mangrove ecosystems.
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2. ECOGEOMORPHOLOGY OF MANGROVES (M ODEL 1) The ecogeomorphic and ecological classifications of mangrove wetlands, which are clearly delineated by natural boundaries, reflect a combination of spatial hierarchical levels that form patterns at different temporal scales that can explain global patterns across different coastal landforms (Figure 2). The ecogeomorphic classification of mangroves describes the relative role of geophysical processes across different geomorphological settings and their interaction with ecological processes, which can account for much of the diversity in mangrove community structure and ecosystem function (Thom, 1984; Woodroffe, 1992, 2002; Twilley, 1995; Twilley et al., 1996). Climate and landform characteristics of a coastal region together with local geophysical processes (river, tide, and waves) control the basic properties of mangrove wetlands (Thom, 1984; Twilley, 1995). Global distribution constrains the latitudinal expanse of mangroves along coastal landforms, and biogeographic provinces define what mangrove species will colonize the landscape. These coastal geomorphic settings can be found in a variety of humidity provinces (Blasco, 1984) that depend on regional climate and oceanographic processes, along with river input from catchments that are far inland of coastal climates. Oceanic processes and river input will also determine the source of sediments as either clastic or biogenic as discussed above (Woodroffe, 2002). River input of sediments is a function of catchment size, climate, and lithology. These factors along with geologic setting, topography and hydrodynamic processes determine sedimentation that defines geomorphologic types. However, these global and geomorphologic types are not clearly delineated but represent a continuum in the relative effects of climate, river, waves, and tides on coastal processes that determine mangrove distribution. We are using muddy, estuarine, and carbonate as a general classification of the types that represent extreme conditions that may define nutrient biogeochemistry of mangroves. There will certainly be research results from sites that fit within the continuum; and thus confuse any particular pattern. And these are all very dynamic landforms with associated changes in habitat types with sedimentation patterns, largely associated with changes in sea level. It is important not to consider any association between landform and ecological function as a static model; but part of the evolutionary pattern of coastal settings, particularly given the degree of climate changes proposed. These landform conditions establish the constraints of finer grain processes (referred to as ecological processes) that explain the variety of habitats within any one type of geomorphological setting. In each of the three geomorphological settings that is the focus of this review (muddy, estuarine, and carbonate coasts), local variations in topography and hydrology result in the development of distinct ecological types of mangroves such as riverine, fringe, and interior forests (Figure 1; Lugo and Snedaker, 1974; Lugo, 1990; Woodroffe, 1992, 2002). Local scale gradients in tidal and river processes (Figure 1) form diverse mangrove types along the shoreline; and these processes change with increased distance perpendicular from the shore to more interior locations that can form mangrove zones (Watson, 1928; Walsh, 1974; Chapman, 1976). This chapter will describe how local patterns of nutrient biogeochemistry respond to the degree of inundation as a critical control of biochemical processes.
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Several conceptual models have described inundation classes (number of tidal inundations per year) as a way to explain variation in community composition of mangrove forests (Watson, 1928; Chapman, 1976). These classes of inundation also explain distinct patterns in nutrient dynamics of mangrove soils from fringe to interior locations. And finally, interior mangroves are particularly sensitive to local climate conditions, again due to effects of water budget and soil salinity (Twilley and Chen, 1998). We will investigate if the nutrient biogeochemistry of interior mangroves differs in wet compared to dry climate zones. Finally, the habitat units that are the focal point of this review are sites that define research results of nutrient biogeochemistry in mangrove wetlands. We will develop three mechanisms that control the link in geomorphological setting and ecological function as relative to concentrations of resources and regulators across gradients in hydroperiod. These three gradients are a function of the same processes and conditions that form the diverse geomorphological settings of a coastline; and the sediment source in each of these habitat units will largely influence patterns in nutrient biogeochemistry. The remaining four models in this review will help associate the diverse mechanisms under different geomorphological settings that establish links to nutrient biogeochemistry in mangroves. The expression of ecological patterns within geomorphological settings can be defined by these three gradients of resources, regulators, and hydroperiod; and all three are linked to nutrient biogeochemistry of mangroves.
3. A M ULTIGRADIENT M ODEL (M ODEL 2) A combination of ecological types of mangroves can occur within any one of the geomorphological settings described above (Figure 2) depending on the distribution of soil resources and abiotic regulators along with hydroperiod gradient (Figure 3). Ecological processes are constrained by a variety of stress conditions in mangrove soils associated with resources, regulators, and hydroperiod (Huston, 1994; Twilley, 1997; Twilley and Rivera-Monroy, 2005). Resource gradients such as nutrients, light, and space are variables that are consumed in the process of contributing to mangrove productivity; compared to regulator gradients such as salinity, sulfide, pH, and metals that are not consumed in the process of controlling mangrove growth (Tilman, 1982). The third gradient, hydroperiod, is the frequency, duration and depth of inundation that controls wetland productivity (Gosselink and Turner, 1978). At low levels of stress for all three environmental gradients, such as low salinity, high nutrients, and intermediate flooding, mangrove wetlands will reach their maximum levels of biomass and net ecosystem productivity. Coastal settings that result in higher stress conditions for any one or more of these gradients, and thus result in plots closer to the origin of the three axes, should reflect a lower total NPP. The interactions of these three gradients have been proposed as a constraint envelope for defining the structure and productivity of mangrove wetlands based on the relative degree of stress conditions (Figure 3) (Twilley and RiveraMonroy, 2005). And these conditions of stress among these three gradients are associated with various combinations of geophysical processes and geomorphological settings. Using this model, we will describe critical feedback effects of resources,
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regulators, and hydroperiod that describe the significance of biogeochemical processes on community structure and ecosystem function of mangroves. The biogeochemistry of mangrove soils has a significant effect on limiting mangrove productivity based on the mechanisms associated with resources, regulators, and hydroperiod described in Figure 3. One of the most critical regulator gradients controlling propagule establishment, seedling survival, growth, height, and zonation is salinity (Scholander et al., 1962; Chapman, 1976; Tomlinson, 1986; Ball, 1988; Chen and Twilley, 1998). Studies show that Rhizophora mangle and Laguncularia racemosa have a narrower salt tolerance than Avicennia germinans. It is reported that the interspecific differential response of mangrove propagules to salinity occurs at salinities from 45 to 60 g/kg, which inhibited the growth of R. mangle and L. racemosa significantly more than that of A. germinans (McKee, 1993). A. germinans is generally dominant in coastal settings with dry (perhumid) climate where evaporation potential exceeds precipitation and soil salinities of interior mangroves are >120 g/kg (Castan˜eda-Moya et al., 2006). Hydroperiod gradient is another critical factor associated with frequency and duration of flooding that controls seedling establishment and growth in mangroves. Flooded mangrove soils may be anaerobic and reducing with high concentrations of sulfides present depending on duration of standing water (details described in the geochemical model above). Greenhouse experiments show differential tolerance of mangrove seedlings to flooding (Cardona et al., 2006; Krauss et al., 2006). Yet, field experiments also show that adult mangrove plants can grow in soils with high concentration of sulfide such as 120 mg/L for A. germinans (Nickerson and Thibodeau, 1985), 1.63 mmol/L for R. mangle and 1.44 mmol/L for A. germinans (Mckee et al., 1988), and 3 mmol/L for A. germinans and 1.7 mmol/L for R. mangle (McKee, 1993). Coasts with high geophysical energies and moist climates have lower stress conditions associated with the interaction of hydroperiod and regulators, particularly along the fringe zone. The effect of these two gradients on interior mangroves will depend on climate and upland runoff (moist vs. dry conditions). Estuarine and carbonate settings have interior mangroves more commonly limited by redox and salinity gradients, particularly in dry coastal settings. In the absence of stressors, soil nutrient availability is often implicated as the principal factor determining variation in forest structure and biomass production. Comparative analysis of mangrove forests and soil characteristics, in addition to fertilization studies, revealed that P availability was one of the major factors limiting annual growth of mangrove forests (Boto and Wellington, 1984a; Boto et al., 1985; Feller, 1995; Twilley, 1995). A study of mangrove forest structure along the Shark River estuary in south Florida showed that the landward decline of forest structure was associated with a decrease in soil P concentration (Chen and Twilley, 1999b). In addition, the mean N:P atomic ratios of mangrove soils along this estuary varied from <20 in the downstream locations (tree height = 15 m), 40 in the intermediate zone (tree height = 8 m), to >80 in the upstream location (tree height = 4 m). We used a Monod model (Bridgham et al., 1995) to generalize the nature of P limitation of mangrove forest development in New World mangroves (Figure 7). The relationship between forest development and nutrient resource availability in mangroves is based on sites where soil salinity were all <50 g/kg, thus minimizing the complication of regulators on forest development. The model is as follows (Bridgham et al., 1995):
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Basal area (m2/ha)
50 40 30 20 R 2 = 0.77
10 0 0
0.125 0.250 0.375 Total P (mg/cm3)
0.500
0.625
Figure 7 Model of P concentration in soil versus basal area of mangrove forests in NewWorld.
PR =
ðRac Rmin ÞPRmax Rac Rmin þ
where PR is the production, PRmax is the maximum production, Rac is the amount of resource acquired, Rmin is nutrient availability at zero production, and is half-saturation constant with respect to Rac (see Bridgham et al., 1995 for details). To apply this Monod model to results in this review, we used basal area as an index of biomass production and total P (per unit volume of soil) as an estimate of nutrient availability (Bridgham et al., 1995). Basal area is a more direct estimate of forest production than biomass, because errors can be introduced by different allometric equations used to calculate biomass in different geographical areas (Day et al., 1987; Simard et al., 2006, 2008). We included mangrove sites in an estuary setting along Florida Coastal Everglades (FCE) (Shark River estuary); scrub mangroves along southeast carbonate Florida coast; interior mangroves in southwest Florida Lagoon (Rookery Bay); fringe and riverine mangroves in a lagoon–delta complex (Terminos Lagoon, Mexico); and mangrove forests in riverine, tidal, and interior forests of watersheds in Pacific islands (Korsae and Phonpei, Federated States of Micronesia) (Figure 7). Basal areas were regressed against respective contents of soil P using the Monod model above and a nonlinear fitting technique by minimizing residual sum of squares. There was a significant range in total P concentrations among the mangrove sites, and basal area was strongly correlated with total soil P (Figure 7). The fit of the Monod model was significant (r2 = 0.77), and Rmin was positive with an optimum soil P content of 0.25 mg/cm3. The halfsaturation constant of soil P based on the nonlinear model was 0.12 mg/cm3. This model includes mangroves from different geomorphologic settings that support a more global understanding of P limitation in mangrove wetlands that lack other stress effects of salinity, sulfide or flooding. Strong P limitation of mangroves in this analysis (<0.07 mg/cm3) is associated with carbonate settings that lack terrigenous sediment input. Mangroves along the mouth of Shark River estuary, which is also located in the carbonate setting of the FCE, has sediment input from Gulf of Mexico during hurricanes, which is proposed as the source of P that supports optimum
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mangrove growth in this region (Chen and Twilley, 1998). Field experiments have shown that mangrove forests growing under nutrient-limited conditions respond quickly to fertilization (Feller, 1995). Studies in carbonate-dominated environments in the Caribbean show that P enrichment significantly increases productivity of interior forests along a microtidal gradient whereas trees were generally N limited in the fringe zone (Feller et al., 1999, 2003). This response is associated with P concentrations of about 0.06 mg/cm3, below the half-saturation concentration of our analysis (Figure 7); and similar to concentrations found in high carbonate sediments in more peaty interior soils. Additions of P in scrub mangroves on marl sediments east of the FCE sites also stimulated mangrove productivity (Koch, 1996). Evidence of nutrient resource limitation in mangroves can also be evaluated by summarizing existing information on soil nutrient density among mangrove forests (68 measurements summarized in Figure 8). Nearly one third of the sites have total 30
Frequency
25 20 15 10 5 0 0.0 0.25 0.50 0.75 1.00 1.25 1.50 1.75 2.00 2.25 2.5 Total nitrogen (mg/cm3) 30
Frequency
25 20 15 10 5 0 0.0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0 Total phosphorus (mg/cm3) 30
Frequency
25 20 15 10 5 0 0
15
30 45 60 75 90 105 120 135 150 Nitrogen: Phosphorus ratios
Figure 8 Survey of nutrient density (upper panel is total nitrogen and middle panel is total phosphorus) from 68 mangrove sites around the world (based on mg per volume of sediment). Lower panel is the nitrogen:phosphorus ratio (atomic) of these nutrient concentrations for each respective site.
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P density <0.10 mg/cm3 representing mangroves in either interior carbonate settings (isolated from hurricane deposits), or interior lagoon settings with limited sediment input (such as basin sites at Rookery Bay and Terminos Lagoon). Nearly half of these mangrove sites surveyed have N:P ratios >15, indicating that P is potentially limiting under those conditions (Figure 8). However, both of these surveys demonstrate that N is potentially limiting in many locations where sediment deposition in estuarine and muddy coasts provide P density in soils sufficient to saturate growth conditions. This trend in the relative content of N and P among mangrove sites in muddy, estuarine, and carbonate settings is probably one of the strongest linkages between geomorphological settings and ecology of mangrove wetlands (Thom, 1967).
4. G EOCHEMICAL M ODEL (M ODEL 3) The mechanisms that establish gradients of resource availability and soil regulators, and how hydroperiod influences both of these gradients, are described by a geochemical model of mangrove soils developed by Clark et al. (1998) (Figure 4). Many of these mechanisms are associated with the different zones of oxidation and reduction depending on routes that dissolved oxygen is exchanged with the atmosphere. Oxygen distribution, which mostly is restricted to upper sediments but can also be found below upper reduction zones, can be important to the organic matter respiration and biogeochemistry of mangrove sediments. There are four redox zones that can be identified with depth in mangrove soils (Figure 4; Clark et al., 1998). The upper oxidation zone is established when the supply of oxygen is sufficient to balance aerobic respiration; supply being largely a function of oxygen physically exchanging with surface mangrove sediments. The presence and absence of water, and the dissolved oxygen concentration of that water when present, establishes that oxygen diffusive flux. The seasonal variation in the upper oxidation zone is very sensitive to the hydroperiod gradient that occurs with distance from fringe to interior wetlands, and on the depth of the water table. Below the oxidation zone is a reduction zone, where sulfate-reducing bacteria dominate the respiration of organic matter resulting in redox values well below –100 mV. Biological exchange of oxygen into deeper sediments, usually below the upper reduction zone, results in a lower oxidation zone; below that is the lower reduction zone (Figure 4). Upper and lower zones that are either oxidized or reduced result from the balance of oxygen supply relative to oxygen consumption by microbial respiration, which may vary with time and depth. And oxygen supply is strongly linked to water levels. We will describe how these zones vary in response to physical and biological processes that vary across muddy, estuarine, and carbonate coastal settings.
4.1. Redox zones in mangrove soils A primary determinant of the biochemical processes in mangrove soils is the relative distribution of oxidation and reduction zones that are dependent on the presence of dissolved oxygen. The upper oxidation zone in mangrove soils is controlled by the diffusive limit of atmospheric oxygen, which is usually limited by presence of water in
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pores of wetland soils (0.205 cm2/s in air vs. 0.227 10–4 cm2/s in water). In addition, the normally high concentration of organic matter in mangrove soils, due to both litter and root productivity, establishes high sediment oxygen demand. Oxygen is the preferred electron acceptor in biochemical respiration of this organic matter supply, generating – 475 kJ of Gibbs free energy per mol of organic matter (CH2O) along with inorganic nutrients of N and P with stoichiometry of 16:1 (Klump and Martens, 1983) as follows: ðCH2 OÞ106 ðNH4 Þ16 H3 PO4 þ106O2 ! 106CO2 þ16NH3 þH3 PO4 þ106H2 O ð1Þ Most mangrove soils lack oxygen and have redox potentials below 200 mV (which usually indicates oxygen as terminal point of electron flow during respiration). Several mechanisms can enhance oxygen supply by increasing diffusion into otherwise oxygen deficient soils (Clark et al., 1998) including (1) sediment texture, which dictates porosity and permeability that control diffusive flux (Hutchings and Saenger, 1987); (2) the position of the water table, which may vary seasonally with hydroperiod (balance of tides and rain events; Twilley and Chen, 1998); and (3) the influence of biota (directly and indirectly) on water and oxygen exchange with sediments (Thongtham and Kristensen, 2005; Kristensen, 2008). When the dissolved oxygen necessary to supply aerobic respiration is absent, there are a series of alternate electron acceptors used by facultative and obligate anaerobes to terminate electron flow during respiration. Each alternate electron acceptor is associated with decreasing efficiency of free energy per mol of organic matter (CH2O). Collectively, these respiration pathways in the absence of oxygen represent the reduction zone (Figure 4). The ecological significance of each oxidation–reduction reaction is very distinct in mangrove soils due to the effects of the various by-products of each microbially mediated reaction. The collection of electrons by each of these reactants results in the following products under reduced conditions: manganese (Mn3þ) and iron (Fe3þ) are reduced to Mn2þ and Fe2þ, NO 3 is reduced to N2 or 2– N2O (discussed below), SO2 4 is reduced to sulfide (S ), and CO2 is reduced to methane (CH4). Since sulfate is normally the most abundant electron acceptor in the absence of oxygen in most mangrove soils (due to presence of salinity), the “reduction zones” are characterized by the presence of decomposing organic matter and substantial populations of sulfate-reducing bacteria that produce sulfides (Figure 4). The reduction of sulfate to sulfide in the presence of metals produces a variety of reduced metal sulfides (Jørgensen, 1982; Berner, 1984). Most sulfides in mangrove soils are dominated by pyrite (FeS2) and mackinawite (FeS) (Pons et al., 1982).
4.2. Transition from reduced to oxidized zones The anaerobic conditions that usually dominate mangrove soils can be converted to oxidized zones either by enhancing oxygen transport in flooded conditions (increase geophysical energy of flow) or enhanced oxygen transport in the absence of flooding as water levels decrease and expose soil to the atmosphere. Oxygen diffusion in surface sediments can occur in interior mangroves when sediments dry out periodically or when they become well aerated by biological processes, for example, bioturbation or oxygen release from mangrove roots. Exposure of anaerobic soils to oxygen may cause sulfides in the sediment to oxidize (Bloomfield and
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Coulter, 1973). Sulfides are stable in the reduced zone; however, as reduced zones transition to oxidized zones, these minerals are readily oxidized following a couple of possible reactions producing byproducts that are chemically and ecologically significant (Singer and Stumm, 1970): FeS2 þ
7 O2 þ H2 O ! Fe2þ þ 2SO42 þ 2H þ 2
FeS2 þ 14Fe3þ þ 8H2 O ! 15Fe2þ þ 2SO42 þ 16H þ
ð2Þ ð3Þ
Large amounts of sulfuric acid are produced by these reactions, thereby lowering pore water pH and potentially resulting in mobilization of metals in the sediment (Dent, 1986; Evangelou, 1995). Nonferrous metal sulfides can also decompose by a similar sequence of reactions. Both the pyrite oxidation by either oxygen or oxidized Fe and subsequent jarosite formation yield large quantities of Hþ that can reduce pH of mangrove soils. If this acidity is not buffered in some way within the soil system, then pH values <4 can occur, which are typical of acid sulfate soils.
4.3. Hydroperiod effects on transition zones Hydroperiod has a significant influence on the depth and duration of the “upper oxidation zone” as a result of exposing pore space of mangrove soils to atmospheric oxygen. This exposure will vary from fringe (and riverine) to interior wetlands based on the duration that the water table is below the sediment surface. Interior mangroves, such as demonstrated with a hydrology model of basin mangrove forests at Rookery Bay, can have soils exposed to the atmosphere for several weeks during seasonally dry part of the year (Twilley and Chen, 1998). This condition in interior mangroves, with corresponding elevated redox potentials, demonstrates that surface sediments oxidize as oxygen diffuses across the sediment–air interface during prolonged periods of lower water table. When rainy season occurs or higher seasonal sea levels cause tidal inundation further inland, water tables will rise thus allowing sulfate reduction to occur nearer to the sediment surface resulting in lower redox values. Prolonged waterlogging can cause the reduction zone to migrate up to the sediment surface, eliminating the “upper oxidation zone.” However, fringe, riverine, and interior mangrove forests in the wet tropics of Micronesia exhibited very little difference in redox either with site or season (Ewel et al., 1998). This was attributed to heavy annual rainfall (nearly 5 m) and/or strong influence of shallow groundwater flow from upland catchment basins. In this wet climate regime, specific patterns of redox are more closely associated with mangrove species (Gleason et al., 2003) rather than mangrove habitat type (fringe, riverine, interior). The geochemical model in Figure 4 represents a drier coastal setting and describes more influence of fringe versus interior mangrove types on redox patterns. Soil conditions in perhumid environments can be strongly linked to strong gradients within 25 m of shoreline (Castan˜eda-Moya et al., 2006), indicating the strong influence of climate boundary conditions on general models of redox conditions within interior mangrove forests (Twilley and Chen, 1998). Seasonal changes in the position of the water table in mangrove sites in Australia (Wynnum sites, Clark et al., 1998) are most noticeable in the landward part of the
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mangrove forest and in the area of dead mangroves, whereas the influence of tides is most evident in the seaward part of the mangrove forest and in the tidal flat sediments (Clark et al., 1997). In the more landward parts of the study area, depression of the water table during periods of low rainfall allows the upper portions of the sediment to dry and oxidize, whereas during periods of high rainfall, the water table rises and sediment redox conditions decrease due to waterlogging (Beckwith et al., 1975; Armstrong, 1978; Gleason and Zieman, 1981; Clark et al., 1997). When the water table falls, sulfides in near-surface sediment can oxidize, releasing sulfide-bound metals and acid that may leach metals from nearby sediment, such that the zone of maximum metal concentration in the sediment can shift seasonally in response to vertical changes in the redox profile (Clark et al., 1998). The shallow depth of soils (2–5 cm) in the fringe mangrove zone along the shore is more consistently oxidized due to more frequent inundation with oxygenated waters that maintain elevated redox potentials. Under conditions of prolonged water deficit in interior mangroves, lower water tables may allow the oxidation zone to reach depths of 10 or 20 cm; but the depth of the oxidation zone is more variable, being entirely reduced during the wet season of the year. These patterns demonstrate how redox conditions can control metal cycling in mangrove forests depending on tidal frequency in the fringe mangrove zone compared to climate in interior zones. In low-energy coastal settings (microtidal estuarine and carbonate settings) these patterns are particularly conspicuous. The highly oxidized soils of mangrove soils at the mouth of muddy coasts such as French Guiana and the Fly River are associated with high redox potentials up to soil depths of 50 cm and generally lack vertical profiles in dissolved and particulate nutrients, bacterial numbers, and dissolved metal concentrations (Fabre et al., 1999; Marchand et al., 2003, 2006).
4.4. Lower oxidation zone The upper oxidized and reduced redox zones discussed so far are defined by the relative supply and consumption of dissolved oxygen. Below the upper reduction zone is a “lower oxidation zone” that can develop due to the release of oxygen from mangrove roots (Armstrong, 1978, 1982; Thibodeau and Nickerson, 1986) or animal burrows. The resulting oxidation is accompanied by a reduction of pore water pH as iron sulfides are converted to oxy-hydroxides and sulfuric acid (Equations 2,3). Radial release of oxygen from mangrove roots affects a microzone surrounding the root that may only be a few millimeters in depth (rhizosphere); and sediments at a greater distance from the roots can remain reducing and sulfidic. The duration and depth of this lower oxidation zone can depend on the type of mangroves (radial roots vs. prop roots) and the density of mangrove roots in response to soil conditions (McKee et al., 1988; Clark et al., 1998; Gleason et al., 2003). Where mangroves are sparse or absent, the “lower oxidation zone” will be poorly defined or absent. Thus the depth distribution of roots not only provides a source of organic matter that promotes respiration but also serves as source of atmospheric oxygen for aerobic respiration of this organic matter in an otherwise anaerobic environment. Oxide-bound and exchangeable metals are more common in the “lower oxidation zone,” yet sulfide-bound metals are also abundant and become much more so in areas where there are few mangrove roots due to diffusion from the lower reduction zone below (Clark et al., 1998).
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The lower oxidation zone may also be formed below the upper reduction zone due to animal tunnels formed by burrowing organisms, in addition to the ventilation activity of animals residing in the borrows (Figure 4; Clark et al., 1998). Sediment turnover by burrowing fauna can have a major influence on the geochemical and physical character of sediment. In mangrove forests, mud lobsters Thallasina anomola commonly provide obvious examples of sediment turnover (Macnae, 1966; Bennett, 1968). But the numerous burrows of small mud crabs, shrimp, and various worms are clearly responsible for most sediment turnover in the Wynnum area (Australia) where the geochemical model was developed (Clark et al., 1998). In addition to sediment turnover, burrows and water pumping by their inhabitants provide a means for oxygen to be transferred into the otherwise anoxic sediment, and increase subsurface water flow (Hutchings and Saenger, 1987). The resulting ventilation of anoxic sediments by oxygenated waters can increase the Cd, Cu, and Ni concentrations in oxidized surface sediments and enhance the fluxes of metals and other dissolved materials, for example, CO2, HS, and CH4 between the sediment and water in contact with it (Aller, 1978; Morris et al., 1982; Emerson et al., 1984). The movement of water through burrows and mangrove root and pneumatophore casts also provides a means whereby sulfate reducing bacteria in the sediment are supplied with sulfate (Bloch and Krouse, 1992) and the by-products of bacterial metabolism (e.g., HS) can be flushed from the sediment (Aller, 1978; Emerson et al., 1984).
4.5. Linkages in multigradient and geochemical models There are linkages and feedbacks between the multigradient and geochemical models described above that are a function of the respiration of organic matter in mangrove soils. Landscape gradients in hydroperiod establish the depth that oxygen diffuses into surface sediments, particularly in seasonally dry forests. Hydroperiod particularly establishes strong seasonal gradients in interior forests where water controls the depth of the surface oxidized zone and allows reduction zones to migrate to the soil surface. As described above, the relative distribution of reduction zones produces by-products of anoxic respiration including HS and Hþ ions that both represent regulator gradients of the multigradient model. The effects of these regulators on carbon exchange (NPP) may be direct since HS and Hþ ions can be toxic to plants; but can also determine plant productivity by reducing nutrient uptake rates due to the presence of these byproducts of anaerobic respiration. These regulators and their effect on nutrient uptake will determine mangrove zonation patterns; in the New World mangroves this will influence the relative distribution of Avicennia and Rhizophora. There are feedback mechanisms of these mangroves on the distribution of the oxidized zone by release of oxygen from roots, which may limit the toxic effects of these anaerobic byproducts. Yet the indirect effect of these anaerobic byproducts on utilizing resources is less evident in determining mangrove productivity (Gleason et al., 2003). The multigradient and geochemical models may explain variation in biogeochemistry among carbonate, estuarine and muddy coasts associated with sulfate reduction in anaerobic environments and different forms of P in mangrove soils. Sulfate reduction forms hydrogen sulfide, which will react to form pyrite, elemental sulfur, or iron monosulfides (Holmer et al., 1994). Mangrove soils show that
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aerobic respiration and anaerobic respiration by sulfate reduction are about 50% of total organic matter respiration on mangrove soils. In iron-rich muddy coasts, iron respiration can reach rates of 35 mmol/cm3 or about 80% of carbon oxidation by iron reduction (Bouillon et al., 2008a). In these types of environments, CH4 flux and CO2 respiration is a very small component of total respiration at about 0–5 mmol/m2/day. In mature forest in estuarine environments with highly reduced muds, sulfate reduction is about 80% of total organic carbon respiration. In similar coasts but with sandy sediments and more tidally inundated zones, sediments may be more oxidized resulting in less sulfate and iron reduction (Alongi and McKinnon, 2005). These patterns in biochemical pathways that dominate formation of redox zones can also strongly influence phosphate availability along inundation gradients. Under oxidized conditions in presence of iron, P becomes adsorbed onto iron oxyhydroxides forming occluded forms of P reducing the concentration of soluble reactive P in mangrove pore waters (Holmer et al., 1994; Sherman et al., 1998). Under anoxic conditions in the presence of sulfate and iron, sulfate reduction results in the formation of iron sulfides, mostly pyrites, and P goes back into solution in pore waters (Krom and Berner, 1980). This process of increasing P solubility in mangrove pore waters may vary from fringe to interior forests depending on the supply of oxygen, supply of sulfate, and the availability of iron. When oxygen is absent and both sulfate and iron are present, the pyrite forming potential is high increasing P solubility potential (Sherman et al., 1998; Gleason et al., 2003). But even in the absence of oxygen, either the supply of sulfate or iron can influence the relative solubility of P. When pyrite is oxidized in the transition from reduced to oxidized zones in mangrove soils, sulfate and acidity are byproducts as is the formation of iron oxyhydroxides that reduce the solubility and migration of P in the soil column (such as across the sediment–water interface). Pyrite formation is iron limited in marine carbonate environments (Berner, 1984; Chambers et al., 2001) but iron is plentiful in coastal systems with clastic sediment inputs (Duarte et al., 1995). Analysis of sediment properties along four gradients of siltation in Southeast Asia demonstrated that silty types of habitats such as mangroves have higher contents of organic matter and nutrients (particularly N, P, and Fe) compared to more carbonate systems such as coral reefs and seagrass beds with higher calcium and lower organic matter (Kamp-Nielsen et al., 2002). Under these iron-limited conditions, hydrogen sulfide concentrations are high as byproduct of sulfate reduction rather than the iron sulfide minerals (Sherman et al., 1998). Without much Fe, P sorption to calcium carbonate minerals maintains low P concentrations in solution (Rosenfeld, 1979; Morse et al., 1985; Chambers et al., 2001). And under these circumstances primary producers are P limited. So in the presence of iron under anaerobic respiration there are two important reactions that affect regulator and resource gradients in mangrove soils: (1) iron reacts with sulfide to reduce hydrogen sulfide toxicity and P is released into solution of mangrove pore waters; and (2) in absence of iron, sulfate reduction forms hydrogen sulfide and P becomes bound to calcium carbonate minerals that remain insoluble under these conditions. So the multigradient model of resources could include iron as a key nutrient that will determine the level of regulators that
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control mangrove productivity. This interaction between resource (iron) and regulator (HS) is a function of hydroperiod from fringe to interior topographic gradients. And these complex mechanisms could explain the elevated biomass and net productivity of mangroves along muddy coasts with sufficient iron and reduced regulator concentrations of salinity and sulfide in contrast to carbonate coastal settings.
5. SOIL BIOGEOCHEMISTRY MODEL (NUMAN, M ODEL 4) Soil chemical and physical characteristics are significant constraints of forest structure and productivity as described in the multigradient model of mangrove NPP (Boto and Wellington, 1984b; Twilley, 1995; Alongi et al., 2002, 2004; Alongi and McKinnon, 2005). Moreover soil formation is an important process contributing to biogenic carbon sinks in tropical coastal regions (Twilley et al., 1992; Parkinson et al., 1994; Bouillon et al., 2008a). Soil formation in mangrove wetlands, as in other intertidal wetlands, is the combination of several ecological processes including organic matter production (above- and belowground components), export, decomposition, and burial; as well as sedimentation of allochthonous inorganic matter (Figure 5; Chen and Twilley, 1999a). Biogeochemical models that reconstruct sediment profiles can provide insights into the relative significance of these processes to the accumulation of organic matter and nutrients in wetland soils (Morris and Bowden, 1986). The NUMAN model is one of the only published biogeochemistry mangrove models that simulate organic matter, N and P concentrations with soil depth. Another model with some structural similarities to the NUMAN model is the “Relative Elevation Model” developed by Rybczyk et al. (1998; Chapter 30) for marshes in coastal Louisiana, which has been adapted by Cahoon et al. (2003) to estimate changes in total sediment height after major hurricane disturbances in mangrove forest of the Caribbean region. Sediments suspended in the water column are deposited in mangroves during flooding and this allochthonous material enriches mangrove soils. Mangroves are considered land builders due to the high sediment binding capacity of the root system (Scoffin, 1970). Sedimentation rates were estimated for sites in New World tropics using techniques develop by Lynch et al. (1989). Accumulation of organic matter was similar among all sites while accumulation of inorganic matter was higher in the forests influenced by river discharge. The contribution of inorganic material ranges from 133 to 5,151 g/m2/year with higher values occurring in riverine mangrove forests such as the Guayas River estuary, which represents a riverine type of coastal setting (Table 1). Those sites with sedimentation rates <500 g/m2/year include the fringe and basin sites of Rookery Bay and the lowest rate of 94 g/m2/year of sediment input is to a fringe carbonate site in Belize (Table 1). The fringe and basin sites at Terminos Lagoon have sedimentation rates of 1,094 and 618, respectively, while all of the riverine sites had sedimentation rates >1,200 g/m2/year (Table 1). All of the riverine sites in the Guayas River estuary had sedimentation rates >4,000 g/m2/year, more than double the riverine sites in
Table 1 Estimates of nutrient accumulation in different types of mangrove forests around the world Site description
Ecological type
Southwest Florida
Rookery Bay
Belize
West Pond Sittee River Estero Pargo
Terminos Lagoon
Boca Chica Ecuador
Gulf of Thailand Matang Forest
M1 M3 Sawi Bay
Inorganic sediments accumulation (g/m2/year)
Organic sediments accumulation (g/m2/year)
Interior Interior Interior Interior Fringe Fringe Interior Riverine Interior Interior Riverine Interior Riverine Riverine Riverine S1
239.19 184.57 141.59 172.26 307.82 173.16 93.64 1,286.43 744.80 386.77 1,440.93 1,136.54 3,581.60 4,318.88 3,076.41 7,500
246.83 227.48 280.17 272.05 54.78 100.29 276.19 307.92 348.93 231.66 400.76 145.51 803.35 813.94 997.47
MVR M18 M5
2,450 3,800
Total carbon accumulation (g/m2/year)
Tot nitogen accumulation (g/m2/year)
114.50 106.50 132.99 126.15 32.27 51.59 122.41 129.68 157.03 104.25 189.83 65.46 287.84 265.51 387.57 225.6
6.64 6.32 7.22 7.60 1.34 1.97 5.77 7.95 11.01 4.83 8.44 2.46 10.07 10.41 13.37 8.18
127 109.5 100.7
13.29 14.31 17.89
Tot phosphorus accumulation (g/m2/year) 0.23 0.25 0.24 0.21 0.27 0.62 0.16 1.06 0.63 0.21 0.82 0.73 3.01 3.62 0.92
IS:OM accumulation
0.97 0.81 0.51 0.63 5.62 1.73 0.34 4.18 2.13 1.67 3.60 7.81 4.46 5.31 3.08
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Belize or Terminos Lagoon, and 10-fold higher than the basin mangrove sites in Rookery Bay. The IS:OM ratio of inorganic (IS) and organic (OM) sedimentation explains the effects of geophysical processes on sedimentation patterns in mangroves. The ratios are <1 in the basin sites at Rookery Bay and the carbonate site in Belize. The overwash site at Rookery Bay and fringe and basin site in Terminos Lagoon have ratios that range from 1.7 to 2.1, while the subtidal fringe site has a ratio of 5.6. All the riverine sites have ratios that range from 3.6 to 7.8 (Table 1). Total N accumulation ranged from 1.34 g/m2/year in the subtidal fringe in Rookery Bay to 11.01 g/m2/year in the fringe forest at Terminos Lagoon (Table 1). There was no clear pattern among ecological types or geographical locations. The accumulation of N among the riverine and basin forests was similar at about 5.5 g/m2/year, which is higher than the range of denitrification and N fixation described below. Thus N storage in sediment soils is an important fate of N in mangrove ecosystems. The accumulation of carbon (59–185 g/m2/year) and N (1.55–5.80 g/m2/year) were associated with deposition of organic matter and rates were similar among sites. Atomic C:N ratios of accumulated material at the riverine sites were >30, while sites with less riverine input had C:N ratios <20. Accumulation of P ranged from 0.11 to 0.78 g/m2/year and the higher rates occurred at sites with high inorganic matter loading. The riverine forest in Terminos Lagoon receives increased inputs of inorganic sediment and P from the Palizada River, whereas sites dominated by tides accumulate a greater proportion of organic matter and N. The elevated phosphate input rates into riverine mangrove sites are associated with higher levels of litter productivity compared to tidal mangroves that have less P input and lower productivity. All of the riverine sites had TP accumulation rates >0.7 g/m2/year, with a range from 0.73 to 3.62 g/m2/year (Table 1). All of the higher rates (>1.0 g/m2/year) were at the riverine sites in Guayas River estuary and Belize, whereas the rates in the riverine sites at Terminos Lagoon were 0.82 and 0.73 g/m2/year for the fringe and interior site, respectively. Sedimentation and nutrient burial, particularly carbon and N, in mangrove wetlands not only include allochthonous inorganic matter input (sedimentation), but net organic matter input resulting from high rates of root production relative to decomposition (Chen and Twilley, 1999a). Several studies have found that decay of belowground material is slower than leaf litter (Hackney and de la Cruz, 1980; van der Valk and Attiwall, 1984; McKee and Faulkner, 2000; Middleton and McKee, 2001; Poret et al., 2007). A large part of sedimentary organic matter in mangroves is derived from root organic matter (Alongi et al., 2001) and in many forest systems can be the principle source of organic matter in the deeper soil layers (Ludovici et al., 2002). For example, deposition and slow degradation of mangrove roots may contribute more to organic matter accumulation and vertical building of mangrove islands in Belize than total litter fall (Middleton and McKee, 2001; McKee et al., 2007). In carbonate settings, like those of south Florida, belowground peat production is the primary control of sediment accretion (Lynch et al., 1989; Parkinson et al., 1994). Root decomposition of mangroves in south Florida (Poret et al., 2007) does not seem to follow the predictable pattern of higher rates with decreased C:N ratios and lower lignin:N ratios as observed for other root degradation studies (McClaugherty et al., 1982, 1984; Richert et al., 2000; Scheffer and Aerts, 2000).
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However, cotton strip assays found that belowground decomposition of mangroves in a reef system off the coast of Belize increased with increasing nutrient availability (Feller et al., 1999); and P fertilization caused a dramatic increase in belowground decomposition (Feller et al., 2002). Results from the FCE region point to greater control of belowground processes by the particular environmental setting of each site rather than the chemical components of the roots. This suggests that differences in hydroperiod, soil fertility, tidal regime, topography, and other microscale changes in sites could have a greater effect on decomposition than the plant material itself (Poret et al., 2007).
6. M ASS BALANCE EXCHANGE (MODEL 5) The processes associated with nutrient accumulation in mangroves soils described in the NUMAN model (Figure 5) above can also be viewed as residual fluxes that remain following mass balance exchange of nutrients with atmosphere and/or coastal waters. This section will focus on nutrient biogeochemistry of mangrove wetlands as a nutrient source, sink, or transformer depending on the net flux across two mangrove boundaries including the atmosphere and tidal exchange (Figure 6; using N as example). Several hypotheses establish estuarine exchange as strictly dependent on the geophysical properties of an estuary, while others focus on the role of biological factors within the mangrove controlling the fate of organic matter and nutrients. As described in Figure 1, the type of exchange (strong outwelling vs. bidirectional flux) is associated with the river to tidal gradient established as focal point in this review. Defining the relative influence of ecological and geophysical processes associated with nutrient exchange in mangrove wetlands is an excellent system to define how ecogeomorphology controls the nutrient dynamics of coastal landscapes.
6.1. CO2 efflux from mangrove sediments and tidal waters Total mineralization based on CO2 fluxes from mangrove sediments is available for a wide range of mangrove systems (Bouillon et al., 2008a). Mangrove creek waters have consistently been found to show high CO2 oversaturation, and hence, are a net source of CO2 to the atmosphere, with an overall average of 59 + 52 mmol/m2/day (n = 21) (references from Bouillon et al., 2008a include Ghosh et al., 1987; Ovalle et al., 1990; Millero et al., 2001; Borges et al., 2003; Bouillon et al., 2003, 2007a–c; Biswas et al., 2004). Dark fluxes from sediments average of 61 + 46 mmol/m2/day (n = 82), while about half of the available flux data under light conditions show a net CO2 uptake of –15 + 54 mmol/m2/day (n = 14). These sediment and water column estimates relate only to net CO2 fluxes, and not to overall mineralization rates (see discussion below). Upscaling CO2 fluxes for sediments and the water column separately is somewhat problematic, since the relative surface areas in these intertidal systems shift during tidal events. Bouillon et al. (2008a) summarized that the similar magnitude in CO2 efflux from both sediments and water column results
Ecogeomorphic Models of Nutrient Biogeochemistry
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in overall CO2 efflux from mangrove systems to be about 60 + 45 mmol/m2/day (ignoring the CO2 efflux under light conditions). These CO2 emission rates are grossly underestimated partially due to the lateral transport of dissolved inorganic carbon (DIC) resulting from mineralization of organic rich pore waters that drain from mangrove wetlands (Bouillon et al., 2008b). Recent analysis of data sets consistently shows that DIC in mangrove creeks far exceeds dissolved organic carbon (DOC), by a factor of 3–10 (Bouillon et al., 2008a). Under the assumption that both originate mainly from the tidal exchange and therefore follow the same tidal variations (Bouillon et al., 2007c), this implies that DIC export exceeds DOC export to the same degree. Crab burrows can greatly enhance the surface area of the sediment–air or sediment–water interface where exchange of CO2 or DIC can take place, serving as significant conduits for enhancing CO2 exchange between mangrove sediments and atmosphere (Kristensen, 2008). These updated estimates of carbon flux from sediments and waters of mangroves have implications to the debate on how coastal ecosystems influence the global carbon budget (Twilley et al., 1992; Saenger and Snedaker, 1993; Lee, 1995; Duarte et al., 2005; Jennerjahn and Ittekkot, 2002). A reassessment of global mangrove carbon budgets by Bouillon et al. (2008a) estimates that more than 50% of the carbon fixed by mangrove vegetation, estimated 217 + 72 Tg C/year, appears to be unaccounted for based on estimates of various carbon sinks (organic carbon export, sediment burial, and mineralization). This missing carbon sink is conservatively estimated at 112 + 85 Tg C/year, equivalent in magnitude to 30–40% of the global riverine organic carbon input to the coastal zone. The analysis above suggests that inorganic carbon flux from sediments and mangrove waters is severely underestimated, and that the majority of carbon export from mangroves to adjacent waters occurs as DIC. Using the average rate of DIC flux above, global levels of CO2 efflux from both sediments and the water column can be estimated at 42 + 31 Tg C/year. This analysis suggests that mineralization and export of carbon as DIC is a quantitatively important pathway (178 + 165 Tg C/year; Bouillon et al., 2008a). The magnitude of this process could be similar to that of the missing carbon sink, and may vary in range among different mangrove systems from muddy coasts to carbonate settings.
6.2. The balance of N fixation and denitrification The net exchange of N gas in mangrove ecosystems depends on the inputs of N via fixation relative to the loss via denitrification. The fixation of N depends on organisms that utilize the nitrogenase enzyme in specific environments. The flux of N gas out of mangrove ecosystems is associated with nitrate reduction as part of organic respiration in the reduction zone known as denitrification. Thus the distribution of oxidized and reduced zones, as described in the geochemical model, together with the supply of oxidized form of N as NO3, will determine the net balance of this nutrient in mangrove forests. N fixation represents the “new” source of N to the system and is the only process that can compensate for N removal by denitrification (Howarth et al., 1988). Nitrogenase activity has been observed in decomposing leaves, root surfaces
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(prop roots and pneumatophores), and sediment, but few of these studies have interpreted these rates relative to the N budget of mangrove forests (Tables 2 and 3). These studies have shown that decaying mangrove leaves are sites of particularly high rates of fixation, and thus may account for 15–64% of N immobilization during decomposition of mangrove leaf litter on the forest floor (Gotto et al., 1981; van der Valk and Attiwill, 1984; Woitchik et al., 1997). Sediment and roots fix N in range from nondetectable to about 15 nmol N/g of substrate/h, rates for leaf litter are common >250 nmol N/g of substrate/h (Table 2). Pelegri et al., (1997) estimated that N fixation could account for 45–100% of the total N immobilized in leaf litter (1–8 mg N/dry wt N enrichment) in the Shark River mangrove sites of the FCE. However, N fixation only could supply 7% (8.3 mg N/m2/day) of the N required (53 mg N/m2/day) for mangrove growth. Most of the studies of N fixation in mangroves average around 50 mmol/m2/h, which is about 0.4% of the NPP of mangroves (Bouillon et al., 2008a). Results from mangrove sediments in south Florida indicate that N fixation rates range from 0.4 to 3.2 g N/m2/year (Kimball and Teas, 1975; Zuberer and Silver, 1978), similar for the natural rates of denitrification. However, the spatial analysis of N fixation is still inadequate to provide a clear estimate of this contribution to the N budget of mangrove wetlands (Twilley, 1997; Rivera-Monroy et al., 1999). Sponges, tunicates, and a variety of other forms of epibionts on prop roots of mangroves are highly diverse (Sutherland, 1980; Ru¨tzler and Feller, 1988; Ellison and Farnsworth, 1992), especially along carbonate shorelines with little terrigenous input. The diversity and biomass of these communities and associated ecological functions may be limited to specific geomorphologic types that are protected from turbid waters. There are a few studies on ecosystem function that indicate that these communities are sites of N fixation that influence the N budget of mangrove canopies. There is evidence that epibionts on prop roots may be a source of nutrition for higher-level predators as well as influencing various processes in mangrove fringe forests. These processes of nutrient regeneration associated with sponge communities that colonize aerial root systems of mangroves have received comparatively little attention; but they may influence the productivity of fringe mangrove forests, as well as enhance the exchange of nutrients with coastal waters (Ellison et al., 1996). The specific contribution of these productive and diverse epibiont communities in predominately carbonate environments may demonstrate an important linkage between biodiversity and ecosystem function. Denitrification is primarily dependent upon presence of a reduction zone, an energy source, and availability of NO3 source. Based on the sources of NO3 , there are two types of denitrification; direct denitrification that is fueled by NO3 diffusing from the overlying water column into sediments, and coupled nitrification– denitrification that is supported by NO3 from nitrification in the sediments (Nielsen, 1992). The contributions of these two NO3 sources are regulated by different mechanisms. Direct denitrification is typically a linear function of the NO3 concentration in the overlying water and an inverse linear function of the oxygen penetration depth in the sediments (Christensen et al., 1990; Nielsen et al., 1990). Coupled nitrification–denitrification is regulated by the rate and position of
Substrate
Rate (nmol N/g substrate/h)
Source
Tampa Bay (USA)
Sediment (Avicennia germinans, Laguncularia racemosa, Rhizophora mangle) Sediment (Avicennia marina)
0.14a to 0.61a
Zuberer and Silver (1978)
5.3 + 4.1b,c 0a,c to 20.9a,c 0.07a,c to 0.27a,c
Pelegri and Twilley (1998)
Tampa Bay (USA) Tampa Bay (USA) Tampa Bay (USA) West Port Bay (Australia)
Sediment (Avicennia germinans, Laguncularia racemosa, Rhizophora mangle) Sediment (Avicennia germinans, Laguncularia racemosa, Rhizophora mangle) Roots (Rhizophora mangle) Roots (Laguncularia racemosa) Roots (Avicennia germinans) Roots (Avicennia marina)
van der Valk and Attiwill (1984) Pelegri et al. (1997)
Florida Coastal Everglades (USA) Makham Bay (Thailand) Key Biscayne (USA)
Pneumatophores (Avicennia germinans) Roots (Rhizophora apiculata) Leaf litter (Rhizophora mangle)
Zuberer and Silver (1978) Zuberer and Silver (1978) Zuberer and Silver (1978) van der Valk and Attiwill (1984) Pelegri et al. (1997) Kristensen et al. (1998) Gotto and Taylor (1976)
Tampa Bay (USA) West Port Bay (Australia)
Leaf litter (Rhizophora mangle) Leaf litter (Avicennia marina)
Auckland Province (New Zealand) Gazi Bay (Kenya) Gazi Bay (Kenya) Florida Coastal Everglades (USA) Florida Coastal Everglades (USA)
Leaf litter (Avicennia marina)
3c,d to 10c,d 7c,d to 35c,d 9c,d to 11c,d 2.3b,c,e,f and 7.5 + 3.9b,c,e,g,h 0a,c to 3.2a,c 11 + 4a,c,e 50.0 + 50.0b,c,e to 900.0 + 242.9a,d,e 23b and 25a 235 + 37b,c,e,h and 313 + 142b,c,h 7.4 + 4.1 to 27.6 + 5.5
Zuberer and Silver (1978) van der Valk and Attiwill (1984) Hicks and Silvester (1985)
250b,h,i and 780b,h,j 156b,h,i and 760b,h,j 44.1a,c,h 43.1a,c,h
Woitchik et al. (1997) Woitchik et al. (1997) Pelegri and Twilley (1998) Pelegri and Twilley (1998)
West Port Bay (Australia) Florida Coastal Everglades (USA) Florida Coastal Everglades (USA)
Leaf litter Leaf litter Leaf litter Leaf litter
(Rhizophora mucronata) (Ceriops tagal) (Rhizophora mangle) (Avicennia germinans)
665
Site (country)
Ecogeomorphic Models of Nutrient Biogeochemistry
Table 2 Nitrogen fixation rates for different types of ecological substrates incubated under different types of experimental conditions in mangrove forests around the world
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Table 2
(Continued )
Site (country)
Substrate
Rate (nmol N/g substrate/h)
Source
Florida Coastal Everglades (USA) Florida Coastal Everglades (USA) Florida Coastal Everglades (USA) Iriomote Island (Okinawa)
Leaf litter (Rhizophora mangle) Leaf litter (Laguncularia racemosa) Leaf litter (Avicennia germinans) Warty lenticellate bark (Bruguiera gymnorrhiza) CPOMk (Avicennia marina)
4.9a,c to 149.0a,c 22.5a,c to 222.9a,c 17a,c to 359.9a,c 5b,c to 8b,c
Pelegri et al. (1997) Pelegri et al. (1997) Pelegri et al. (1997) Uchino et al. (1984)
20.5
Hicks and Silvester (1985)
Decomposing logs
0.63 + 0.30l to 4.46l
Boto and Robertson (1990)
Auckland Province (New Zealand) Missionary Bay (Australia)
Robert R. Twilley and Victor H. Rivera-Monroy
Original rates were converted to nmol of N based on 3:1 ratio of ethylene to mole of N2. All rates were obtained using the acetylene reduction technique. Rates originally reported as moles C2H4 produced were converted to moles of N2 using the theoretical 3:1 ratio. An effort was made to present the range of rates measured in each study. If study examined the rate of N fixation during decomposition of a substrate (i.e., leaves), the maximum rate was presented. If study measured a rate of 0, this was reported. The exception was if the study looked at rates of N fixation during decomposition (see note above). An effort was also made to keep track of the conditions of the incubations (i.e., light vs. dark, aerobic vs. anaerobic). If the conditions were unclear, then they were not specified. a Anaerobic. b Aerobic. c Dark. d Light. e Inundated. f Live. g Dead. h Maximum. i Dry season. j Rainy season. k Coarse particulate organic matter. l 24 h.
Nitrogen fixation rates per unit area of mangrove forest based on experimental conditions in mangrove forests around the world
Site (country)
Substrate
Rate (mmol N/m2/h) a
a
Source
0.3 to 33 0b to 100.0b
Hicks and Silvester (1985) Nedwell et al. (1994)
Missionary Bay (Australia)
Sediment (Avicennia marina) Sediment (Rhizophora mangle, Avicennia germinans) Sediments (Rhizophora spp., Ceriops spp.)
0c to 9.9 + 1.8c
Joyuda Lagoon (Puerto Rico)
Sediment (not specified)
13.3d to 31.5d
Makham Bay (Thailand) Mekong Delta (Vietnam)
Sediment (Rhizophora apiculata) Sediment (Rhizophora apiculata)
Mazizini (Tanzania)
Sediment (Avicennia marina, Sonneratia alba)
12 + 0.4b,d,e 20 + 11a,e to 119 + 39a,e 8.7a to 112.0a
Boto and Robertson (1990) Morell and Corredor (1993) Kristensen et al. (1998) Alongi et al. (2000)
Sawi Bay (Thailand)
Sediment (Avicennia alba, Ceriops decandra, Rhizophora apiculata) Sediment (Rhizophora apiculata)
Auckland Province (New Zealand) Oyster Bay (Jamaica)
Matang Mangrove Forest Reserve (Malaysia) Kisakasaka (Tanzania) Jiulongjiang Estuary (China)
Sediment (Avicennia marina, Bruguiera gymnorrhiza, Ceriops tagal, Rhizophora mucronata, Sonneratia alba) Sediment (Kandelia candel, Aegiceras corniculatum)
0a,e,f to 24 + 24a,e,f
Ecogeomorphic Models of Nutrient Biogeochemistry
Table 3
Lugomela and Bergman (2002) Alongi et al. (2002)
0 + 0a,e,f to 250.0 + 108.3a,e,f 1.3a,c to 41.3a,c
Alongi et al. (2004)
0a,e,f and 4 + 4a,e,f
Alongi et al. (2005)
Sjo¨ling et al. (2005)
667
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Table 3
(Continued )
Site (country)
Substrate
Rate (mmol N/m2/h)
Source
Twin Cays (Belize)
Sediment (Avicennia germinans, Laguncularia racemosa, Rhizophora mangle) Pneumatophores (Avicennia marina, Sonneratia alba)
0a to 100.8a
Lee and Joye (2006)
40a to 790.0a
Lugomela and Bergman (2002)
Mazizini (Tanzania)
Robert R. Twilley and Victor H. Rivera-Monroy
Original rates were converted to nmol of N based on 3:1 ratio of ethylene to mole of N2. All rates were obtained using the acetylene reduction technique. Rates originally reported as moles C2H4 produced were converted to moles of N2 using the theoretical 3:1 ratio. An effort was made to present the range of rates measured in each study. If study examined the rate of N fixation during decomposition of a substrate (i.e., leaves), the maximum rate was presented. If study measured a rate of 0, this was reported. The exception was if the study looked at rates of N fixation during decomposition (see note above). An effort was also made to keep track of the conditions of the incubations (i.e., light vs. dark, aerobic vs. anaerobic). If the conditions were unclear, then they were not specified. a Aerobic. b Inundated. c 24 h. d Anaerobic. e Dark. f Light.
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669
nitrification activity, which in turn, is regulated by the nitrification capacity and the availability of ammonium and oxygen, and generally is enhanced by deeper oxygen penetration (Henriksen and Kemp, 1988; Blackburn et al., 1994). The difference between these two regulating mechanisms establishes the significance of the interface between the oxidized and reduction zone of mangrove soils described above in the geochemical model (Figure 3). For instance, a decrease in oxygen supply to mangrove soils may either stimulate denitrification if the water column is the major source of NO3 or inhibit denitrification if NO3 produced by nitrification is the major source of NO3 (Rysgaard et al., 1994). Coupled nitrification–denitrification is thought to be quantitatively more important than influx of NO3 in most aquatic sediments (Christensen et al., 1990). Another potentially important N sink in mangrove sediments is anaerobic ammonium oxidation (anammox) (Francis et al., 2007; Hayatsu et al., 2008) þ (Figure 6). Anammox is the anaerobic conversion of NO 2 and NH4 to N2, and although its presence was suggested as early as 1965, first direct evidence for this reaction was documented more than a decade ago (Kartal et al., 2007). Several studies have reported the presence of anammox in estuarine and offshore sediments, in permanently anoxic bodies of water, and in multiyear sea ice (Jetten et al., 2003; Thamdrup et al., 2004, 2006; Dalsgaard et al., 2005; Jensen et al., 2007). Anammox could be an important pathway in global N cycling, since it can account for as much as 67% of benthic N2 production, with the remainder being produced by denitrification as determined in oceanic and coastal N budgets (Meyer et al., 2005; Schmid et al., 2007). However, estimates of anammox rates in mangrove sediments are scarce. The only study in a subtropical mangrove forest in Australia indicates that anammox to sediment N2 production in subtidal sediments was relatively low (0–9%) and comparable to other temperate estuaries (Meyer et al., 2005). Yet, more studies are needed in tropical latitudes to further understand the significance of anammox in the N cycle of mangrove-dominated ecosystems. Mangrove sediments have a high potential for the removal of N from surface waters with a large range in denitrification estimates (Table 4). A general average rate of denitrification among mangrove sediments is about 180 mmol/m2/h, about four times that of N fixation (cf. Tables 3 and 4). This is similar to the medium range of rates measured in mangrove sediments described by Nedwell (1975). Estimates of denitrification based on NO3 uptake range from a low of 0.53 mmol/m2/h at Hinchinbrook Island (Iizumi, 1986) where concentrations are <10 mM, compared to ranges of 9.7–261 mmol N/m2/h in mangrove forests receiving effluents from sewage treatment plants with concentrations >1,000 mM (Nedwell, 1975; Corredor and Morell, 1994). Measures of denitrification using small amendments of 15NO3 followed by direct measures of 15N2 production have shown that denitrification accounts for <10% of the applied isotope suggesting that NO3 uptake is not a sink of N (Rivera-Monroy et al., 1995b). Rates of denitrification linked to nitrification in mangrove sediments are also low and thus little exchange of N gas can be associated with disssolved inorganic nitrogen (DIN) uptake (Rivera-Monroy and Twilley, 1996). The lack of 15N2 gas production indicates that much of the net DIN exchange at the boundary of mangroves described above may not be loss to the atmosphere via denitrification, but
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Table 4 Denitrificaiton rates using different techniques and various experimental conditions in mangrove forests around the world Site (country)
Mangrove type
Techniques
Incubations
Rate (mmol N/m2/h)
Source
Joyuda Lagoon (Puerto Rico) Oyster Bay (Jamaica) La Parguera (Puerto Rico)
Fringe
A
0.74 to 160.9
A A
Morell and Corredor (1993) Nedwell et al., (1994) Corredor and Morell (1994) Corredor et al. (1999)
Terminos Lagoon (Mexico)
Fringe, center, reara,b Fringe receiving sewage effluenta–c Fringe receiving sewage effluenta–c Fringe, transition, dwarf a–c Fringe, basin, riverinea–c
Dark, flooded, anaerobic Flooded Dark, flooded, anaerobic Dark, flooded, anaerobic Diel, aerobic
0.08 to 250
Makham Bay (Thailand)
Mid-intertiald
B
Hinchinbrook Channel (Australia) Mekong Delta (Vietnam)
Matured–f
C
Managed, high-intertidald
C
Sawi Bay (Thailand)
Managed, mid-, and high-intertidald,g,h Managedd,f,i,j
C
Low-, mid-, highintertidalk,l
C
Dark, flooded, aerobic Dark, flooded, aerobic Flooded, aerobic Flooded, aerobic Dark, flooded, aerobic Dark, flooded, aerobic Dark, flooded, aerobic
La Parguera (Puerto Rico) Twin Cays (Belize)
Techniques: A = acetylene blockage; B = isotope enrichment; C = N2 gas flux. a Rhizophora mangle b Avicennia germinans c Laguncularia racemosa d Rhizophora apiculata e Avicennia alba f Ceriops decandra g Kandelia candel h Rhizophora stylosa i Bruguiera gymnorrhiza j Aegiceras corniculatum k Avicennia marina l Acrostichium aureum
A B
C
49.9 + 0.3 to 89.9 + 83 0 to 92.5
1.9 + 0.4
Lee and Joye (2006) Rivera-Monroy and Twilley (1996) Kristensen et al. (1998)
342 + 264 and 460 + 266 0 and 183.3 + 41.7
Alongi et al. (1999)
0 to 160 + 300
Alongi et al. (2002)
33.3 + 50.0 to 916.7 + 1,066.7 92 + 53 to 315 + 186
Alongi et al. (2004)
Alongi et al. (2000)
Alongi et al. (2005)
Robert R. Twilley and Victor H. Rivera-Monroy
Matang Mangrove Forest Reserve (Malaysia) Jiulongjiang Estuary (China)
A
0 to 83.3 9.7 to 183.0
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671
accumulated in the litter on the forest floor. However, recent measurements of direct N2 release from mangrove sediments suggest that more of this NO3 may flux to the atmosphere than suggested in the studies from Mexico (Terminos Lagoon). The N2 gas flux estimates are commonly above 180 mmol/m2/h, with some rates >900 mmol/m2/h (Table 4). A more comprehensive comparison on fixation and denitrification is needed to balance the N budget relative to carbon budgets for mangroves discussed above. Given the general observation that denitrification >N fixation, N recycling must meet most of the demand of NPP in mangrove forests.
6.3. Tidal exchange In addition to the lack of flux studies, methodological limitations have also hampered comparison among mangrove systems since material exchange is actually occurring at different spatial scales. In fact, a general consensus on nutrient outwelling from mangroves to coastal waters has not yet been reached due in part to inconsistencies in the published data as result of methodological differences (Dittmar and Lara, 2001b). Historically, material exchange in mangrove-dominated systems has been evaluated using a “Eulerian” approach (Boto and Wellington, 1988; Boto and Bunt, 1981; Wattayakorn et al., 1990) where material fluxes are estimated along tidal creeks and embayments over several tidal cycles. This approach allows the estimation of annual exchanges between large mangrove areas and the estuary through the product of water discharge and material concentration over a tidal cycle to obtain net fluxes. However, results from these studies integrate biogeochemical processes in the forest, tidal creeks, and the coastal boundary system thus making it difficult to separate the effect of the mangrove forest from the water column on these nutrient exchanges. Evaluating material fluxes across the boundary of mangrove forest with tidal channels and coastal waters requires the implementation of studies at different spatial scales. Carbon export from mangrove ecosystems ranges from 1.86 to 401 gC/m2/year (Twilley et al., 1992), with an average rate of about 210 gC/m2/year (Table 3). Carbon export from mangrove wetlands is nearly double the rate of average carbon export from salt marshes (Nixon, 1980), which may be associated with the more buoyant mangrove leaf litter, higher precipitation in tropical wetlands, and greater tidal amplitude in mangrove systems studied (Twilley, 1988). The fate of mangrove primary production has been a major topic of debate in the literature during the past decades (Bouillon et al., 2008a). In particular, the “outwelling” hypothesis, first proposed by Odum and Heald (1972) suggested that a large fraction of the organic matter produced by mangrove trees is exported to the coastal ocean, where it would form the basis of a detritus food chain and thereby support coastal fisheries. Despite the large number of case studies dealing with various aspects of organic matter cycling in mangrove systems (Kristensen et al., 2008), there is still no consensus on overall mangrove primary production and the ecological fate of the organic matter produced. Several authors have suggested that mangrove-derived organic matter is of global significance in the coastal zone (Robertson and Alongi, 1995; Schlu¨nz and Schneider, 2000; Dittmar and Lara, 2001a). Estimates indicate that mangrove forests could be responsible for 10% of the global export of
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terrestrial particulate and dissolved organic carbon (POC and DOC) to the coastal zone (Jennerjahn and Ittekkot, 2002 and Dittmar et al., 2006, respectively), and for 10% of the global organic carbon burial along with seagrasses in the coastal ocean (Duarte et al., 2005). The exchange of carbon between tidal wetlands such as mangrove forests or salt marshes and the coastal ocean, and its ultimate fate in the ocean is therefore increasingly recognized as a potentially important component in the ocean carbon budget (Twilley et al., 1992). This may particularly be evident in river-dominated mangrove systems such as muddy coasts and deltas, where organic material exchanged is greater at the boundary of the forests compared to other coastal settings (Twilley, 1985). Surveys of N exchange demonstrate some of the principles of determining the function of mangrove wetlands as a nutrient sink and transformer of some nutrient species (Figure 6). Export and import flux studies of estuarine coasts in Mexico (Rivera-Monroy et al., 1995a) and Australia (Boto and Wellington, 1988) observed an import of NH4 compared to slight release of NO3 in Australia and import in Mexico. One of the main differences between the two sites is the flux of DON, which is imported into the wetlands in Australia and exported from the site in Mexico. The largest N flux at both sites is export of particulate nitrogen, which coincides with net flux of organic carbon from most mangroves (described above). Compared to other flux studies of mangroves, there seems to be a pattern of net inorganic fluxes into the wetlands and corresponding flux of organic nutrients out (Rivera-Monroy et al., 1995a). Given large statistical variation around these flux estimates, the best conclusion may be that mangrove wetlands transform the tidal import of inorganic nutrients into organic nutrients that are then exported to coastal waters. Small-scale estimates of nutrient fluxes within the (<20 m2) carbonate zone (Taylor Slough) of the FCE (Davis et al., 2001a,b, 2003, 2004) showed there was consistent uptake of ammonium (6.6–31.4 mmol/m2/h) and of total N (98–502 mmol/m2/h), whereas oxidized forms of inorganic N (7.1–139.5 mmol/ m2/h) was released to the water column (Davis et al., 2001a). Mangroves in both the sediment-rich (Terminos Lagoon) and sediment-poor (Taylor River) sites export organic nutrients to adjacent coastal waters; the oligotrophic system with diverse prop root communities release nitrate whereas the sediment dominated site removes nitrate from overlying water.
7. CONTRASTING C OASTAL SETTINGS AND B IOGEOCHEMICAL MODELS We used five models to capture the mechanisms associated with nutrient biogeochemistry across different geomorphological settings that can be subdivided by source of sediment and relative levels of geophysical processes, particularly river and tides. This presentation was to better describe the hypothesis that geomorphological settings dominate the diversity of mangrove ecosystem function across coastal landforms, with feedback effects on soil properties that contribute to evolution of morphodynamics of coastal settings. We will compare and contrast
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a few sites across key elements of the five models to develop some summary questions that may be of interest in further developing this ecogeomorphology approach to mangrove biogeochemistry. The FCE represent mangroves in contrasting basins with very different sediment inputs representing the significance of clastic versus biogenic sediment source that exists on a carbonate platform. Mangroves at the mouth of Shark River estuary have allochthonous sediments rich in P imported from the coastal shelf deposited during storm events (SRS5 and SRS6) compared to more inland mangroves that receive much less sediment input (Chen and Twilley, 1999b). Allochthonous sediment input results in higher P concentrations and lower N:P ratios compared to mangroves farther upstream, resulting in much greater forest biomass and NPP near the mouth of the estuary. In contrast, mangroves along Taylor Slough receive much less sediment during storm events due to a geologic ridge that isolates the interior mangrove forests from Florida Bay and is more remote from P-enriched shelf sediments. This isolation results in biogenic sediments developed above calcareous substrate with much lower P concentrations resulting in scrub mangrove formations (Davis et al., 2004; Ewe et al., 2006). These two adjacent geomorphological settings with the same catchment inputs differ in mangrove productivity and community structure largely due to sediment source. The two basins also have different inundation periods, which are much longer in Taylor Slough resulting in lower redox conditions that may also contribute to patterns in nutrient biogeochemistry and forest structure of the two basins. The two basins contrast in both phosphorus resource and hydroperiod, with anoxic zones existing nearly to the surface during most of the year in Taylor Slough. Yet the iron concentrations are similar among these two transects that may explain the relative low concentrations of sulfide, even though they have different sources of sediment (Chambers and Pederson, 2006). The biogeochemistry of Taylor Slough is very similar to mangroves growing on the carbonate reef system in Belize, where P limits interior forest structure and productivity, and root production is critical to accretion and carbon accumulation. The low nutrient concentration in soils of Taylor Slough results in high amounts of root biomass relative to aboveground, resulting in accretion rates and storage of carbon and N that are similar to carbonate-based systems of Belize with low inorganic to organic sediment ratios. The fringe mangroves of both systems have clear waters lacking suspended sediment that inundate prop roots daily, resulting in a vast community of epibionts and episediment that dominate nutrient biogeochemistry. The importance of these ecological communities on nutrient biogeochemistry in carbonate settings has not been fully described. But autotrophic processes in systems with plenty of light (low sediment input) but scarce nutrient resources may dominate biogeochemistry and patterns of nutrient dynamics. A geochemical model just for carbonate sediment systems with biogenic sediments may be unique to the geochemical model described in this review for estuarine coasts. The recent passage of Hurricane Wilma in 2005 demonstrated that both sediment and P input were associated only with sites along the mouth of Shark River estuary; in contrast to very little sediment input to mangrove sites along Taylor
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River Slough. Thus an actual hurricane event corroborated historical evidence of this fertilization mechanism in soil cores over the last 100 years. So the presence of clastic sediments in Shark River estuary from the shelf is a key process to describe the landscape pattern of mangrove forest productivity and nutrient biogeochemistry. However, sediment-rich sites at the mouth of Shark River estuary in FCE are very different to mangroves found along the muddy coast of French Guiana (Fabre et al., 1999). The phosphate concentrations in muddy coast mangroves are high and mostly in organic form (52–58%) in fringe and interior mangroves. P concentration range from 0.638 to 0.804 mg/g dry mass in pioneer and mixed mangroves, respectively; higher than the concentrations in Shark River estuary, and much higher than Taylor and Belize carbonate systems. And extractable iron concentrations vary from 10.3 to 15 mg/g dry mass; much higher than the FCE mangroves. The nutrient biogeochemistry of carbonate settings that are biogenic (Taylor Slough and sites in Twin Cays, Belize) compared to sites from a muddy coast demonstrates the complex patterns that are associated with the delta–estuarine– carbonate continuum of geomorphological settings. The iron and P associated with clastic sediment systems along deltaic coasts, together with high-energy mixing of soils in fringe versus interior forests, results in much higher levels of mangrove net productivity and forest structure. The interior sites of mangroves along high-energy deltaic coasts may have more anoxic conditions that can promote P solubility and reduce sulfide toxicity with formation of pyrites. This interaction of sediment source and redox zones on resource and regulator availability needs more clarity across geomorphological settings. And these zonation patterns from fringe to interior sites can be obscured in wet tropical climates, compared to dry climates that have strong seasonal cycles of water levels and thus redox zones. These differences are highlighted in the conditions described in models 2 to 5 presented in this review that promote mangrove ecological productivity, and thus help define the links between geomorphological setting and mangrove ecosystem function. Thus the hierarchy of ecogeomorphic settings in Figure 2, and the mechanisms described in this chapter, explain some of the diverse patterns in nutrient biogeochemistry in mangrove ecosystems. But more comparisons of sites are needed to build a more complete model of biogeochemical processes. Finally, one of the early connections between geomorphological settings and ecological function of mangroves was associated with outwelling concept. As described in Figure 1 the relative process of river and tides influences the flux of organic matter and nutrients across the mangrove boundary with the coastal ocean. There has been significant advance in recent years on the carbon dynamics associated with this flux, particularly the export of inorganic carbon to coastal oceans along deltaic coasts. Yet there has been very little improvement in understanding the dynamics of N, across both coastal waters and atmosphere. The accumulation of N in mangrove soils follows the rate for carbon. But it is not clear if N outwelling from mangroves in deltaic and estuarine coasts has similar correlation. One of the key missing links in this analysis is the fate of dissolved organic N in waters exporting from mangrove forests. There is evidence that inorganic N regeneration (ammonification) is correlated with P availability, which may influence the availability of N for NPP. But there is no clear pattern on the relative rates of N fixation
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as source of N for new production of mangrove biomass. And the net exchange of N with the atmosphere as a consequence of denitrification rates is still a missing part of our biogeochemical models. Given the huge impact of humans on the global N cycle, and particularly eutrophication of coastal waters, such a gap in our models should be a focus over the next decade. We may find some very interesting patterns across the muddy–estuarine–carbonate continuum with how N is processed by mangrove ecosystems.
ACKNOWLEDGMENTS The authors want to thank Kelly Henry and Edward Castan˜eda, graduate students at LSU in the Department of Oceanography and Coastal Sciences, for developing the excellent review of N fixation and denitrification results for mangrove ecosystems described in this review. We would also like to thank Steve Davis of Texas A&M University, Mark Brinson of East Carolina University, and Don Cahoon of USGS for very helpful suggestions to improve this manuscript. The authors were supported by the National Science Foundation under Grant No. DBI-0620409 and Grant No. DEB-9910514 (Florida Coastal Everglades, Long-Term Ecological Research). The development of ecogeomorphology concept was supported by the STC program of the National Science Foundation via the National Center for Earth-surface Dynamics under the agreement Number EAR-0120914. Any opinions, findings, conclusions, or recommendations expressed in the material are those of the authors and do not necessarily reflect the views of the National Science Foundation.
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P A R T
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COASTAL WETLAND RESTORATION AND MANAGEMENT
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C H A P T E R
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S EAGRASS R ESTORATION Eric I. Paling, Mark Fonseca, Marieke M. van Katwijk, and Mike van Keulen
Contents 1. Introduction 2. Regional Activities 2.1. Europe 2.2. Australia 2.3. Oceania 2.4. Southeast Asia 2.5. China and Japan 2.6. New Zealand and the Pacific Islands 2.7. United States 3. Policy Issues Relevant to Mitigation 3.1. Costs of restoration 3.2. Valuation of ecosystem services 4. Conclusions Acknowledgments References
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1. INTRODUCTION It is useful to begin this chapter with definitions of the terms “rehabilitation” and “restoration.” “Seagrass rehabilitation” is a general term with the sense of improving, augmenting or enhancing a degraded or affected area, with the expectation that there will be improvement through return of seagrasses and seagrass ecosystem function. The term “restoration” conveys the meaning of a return to pre-existing conditions. Since this is acknowledged as being an unlikely outcome in practice, “restoration” is usually interpreted as returning the ecosystem to a close approximation of its condition prior to disturbance (USNRC, 1992). By that definition, structure and function of the ecosystem are approximately created, but still with the expectation of producing a natural, functioning and self-regulating system integrated with the ecological landscape. The first recorded seagrass transplantation took place in Europe in 1939 (Reigersman et al., 1939). However, since the 1960s, serious restoration experiments Coastal Wetlands: An Integrated Ecosystem Approach
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and mitigation projects using different seagrass species have been attempted worldwide with varying degrees of success (Fonseca, 1992). The focus, however, has mainly been on the United States (Treat and Lewis III, 2006), Europe (this chapter) and Australia (Seddon, 2004). In this regard, it is worth noting that terrestrial restoration, even of emergent wetlands, has taken place over a much longer time frame – in the case of the former, for centuries. It is unsurprising therefore that seagrass restoration is still an evolving science that, globally, remains quite difficult and challenging (Gordon, 1996). In 1996, while there had been many failed projects, there had also been successes in putting back small areas of lost or damaged seagrasses, particularly with faster growing species such as Zostera marina (eelgrass), Halodule wrightii (shoal grass) and Syringodium filiforme (manatee grass). Gordon (1996), in his review of the status of international seagrass restoration, summarized the key issues related to planning, policy and management affecting seagrass restoration, planting methods, critical issues confronting successful restoration and research gaps identified to further develop the technology. He considered that the return of functional seagrass (or any other wild community) could not be guaranteed although success in attempting to restore, rehabilitate or create seagrass habitat, regardless of location, was more likely where the following factors were considered: selecting suitable sites, developing methodology appropriate to site conditions, improving seagrass spreading and coverage rates, accounting for selffacilitative properties, minimizing donor bed damage, overcoming high labor and time costs and preventing bioturbation. In the decade since Gordon’s (1996) review, several of these factors remain pressing issues (Lord et al., 1999; Greening, 2006). Considerable progress has been made, however, in addressing many of them. Site selection, for example, has improved markedly now that factors affecting seagrass growth are better understood. The application of exposure indices (Fonseca et al., 1998; Kelly et al., 2001; Fonseca et al., 2002) and technology such as sediment erosion sensors (Chisholm, in preparation) has also helped to rapidly characterize potentially successful sites. This and other research has been incorporated into models allowing assessment of site suitability for seagrass transplantation (Short et al., 2002; Biber et al., 2008) and has generated better logistical frameworks to guide restoration programs (Fonseca, 2006). Methodologies have improved in recent years. Although Gordon (1996) considered that overcoming high labor and time costs might be met by the development of mechanical techniques, site-specific manual methods have proven to be relatively efficient in planting appreciable areas (i.e., hectares) fairly efficiently (Davis et al., 2006; Montin and Dennis 2006; Orth et al., 2006b; Paling et al., in preparation). Various mechanical devices have also been developed for seagrass transplantation ranging from plug and sprig planters for shallow areas (Lewis et al., 2006; Orth et al., 2006b), boat-based hydraulic extraction of large (i.e., >1 m2) sods both in the United States (Lewis et al., 2006) and Japan (Nakase and Shimaya, 2001; Suzuki, 2002), and submerged machinery capable of operating to 15 m depth to move large sods in Australia (Paling et al., 2003). However, these mechanical devices typically have limited ranges of application with unknown cost effectiveness (Uhrin et al., in press).
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Other advancements have been the increasing evidence that restored seagrass beds may become self-sustaining in appropriate time frames (Fonseca et al., 1996; Paling et al., in preparation; Bos and van Katwijk, 2007; van der Heide et al., in press) and that many functional attributes return within a few years (Cambridge et al., 2002; Fonseca, 2006). Recovery of donor meadows has also been better demonstrated in a range of environments, providing affirmation as a viable source of transplant material (Fonseca et al., 1987, 1994; Paling et al., in preparation). However, salvage still remains widely used as a technique for mitigation (Lewis et al., 2006; Montin and Dennis, 2006). Despite these advances, problems remain. Bioturbation remains one of the most vexing problems in many areas (Philippart, 1994; Molenaar and Meinesz, 1995; Davis et al., 1998; Townsend and Fonseca, 1998; Siebert and Branch, 2007) although some transplant techniques (Fonseca et al., 1994, 1998; Short et al., 2006) do provide excellent protection from biotic disturbance. Improving spreading rates is also challenging. Nutrient additions to transplants provide mixed results in different locations and sediment types (Powell et al., 1989; Kenworthy and Fonseca, 1992) and research investigating hormone enhancement has been negligible. Ten years ago, techniques had been tested sufficiently to allow small areas of seagrass to be restored with reasonable assurance (Gordon, 1996). In the last decade, we have been able to transplant several hectares, and while returns of hundreds of hectares of seagrass through restoration efforts are still to be realized, tens of hectares have been attempted (Milano and Deis, 2006). This chapter updates the progress that has been made over the last decade and reviews the current status of seagrass restoration and transplant research in those areas of the world where much activity has taken place: Europe, Australasia and the United States. Within each region, a brief background is provided on seagrass geographic distribution, community species composition, causal nature of seagrass decline and an overview of attempts to facilitate recovery via transplanting. The chapter concludes with an evaluation of the ecological and economic appropriateness of restoration as a tool for conserving seagrass.
2. R EGIONAL ACTIVITIES 2.1. Europe In Europe, four seagrass species are widespread: Posidonia oceanica, Cymodocea nodosa, Z. marina and Zostera noltii. The Mediterranean Sea hosts all four species, with P. oceanica and C. nodosa most abundant. Since the 19th century, additionally, Halophila stipulacea has been spreading into the Mediterranean Sea from the Red Sea through the Suez Canal. The Atlantic, North, Baltic and Black Sea coasts are vegetated by Z. marina and Z. noltii (Bostro¨m et al., 2003; Hily et al., 2003; Lipkin et al., 2003; Milchakova and Phillips 2003; Procaccini et al., 2003; Ruggiero and Procaccini, 2004) and the Caspian Sea is vegetated by Z. noltii (Milchakova, 2003). Particularly in Europe, with its ancient history, the economic value of seagrasses has left scattered imprints in written heritage. For instance, the use of seagrass for
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dike enforcements in the Netherlands was referenced in documents of 1328 (Oudemans et al., 1870). The dynamics of seagrass beds were also noticed in the 18th century: “at one place it proliferates, at another it declines, and in still other places it disappears altogether; or, at places where it was never found before, it appears and increases year by year” (Martinet, 1782). People tried to predict abundant yields of seagrass: a rainy and warm spring was thought to enhance growth, hence the proverb (which can be considered as a primitive predictive model) “good hay grass, good seagrass” (Martinet, 1782). These quotes refer to Z. marina in the Wadden Sea, where hundreds of families depended on its harvest. When this species started to disappear at an unprecedented scale from the Wadden Sea at the beginning of the 1930s, related to wasting disease, a local fisherman transplanted seagrass from a remote remaining bed to his former site of harvest. This attempt was recorded by Reigersman et al. (1939) and is to our knowledge the oldest record of seagrass transplantation. It failed. Transplantations with a scientific goal were performed in the same area and era by Harmsen (1936), who found that subtidal and intertidal morphotypes did not survive when transplanted reciprocally but that control transplantations survived. During the 1930s, in northwestern Europe and the Black Sea, wasting disease decimated eelgrass (Z. marina) beds; they recovered only partially, and since the 1970s or 1980s new declines related to anthropogenic activities have occurred (Zaitsev et al., 2002; Bostro¨m et al., 2003; Hily et al., 2003; Milchakova and Phillips, 2003; Duarte et al., 2006). In these areas, seagrass beds are now estimated to cover only 20–35% of their pristine extent (Bostro¨m et al., 2003; Hily et al., 2003). In the Wadden Sea, subtidal beds did not recover at all (Giesen et al., 1990). In the western Mediterranean, losses are estimated to be between 30 and 40% (Procaccini et al., 2003). There is not much known about the eastern Mediterranean, but losses are probably less severe (Lipkin et al., 2003; Procaccini et al., 2003). Losses of deep seagrass beds are generally attributed to increased turbidity resulting from eutrophication and/or construction activities. Other seagrass beds are lost due to mechanical destruction (construction of harbors, beaches, land reclamation, sediment-disturbing fisheries activities, anchoring, etc.) or due to pollution (eutrophication, organic loads, thermal plumes, etc.; reviews in Duarte et al., 2006; Ralph et al., 2006). When circumstances improve, spontaneous seagrass recovery may be rapid, particularly when the losses were local and the species are capable of growing and dispersing rapidly. For example, in Portugal, recovery of Z. noltii was recorded after decreased nutrient loads and decreased fisheries activities. Rapid recolonization following cyclic geomorphological changes was also recorded in Portugal (Cunha et al., 2005). C. nodosa rapidly colonized areas that were buried by moving sand dunes (Marba´ and Duarte, 1995). Z. marina beds are highly dynamic (den Hartog, 1971) and local disappearances and recoveries sometimes occur rapidly, as was shown for the Baltic Sea (Frederiksen et al., 2004a,b), the Wadden Sea – both subtidal and intertidal (Martinet, 1782; Reigersman et al., 1939; van Katwijk et al., 2006; Reise and Kohlus, 2008) – the Atlantic (Gle´marec et al., 1997) and in the Mediterranean (Plus et al., 2003; Olsen et al., 2004). P. oceanica, however, is a slow-growing species, its recoveries are slow and the recovered areas remain vulnerable (Marba´ et al., 1996; Gonzalez-Correa et al.,
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2005; Gobert et al., 2006). Increased coherence of patchy Posidonia beds or other small-scale recovery may be observed after improved wastewater treatment (Jaubert et al., 1999; Pergent-Martini et al., 2002) or after deployment of artificial reefs protecting the beds against trawling (Gonzalez-Correa et al., 2005). Recoveries at the system scale are limited. For example, after large-scale seagrass decline in the 1930s along the northern Atlantic coasts, there was no recovery in the Wadden Sea (Giesen et al., 1990). Partial recovery was recorded in the Baltic Sea and along Atlantic coasts, but often with a time lag of decades (Ostenfeld, 1908; Duarte and Sand-Jensen, 1990; Gle´marec et al., 1997; Bostro¨m et al., 2003; Frederiksen et al., 2004a; Bernard et al., 2005; Gonzalez-Correa et al., 2005; Duarte et al., 2006), as is also found in other parts of the world (Rollon et al., 1998; Moore et al., 2000; Neckles et al., 2005; Walker et al., 2006). These findings suggest nonlinear feedback processes. The existence of multiple stable states in seagrass beds has been postulated (Duarte, 1995; 2002; Duarte et al., 2006; Valentine and Duffy, 2006) and even convincingly indicated by long-term data of Munkes (2005) of an estuary in the Baltic Sea, Greifswalder Bodden. She found that continuing high turbidity prevented recovery of Z. marina, Ruppia and a number of freshwater macrophytes despite marked reductions in nutrient inputs over the last 15 years. In contrast, Z. marina recovered in a similar area in Puck Lagoon, Poland (J. M. Weslawski, personal communication), where a sewage treatment plant, constructed in 1988 (Schiewer et al., 1999), reversed the decimation of the Z. marina meadows that had disappeared between 1957 and 1988 due to nutrient enrichment (Kruk-Dowgiallo, 1991). Similarly in Orbetello Bay, Italy, seagrasses (Ruppia cirrhosa, Z. noltii and single plants of C. nodosa) recovered from almost absence in 1990 to more than 1,250 ha cover (>50% of the bay area), concurrent with macroalgal harvesting, pumping of clean sea water into the lagoon and reduction of nutrient inputs (Lenzi et al., 2003). Many transplantation programs have been set up in Europe as a result of retarded seagrass recovery and also for mitigation purposes. In Limfjord, Denmark, a large-scale Z. marina transplantation program was carried out to restore natural values after nutrient loadings were reduced. Of the several techniques tested, only patches greater than 0.20 0.20 m were able to survive a winter season with heavy storms in shallow waters. In general, growth and survival rates were highest in these blocks, though 20% of the plugs also survived at the most favorable site. Surviving patches increased to 25 times their original size during the first 2 years. Spring transplants performed best in comparison to summer and autumn transplants. Seed dispersal efforts largely failed due to poor germination in the field. Poor seed germination was not typical; a batch collected 1 month previously showed high germination rates in the laboratory (Christensen and McGlathery, 1995; Balestri et al., 1998). Presently, only intertidal Z. marina and Z. noltii beds grow in the Wadden Sea, though during the 1970s these beds also disappeared in the western part. Intertidal seagrass reintroduction programs started in 1987 mainly to reintroduce intertidal Z. marina (Giesen et al., 1990; van Katwijk and Hermus, 2000; Bos and van Katwijk, 2007). Research on habitat requirements revealed that sheltered areas around mean sea level, with some freshwater influence, are preferred in the
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Wadden Sea (van Katwijk et al., 1999, 2000; van Katwijk and Hermus, 2000; van Katwijk and Wijgergangs, 2004). Suitability tests of donor populations indicated that nearby populations were preferred over remote ones (van Katwijk et al., 1998). Several anchoring techniques were tested and seed-bearing shoots were transplanted. The latter was successful for several years, but plants had largely disappeared by 2006. In short, Z. marina can maintain itself at very sheltered sites but is vulnerable to suffocation by macroalgae (Chapter 13). At slightly less sheltered sites, development is better, but seed germination or young seedling survival fails for unknown reasons (Bos and van Katwijk, 2005). Bare root transplantations of Z. noltii were more successful. Transplanted in 1993, patches are still spreading slowly (Figure 1a, van Katwijk et al., 2006). As in Limfjord (Denmark) in the Wadden Sea, spring transplantations generally survive better (Noten, 1983; van Katwijk and Hermus, 2000). In the 1970s, Z. noltii transplantation had also been performed in the United Kingdom to increase wildlife potential, particularly for Brent geese, anticipating mitigation related to the building of an airport. Turf or sod trials were reasonably successful, but larger scale transplantation failed (Ranwell et al., 1974; Boorman and Ranwell, 1977). In Mondego Bay, Portugal, experimental bare root transplantations of Z. noltii were carried out to test the most suitable season for transplanting. In contrast to more northern locations in Europe, late autumn or winter was preferred over spring or summer (Martins et al., 2005). In the Mediterranean, Z. noltii transplantation was carried out using sods during the 1970s and 1980s in Beaulieu, Martigues, and Toulon, France (Meinesz et al., 1990; Boudouresque et al., 2006), and in Venice Lagoon (Curiel et al., 1994; Rismondo et al., 1995). Most efforts used bare root cuttings attached to grids or looped pickets (Figure 1b), or turfs (“matte”) placed in dug holes. Results varied by type of cutting (Molenaar et al., 1993; Molenaar and Meinesz, 1995), season (Meinesz et al., 1992; Boudouresque et al., 2006) and depth (Genot et al., 1994). Higher planting densities yielded higher survival rates (Molenaar and Meinesz, 1995), which was also found for Z. marina in the Wadden Sea (Bos and van Katwijk, 2007). At the same locations except Martigues, C. nodosa was transplanted, (a)
(b)
Figure 1 (a) Three-year-old transplants of Zostera noltii in theWadden Sea, fringing the North Sea. These have continued to expand annually. (b) Transplantation of Posidonia oceanica in France.
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but their fate is unknown after 2 years (Meinesz et al., 1990 and references therein; Curiel et al., 2005). Additionally, in Venice lagoon, transplantations of Z. marina and C. nodosa have been carried out, comparing bare root and sod methods (Curiel et al., 2005). After two growing seasons, cover using the bare root method was nearly equal to that of sods and was more cost effective. Both C. nodosa and Z. marina transplantations were successful and expanded at Venice Lagoon. In France, three patents have been obtained for seagrass transplantation techniques (Meinesz et al., 1993). As seagrasses are protected there, it is assumed that no losses occur (however, see Boudouresque et al., 2006). Therefore, no more restoration programs have been carried out in France since the beginning of the 1990s (Meinesz personal communication). Presently, large-scale – but unmonitored – mitigations of P. oceanica take place in Spain, with quite poor results (Sa´nchezLizaso et al., 2006). Innovative approaches that account for environmental unpredictability and the dynamic population of donor material have been carried out in the Wadden Sea (van Katwijk et al., 2002). Recovery of seagrass in Europe has been possible mostly at a local scale, particularly for slow-growing species like P. oceanica. Large-scale declines may be irreversible at human timescales for two reasons. Firstly, seagrasses modify their environment and once they have disappeared, conditions may have become unsuitable such that only large-scale efforts can overcome problems with erosion and turbidity (Bouma et al., 2005; van der Heide et al., 2006). Secondly, eutrophication and organic loading may have altered both sediment and water column properties, a topic discussed in detail in Chapter 13.
2.2. Australia A number of authors have reviewed seagrass rehabilitation efforts in Australia over the past two decades (Kirkman, 1989, 1992, 1997; Gordon, 1996; Lord et al., 1999; Seddon, 2004). Additional reviews on seagrass communities as a habitat for fisheries have also been carried out (Cappo et al., 1998; Hopkins et al., 1998). These works have led to the development of guidelines for the protection and restoration of fisheries habitats in some places (Hopkins et al., 1998; Western Australian Environmental Protection Authority, 2004). Despite the acknowledged importance and significant losses of seagrass habitat in much of coastal Australia due to both natural- and human-induced causes, seagrass rehabilitation techniques were not seriously considered in Australia until the late 1980s. Major rehabilitation efforts have since been undertaken in Queensland, New South Wales, Victoria and Western Australia (Paling and van Keulen, 2002) and more recently in South Australia. Initial reviews suggested that rehabilitation of seagrasses in southern Australia was almost impossible due to the slow growth rates of the large temperate meadowforming species, reflecting the lack of attention given to pioneering species. Numerous experiments and development of new techniques have now changed that view to one of optimism, with the most recent published review proposing that it is now possible to rehabilitate small-scale seagrass loss (up to several hectares) (Seddon et al., 2004). Seagrass rehabilitation experiments to date have included
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transplantation trials using mature plants as well as seeds and seedlings, improvement or alteration of degraded habitat to encourage natural return of seagrasses and addition of plant growth enhancers including fertilizers and growth hormones. Transplantation experiments initially employed the range of techniques commonly used in the United States and Europe, but with little success. This was thought to be largely because most northern hemisphere seagrasses grow in relatively sheltered conditions in estuaries or coastal lagoons, whereas many southern Australian seagrasses grow in the open ocean, exposed to oceanic swell. Work in Western Australia, using plug planting units (PUs), showed that increasing PU size improved their survival, ostensibly as a result of increased stability against wave action (van Keulen et al., 2003). This realization led to the development of the ECOSUB mechanical technique capable of transplanting very large PUs, up to 0.5 m2 by 0.4 m deep (Figure 2a, Paling et al., 2001a,b). Within more sheltered waters, traditional manual transplantation techniques are still viable. Paling et al. (2007) concluded that seagrass rehabilitation was possible at several locations around Cockburn Sound, Western Australia, following manual transplantation trials using plugs and sprigs. Subsequently, a large-scale manual (a)
(b)
Figure 2 (a) The ECOSUB system is used in Western Australia to mechanically transplant large sods of seagrass. Diver operated to 15 m, the cutting machine is capable of removing 0.5-m deep sods that are 0.7 0.7 m in area. A line of Posidonia coriacea sods are shown here recently planted by ECOSUB2 which is able to plant in any configuration. (b) A viable handplanting method used in sheltered areas of Western Australia. The sprigs of Posidonia australis are attached to wire staples by biodegradable cable ties, spaced (as shown here) and then inserted into the sediment.
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transplanting operation has begun in southern Cockburn Sound using community volunteers to transplant sprigs of Posidonia australis (Paling et al., in preparation). This program is being conducted in conjunction with a study on the role of sediment stability in transplant survival (Figure 2b; van Keulen et al., 2005). Concurrent transplantation exercises underway in Albany, Western Australia, have met with very high levels of success (Bastyan, 2002). In a project targeted for the metropolitan area near Adelaide, sprigs woven into hessian sheets showed better growth characteristics and appeared healthier than plugs of Amphibolis antarctica and Heterozostera tasmanica, but plug survival percentage was greater (Seddon et al., 2004). Slow recovery times of many large meadow-forming species have led to the realization that seagrass harvesting and transplantation as a mitigation technique is not practicable. In response, observations on the role of naturally recruited seedlings in infilling seagrass meadows led to the use of seeds, seedlings and laboratorycultured material as alternatives for rehabilitation. Using various anchoring approaches for seeds and seedlings of Posidonia spp. and Amphibolis spp., Kirkman (1998) had consistently poor results. Overall, Posidonia spp. did not survive well and Amphibolis PUs did not spread after 17 months, apparently due to too much wave action and inappropriate anchoring. Development of seed stock has recently been studied. Seddon et al. (2005) cultured naturally collected seagrass seeds within holding tanks. They concluded that using a large-scale nursery approach to provide seedlings for rehabilitation was not yet practicable because of excess nutrients. Likewise, Wilson (2004) worked to develop in vitro propagation of seagrass seedlings, but ongoing contamination severely limited its application for large-scale rehabilitation work. Observations of seedlings becoming established around natural and transplanted seagrass patches suggest that propagation by seed is a viable means of patch consolidation (Paling et al., 1998). However, subsequent studies by Langridge (2002) confirmed the poor survival of naturally established seedlings in exposed locations. High wave energy in southern Australian makes survival of artificially established seedlings doubtful, with little improvement when sediment-protection meshes are used. Hessian mats were successful by contrast as a means of trapping and protecting seedlings in South Australia. A focus is therefore required on methods of protecting seedlings from erosion and disturbance until they can become firmly established.
2.3. Oceania Australia and the Indo/west Pacific have a diverse tropical seagrass flora characterized in many places by mixed-species meadows that include species of Cymodocea, Thalassia, Enhalus, Syringodium and Halodule (Larkum and den Hartog, 1989). Australia, China and Japan by virtue of their broad latitudinal ranges additionally have subtropical/temperate seagrass floras, dominated by monospecific meadows of Posidonia and Amphibolis spp. in Australia and Zostera spp. in China and Japan (Green and Short, 2003). The important roles seagrasses perform in coastal habitats are slowly being recognized and acknowledged in Oceania. The role of seagrasses as
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nursery areas for juvenile fish has been assumed for some decades; the physical influence of seagrasses on their substrate has been realized only recently and is now receiving careful scrutiny following significant coastal erosion problems resulting from seagrass loss. In the Oceania region, there is a growing concern about coastal habitat loss including seagrass habitats. Many causes are cited as responsible for these losses, both natural and human, but until the reasons for seagrass decline are understood and dealt with, restoration remains difficult. While seagrass rehabilitation efforts are well progressed in Australia and to some extent in Japan, most of the countries in the region are not in a position to seriously tackle issues of large-scale habitat restoration; indeed, the concept of marine conservation is not well advanced in many developing countries in the region. There is a clear push from intergovernment agencies to advance marine conservation with a view to encouraging local ownership of these environments and provide training in grass roots environmental management. Numerous reports have identified the problem of habitat loss and degradation in southeast Asian and Pacific countries, particularly in areas of high population density and human impact, including deforestation, mining and agriculture. While there is considerable concern about these losses, particularly at a local level, efforts to protect and rehabilitate seagrass habitats are hampered by a number of factors. The first of these is a poor knowledge of the distribution and condition of seagrass habitats; this is largely due to a lack of accurate mapping and monitoring, exacerbated by high turbidity in many tropical coastal areas, hampering the effectiveness of aerial photography. The detailed aerial mapping exercises carried out in parts of Australia are not possible in many parts of southeast Asia because of lack of water clarity and economics. The second major factor is the poverty experienced by much of the coastal populations of southeast Asia and the Pacific Islands. Their reliance on subsistence fishing and the urgency of obtaining food leaves little energy or resources for environmental stewardship.
2.4. Southeast Asia Throughout southeast Asia, there has been a growing awareness of the importance of seagrass ecosystems to local fisheries. The Philippines government in particular has been active in promoting the values of seagrass ecosystems. In the early 1980s, the government requested assistance from the Food and Agriculture Organization of the United Nations to investigate the feasibility of seagrass rehabilitation in the Philippines (Thorhaug, 1987). An international consultant was appointed to provide a technology transfer program to introduce seagrass rehabilitation techniques and to undertake an experimental transplantation program. In addition, the program examined transplantation by plugs and sprigs of five common species, as well as of Enhalus seeds. Transplantation was carried out in areas subject to a range of human impacts, including dredging and pollution by sewage, mine tailings and thermal effluent. All species showed moderate success in the trials, with best performance by Enhalus, although the colonizing species Halodule, Cymodocea and Syringodium showed the most lateral spreading. Of the techniques used, sprigs were
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most successful; dredged and filled areas showed the best results overall for transplant survival (Thorhaug, 1987). While the trials of transplanting conducted by Thorhaug appear to have been largely successful, Fortes (1990) acknowledged that perhaps the poverty of these communities precludes the luxury of being able to implement western-inspired conservation measures. To help combat these problems, a recent UNEP/GEF program has established a number of habitat demonstration sites across the region, aimed at demonstrating sustainable environmental management practices and reversing habitat degradation (UNEP, 2004). The habitat sites selected included four for seagrasses located at Bolinao, Philippines; Hepu, China; East Bintan, Indonesia; and Kampot, Cambodia; these have been progressively implemented since 2005 (UNEP, 2006). Project targets include control over seagrass-destructive practices and recovery of damaged areas within demonstration sites, as well as community education and training to encourage community ownership and custodianship of the sites across the southeast and eastern Asian region (Fortes et al., 2006). Artificial seagrass units were deployed in Singapore to provide an artificial habitat in degraded areas (Talbot and Wilkinson, 2001). This was found to be effective in providing a sheltered habitat for benthic organisms, thus enhancing recovery of degraded systems. Similar deployment of artificial seagrasses has been undertaken in the Philippines with a view to improving fisheries production (Fortes, 1988). Talbot and Wilkinson (2001) suggested that seagrass recovery best responds to removing the cause of disturbance but noted that this is difficult where limited resources preclude such broad level changes to environmental management. In the interim, transplanting and artificial seagrass projects were seen to provide a short-term solution. Most seagrass on the Korean coast, primarily Z. marina, has been lost due to anthropogenic disturbances since the 1970s; however, there has been no large-scale rehabilitation program in place to date. Recent research has been conducted using staples, Transplanting Eelgrass Remotely with Frame Systems (TERFS) and oyster shells at different times of year and sediment types (Park and Lee, 2007); the authors noted that staple transplants were most successful but particularly costly when compared to the other two methods that did not require diving (Short et al., 2002). High temperatures were observed to have a strong deleterious effect on transplants established in summer.
2.5. China and Japan During the recent development of the new Hong Kong International Airport, attempts were made to transplant Zostera japonica that was at risk from sedimentation during the land reclamation process at San Tau Beach. While transplanting the Z. japonica at San Tau was successful, attempts to transplant seagrass to other locations were not. It was suggested that there was a lack of appropriate locations for transplanting seagrass to, as most suitable habitat was subject to land reclamation or other coastal development (Fong, 1999). As part of the UNEP/GEF sustainable marine habitat management demonstration program, a site was established at Hepu
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in southern China. The intention again is to improve community and local government awareness of the importance of seagrass ecosystems with a view toward involving the community in seagrass management. Numerous concerns have been aired about the loss of seagrass meadows from the Japanese coast, particularly in response to large-scale coastal developments. While little has been published, there are a number of workshop proceedings; detailed methods and results are not readily available, however. Attempts have been made to enhance germination and growth of seagrass seedlings using a “sowing sheet.” This consists of a cloth sediment protection that prevents loss of seeds and seedlings during the initial growth stages. Results appeared promising after 2 years, although it was concluded that coarser cloth might be better for root penetration (Yoshida et al., 2001). More recently, a seed harvesting mat was developed to trap naturally released Zostera seeds and provide a stable substrate for germination and establishment before transplantation to the restoration site (Taisei Corporation, 2006). Results from controlled mesocosm studies that replicated tidal flows and wave exposure indicated that seagrass restoration would be feasible in Tokyo Bay (Nakamura, 2005). Experimental seagrass transplantation was conducted in Japan as part of mitigation studies for coastal landfill activities in the Awase Tidal Flat, Okinawa. This project transplanted large (1.5 m2) blocks of seagrass and sediment using a modified backhoe. The process is modified from terrestrial turf transplantation techniques. Transplanted blocks did not survive the high wave energy; in fact sand from the disintegrating blocks smothered some of the surrounding patchy seagrass. Subsequently, manual transplantation experiments were deemed appropriate and the landfill project was allowed to proceed (Suzuki, 2002). Penta-Ocean Corporation developed a seagrass transplantation machine built on an excavator platform for intertidal applications. The technique appears similar to the Western Australian ECOSUB system and consists of a bucket harvester that scoops up an intact sod of seagrass; the interior of the bucket can be removed from the machine and transported separately to the transplant site (Penta-Ocean web site, http://www.penta-ocean.co.jp/english/r_d/envi/eelgrass.html). Encouraging results were obtained during trials in Hiroshima Province (Nakase and Shimaya, 2001); however, these authors also cautioned about the need for appropriate wave energy conditions to ensure long-term survival and growth.
2.6. New Zealand and the Pacific Islands While there has been recent concern about damage to seagrass ecosystems in New Zealand’s developed coastal areas, few rehabilitation studies have been conducted. Transplantation trials using sprigs, plugs and 1 m2 sods of Zostera novazelandica in Manukau Harbour in 1993 showed good survival until the onset of winter storms after 6 months. The cause of decline was believed to be increased sedimentation and erosion of the PUs (Turner, 1995). Extensive loss of seagrass from Whangarei Harbour since the 1960s was believed to be due to increased sediment load and turbidity resulting largely from dredging and sediment discharge (Reed et al., 2004). Subsequent evaluation of environmental conditions within the harbor
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determined that conditions were suitable for a proposed transplantation program (Reed et al., 2005; A.M. Schwarz, personal communication). A comprehensive management document for New Zealand seagrasses was released that outlines key concerns and priorities for conservation (Turner and Schwarz, 2006). Coles (1996) visited several Pacific island nations as well as Malaysia and reported an overall support for environmental restoration work, including the rehabilitation and transplantation of seagrasses. It was felt, however, that these were not more than cosmetic attempts and not likely to be particularly effective.
2.7. United States There are approximately 12 species of seagrass in North America and they form substantial habitats in all coastal states except Georgia and South Carolina. Eelgrass (Z. marina) occurs in the temperate zone, along with Ruppia maritima, which appears to be ubiquitous. On the west coast, Phyllospadix spp. (surfgrasses) are found in the rocky intertidal zone, as well as an invasive Z. japonica. In the southeast and Gulf of Mexico, Thalassia testudinum forms extensive, slow-recovering beds. In addition, S. filiforme, H. wrightii and four Halophila species are present. Seagrasses were used by Native Americans, and early 20th century Europeans used Z. marina for insulation. More recently, they have only been valued for their role in supporting waterfowl and fisheries, especially the bay scallop (Aequipecten irradians concentricus). Although recognized for their functional role in coastal ecology as early as the 1920s on the east coast, quantitative studies did not begin in earnest until the 1970s. Since then, published work on seagrasses has grown exponentially. The correlation between human activities near the shoreline and seagrass decline was clear a decade ago (Short and Burdick, 1996). Large-scale losses had been documented in Chesapeake Bay (Orth and Moore, 1981) and in the Gulf of Mexico (Livingston, 1987). Seagrass beds in Tampa Bay were reduced by over 50% (Haddad, 1989) and 35% of the seagrass acreage in Sarasota Bay was lost, as well as 29% of that in Charlotte Harbour, Florida and 76% of that in Mississippi Sound (Eleuterius, 1987). Pulich and White (1991) reported a loss of 90% in Galveston Bay, Texas. Thom and Hallum (1991) reported similar ranges of losses from Puget Sound and large declines have occurred in the San Francisco and San Diego Bays (Kitting and Wyllie-Echeverria, 1992). The loss of seagrasses due to dredge-and-fill activities has been significant although direct impacts from mooring scars and propeller scars (Sargent et al., 1995) also emerged as a major source of habitat loss, along with scallop harvesting and the practices of raking and prop-dredging for other shellfish and crabs. In contrast to physical damage, most of the documented seagrass loss in the United States, as elsewhere in the world, has been due to human-induced reductions in transparency associated with degradation of water quality (Kenworthy and Haunert 1991). Thermal effluents from electric power plants have also caused extensive declines (Zieman and Wood, 1975; Fonseca et al., 1998). Losses due to decreased light availability tend to be irreversible except in rare cases where nutrient discharges are reduced to improve water quality (Johansson and Lewis, 1992).
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Tremendous losses of this habitat have therefore occurred as a result of development within the coastal zone. While disturbances usually kill seagrasses rapidly, their recovery is often comparatively slow (Fonseca et al., 1987). In recognition of this, seagrass ecosystems are now protected in the United States under the federal “no-net-loss” policy for wetlands in recognition of their role in performing important ecological functions. Prior to the decision to transplant, site selection should be carefully evaluated. In many cases where the injury was mechanically derived, restoration can often be achieved, but may also require engineering interventions to “fix” the site, such as filling excavation holes caused by vessel groundings (Figure 3a) or altering water flow. However, when the injured site has been irrevocably altered and offsite selections are made for restoration, more significant problems arise, usually from improper identification of potential sites. The most common mistake in site selection is the assumption that open areas or gaps among existing seagrass are prime sites for restoration when, in reality, the sites either cannot support seagrass or currently support very low densities. This problem has been addressed in detail in several publications (Fonseca et al., 1987, 1998; Fonseca, 1992, 1994). Suffice it to say, Fredette et al.’s (1985) condition “If seagrass does not grow there now, what makes you think it can be established?” best sums the problem and defines the question that must be answered in order to begin a logical process of site selection. Criteria for offsite selection include (1) locations that have similar depths as nearby natural seagrass beds and are similarly
(a)
(b)
Figure 3 (a) Sediment-filled tubes constructed of biodegradable fabric are placed into propeller scars in Florida to prevent additional erosion and to enhance seagrass recovery. Filling the scar with tubes brings the sediment to the surrounding level, which enables adjacent seagrass rhizomes to grow over the scar. (b) Transplanting Halodule wrightii near Beaufort, North Carolina, using the peat pot method. This photograph shows the use of the plugger dropping a small plug into a peat pot in which pelletized fertilizer has been added.
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anthropogenically disturbed; (2) areas not subject to chronic storm disruption; (3) sites not undergoing rapid and extensive natural recolonization by seagrasses; (4) similarity to sites where seagrass restoration had been successful; (5) areas of sufficient acreage to conduct the project; and (6) an area that could be restored to a similar quality of that lost (Fonseca et al., 1998; Calumpong and Fonseca, 2001). By giving attention to these criteria, the probability of successful restoration can be greatly enhanced. Recently, however, significant advances have been made using modeling approaches in site selection (Short et al., 2002; Biber et al., 2008). Since the publication of Fonseca et al. (1998), there have been few observable advances in technological innovation in regard to seagrass transplantation (Figure 3b), with two notable exceptions. Mechanical planting, once only represented by large benthic machines in Western Australia, has now been developed for shallow-water planting in the United States. Versions of the machine that transplant large sods have had successful field trials (Uhrin et al., in press) in contrast to designs for individual planting, which are ineffective relative to manual methods (M. Hall, personal communication, Florida Fish and Wildlife Commission). Another innovation has been the improvement of seeding techniques (at least for Z. marina) through improved holding processes (R. Orth, personal communication, Virginia Institute of Marine Science) and deployment techniques such as reusable seed buoys (Pickerell et al., 2005). The use of appropriate seeding techniques is undoubtedly the most promising method for large areas. Nonetheless, even moderate attention to logistics and manual techniques can have large payoffs. Recent restoration work in Tampa Bay, Florida, was most successful by simply using shovelfuls of seagrass hand-dug, transported with care and installed.
3. POLICY ISSUES R ELEVANT TO M ITIGATION The US federal “no-net-loss” policy, based initially on wetlands, places emphasis on replacing the area and functions of areas that are converted to other uses. “Compensatory mitigation” is a term used for destruction of existing habitat when the agent of loss and the responsible party are known. Compensation assumes that ecosystems can be successfully replaced. While planting seagrass is not technically complex, there is no simple way to meet the goal of maintaining or increasing acreage. Rather, the entire process of planning, site selection, planting and monitoring requires attention to detail if successful outcomes are to be expected. [In contrast, “restoration” is the term that is often used when there is no responsible party. In these cases, community-based approaches place emphasis on planting, with little attention to monitoring the outcome. Because the causative agent is ambiguous in such cases, and may be related to water quality, it is unlikely that planting efforts will be successful (see site selection above).] To prevent continued loss of habitat under compensatory mitigation, decisive action must be taken by placing emphasis on improving site selection, compliance, generating desired acreage and maintaining a true baseline. Many recent projects that involve logical site selection and new planting methods show great promise for reversing the trend of project failure.
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A logical and ecologically defensible goal of mitigation is to replace the lost seagrass species with a surface area that compensates for interim lost resource services and shoot density. Effective examples of plantings quickly provide many of the functional attributes of natural beds. However, when destruction of an impacted site requires planting in another location (i.e., offsite), it is often difficult to find a location with suitable biological and physical conditions. Given these goals and assumptions, the status of seagrass restoration can be viewed from within the US legal framework. Seagrass beds in coastal waters are generally viewed as public trust resources. Injuries to these resources are considered losses suffered by the public, and violations have been successfully prosecuted in US Federal Court. To evaluate this loss in a fair and reasonable manner, however, considerations include not only the actual loss in acreage in an injury – but also the loss of resource services provided by the seagrass bed between the time it is injured and the time it recovers to 100% of preinjury conditions (Fonseca et al., 2000). It is noteworthy that most effective restorations are in response to mechanical disturbance rather than degraded water quality that may only be resolved through more stringent watershed management. Beginning with the assumption that seagrass ecosystems produce irreplaceable goods and services enjoyed by society, how can the value of these attributes be quantified? This issue emerges in two contexts: (1) how much does it cost to restore (and subsequently monitor) seagrass ecosystems and (2) what is the value of services provided by seagrass ecosystems on an annualized basis so that losses can be compared with gains? The former question can be answered using existing reports. The latter is addressed by describing the habitat equivalency analysis (HEA).
3.1. Costs of restoration Seagrass restoration is expensive. Much early experimental research was driven either by entire ecosystem degradation or by loss from nonpoint source factors (e.g., eutrophication) and the desire to return the seagrass, along with the services they provided, to “damaged” ecosystems (i.e., true “rehabilitation”) or the requirement for offsite mitigation for “small-scale” development. These two activities (i.e., large vs. small-scale) are quite different and need separate consideration. In terms of the small-scale setting, what metric should be used to determine the economic viability of seagrass restoration? Sufficient seagrass restoration projects, along with transplant research, have now been attempted to allow their appropriate costing (Fonseca, 2006) and comparison with terrestrial activities. In Europe, recent mitigation activities in the southwest Netherlands have cost approximately $50K/ha (2007 US dollars), although restoration research in the Wadden Sea over the last two decades has amounted to $4,000K. In Australia, mechanical seagrass transplantation (including design and development, construction, testing and associated site selection) can be costed at $1,000K/ha. Recent exercises using volunteer manual planting are far cheaper at $16–$34K/ha, depending on plant unit spacing. The same planting using professionals range from $84 to $168K/ha. In the United States, the most expensive programs have cost between $1,900 (McNeese et al., 2006) and $3,387K/ha (Lewis et al., 2006) depending upon the site and project involved. The lower end of the range is between
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$33 and $99K/ha (Bergstrom, 2006). Even these rough comparisons may or may not include monitoring costs. Underwater restoration is on average far more expensive than that on land and will likely continue to be so in the future (Short et al., 2006). In the United States, an agreed upon “all-inclusive” price, including monitoring, was presented in Treat and Lewis (2006) as between $570 and $972K/ha (Fonseca, 2006). By comparison, terrestrial restoration, while more advanced in practice, is commonly conducted at much larger scales. Costs range between $18 and $353K/ha for land rehabilitation of mining effects and for special purpose restoration (i.e., difficult or “iconic” sites; R. Hobbs personal communication) to much cheaper, large-scale bushland restoration ($2–$8K/ha). So is restoration a viable tool for conserving seagrass in both ecological and economic terms? It would appear that there is sufficient information to indicate that it is ecologically defensible in regard to donor meadow recovery, return of function and self-facilitative properties. From an economic viewpoint, however, it is clearly far more cost effective (e.g., $1.4K/ ha; Stowers et al., 2006) to preserve a seagrass habitat from damage than to restore an area after its degradation. This is especially true at large scales because it will simply be too expensive to put back areas greater than, for example, 20 ha. Thus, there is little hope in using transplant restoration to offset receding seagrass beds worldwide where declining water quality is a major cause. At small scales, however, we possess the ability to successfully generate small seagrass meadows (at high cost), and these are often directed toward offsetting small-scale damage.
3.2. Valuation of ecosystem services The valuation of ecosystem services is now a global exercise (Millennium Assessment, 2005). Within the context of a particular ecosystem or restoration program for replacing ecosystem services, the specifics of a particular habitat need to be acknowledged (Costanza et al., 1997). For seagrasses in the United States, the HEA has been employed as a way to compensate for habitat lost, but also for the loss of services over time. In its most basic application, HEA determines the appropriate scale of a compensatory restoration project by adjusting the project scale such that the present value of the compensatory project is equal to the present value of interim losses due to the injury. Because these services are occurring at different points in time, they must be translated into comparable present value terms through the use of a discount rate, a standard economic procedure that adjusts for the public’s preferences for having resources available in the present period relative to a specified time in the future. For argument’s sake, if 1 ha of seagrass were destroyed today and replanted tomorrow and reached standards of equivalency (e.g., shoot density, biomass and coverage) in 2 years, the interim loss of ecological services over this 2-year period would be relatively low. However, if restoration of this site were not undertaken immediately and it required 7 years to reach its preimpact state, the level of compensation due the public for the interim losses from this same injury would be substantially greater. This highlights the weakness of fixed compensation ratios, for example, replacing 2 ha of restored seagrass for every 1 ha irreversibly altered by a particular project (dredging, filling, etc.). HEA is one of the more frequently used methodologies available to natural resource trustees (Fonseca et al., 2000). The basic approach underlying HEA is to determine the amount
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of compensatory habitat to be restored, enhanced and/or created such that the total services provided by the compensatory project over its functional lifespan are equal to the total services lost due to the resource injury. While HEA is conceptually and computationally straightforward, its proper application requires a detailed understanding of the biological and ecological processes that affect the recovery and productivity of injured and restored habitats. Recently, NOAA developed and applied HEA using basic biological data to quantify interim lost resource services (National Oceanic and Atmospheric Administration, Damage Assessment and Restoration Program, 1997). While sharing many of the same principles as other methods by incorporating interim losses into replacement ratio calculations for wetlands (King et al., 1993; Unsworth and Bishop, 1994), HEA focuses on the selection of a specific resource-based metric(s) as a proxy for the affected services (e.g., seagrass short-shoot density) rather than basing its calculations on a broad aggregation of injured resources. Its use does not reduce the need for appropriate performance standards to ensure that a project provides the anticipated level of services. Well-defined and measurable standards are essential for the success of the project regardless of whether the restoration will be implemented by the parties responsible for the original resource injury or by the trustees using monetary damages that are recovered.
4. C ONCLUSIONS Having examined and updated seagrass restoration in a worldwide framework, it is useful to reflect on generalities that have arisen independent of geographic location and which allow us to evaluate the ecological and economic appropriateness of this activity as a tool for conserving seagrass. The last decade has seen the development of a suite of innovative techniques (Treat and Lewis, 2006) with broad applicability to a range of environments. It remains inevitable that population pressures and economic necessity will generate coastal development and destructive activities in areas currently occupied by seagrass. While minor damage, such as boat groundings, may be reduced through relatively inexpensive means such as exclusion enforcement, legislation and education, it is likely that restoration will still be required as offsite mitigation for port and harbor development. We now possess the ability to successfully generate small seagrass meadows, at high cost, to support this mitigation. However returns of areas greater than 20 ha via restoration efforts are unlikely to be realized in the near future.
ACKNOWLEDGMENTS We thank E. Balestri, A. Bos, P. Christensen, K. Hermus (Figure 1a), I. Martins, A. Meinesz (Figure 1b), A. Rismondo, J.L. Sanchez-Lisazo and J.M. Weslawski for providing data or photographs. Thanks are also due to Professor Richard Hobbs for his editorial advice and land rehabilitation costings and to Mark M. Brinson and an anonymous reviewer for their suggestions.
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T IDAL M ARSH C REATION Stephen W. Broome and Christopher B. Craft
Contents 1. Introduction 2. Principles and Techniques of Tidal Marsh Creation 2.1. Site selection 2.2. Conceptual design 2.3. Hydrology 2.4. Soil 2.5. Establishing vegetation 3. Evaluating Functional Equivalence of Created Tidal Marshes 3.1. Biological productivity and food webs 3.2. Biogeochemical cycles 4. Summary References
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1. INTRODUCTION Tidal salt and brackish water marshes are ecotones between land and sea that occur in the upper intertidal zone of sheltered coastal areas such as estuaries and bays, and behind barrier islands. Their hydrology is characterized by regular or irregular inundation by tidewater and subsequent drainage. Tidal effects produce distinct zonation of the herbaceous vegetation, which is related to frequency, duration, and depth of inundation as well as salinity. Also, the ebb and flow of tides connects the marsh to the adjacent water body, and has been described as an energy subsidy that increases primary productivity (Mendelssohn and Kuhn, 2003) and facilitates the exchange of organic carbon, mineral nutrients, sediments, aquatic organisms, and other material. The emergent vegetation consists of a limited number of salt-tolerant species, most commonly grasses, sedges, or rushes. Plant diversity decreases as salinity increases. Tidal marshes are productive ecosystems that serve as nursery grounds for aquatic organisms (Minello et al., 2003, 2008) and habitat for wildlife. In many coastal environments, much of the primary production that occurs in tidal marshes is exported to adjacent waters in the form of detritus, which becomes a part of the Coastal Wetlands: An Integrated Ecosystem Approach
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estuarine food web providing energy for bacteria, fungi, worms, oysters, crabs, shrimp, fish, etc. Other important functions of tidal marshes include buffering storm surges, storing floodwater, protecting shorelines from erosion, stabilizing dredged material, dampening the effects of waves, trapping waterborne sediments, nutrient cycling and transformations, and serving as reservoirs of nutrients and organic carbon (Mitsch and Gosselink, 2000). The area occupied by coastal marshes has been reduced over time, and existing marshes have been degraded due to human activities and natural processes. Anthropogenic sources of marsh loss include dredging, filling, dike and levee construction, drainage, urban and agricultural development, oil and gas exploration, and construction of port facilities, highways, bridges, and airports. Natural forces such as sea-level rise, land subsidence, and erosion also result in losses of tidal marshes. Marsh functions and values may be lost due to pollutants such as oil or chemical spills. The vulnerability of tidal marshes to anthropogenic and natural destruction and degradation has led to an interest in creation of new marshes to replace their lost structure, function, and value. In many cases, tidal marsh creation is required by regulatory agencies to mitigate losses resulting from development activities.
2. PRINCIPLES AND TECHNIQUES OF T IDAL M ARSH C REATION Wetland creation is defined as the establishment of wetlands through humaninduced changes in the landscape on sites where no wetland existed in the recent past (Lewis, 1990; Streever, 1999; Mitsch and Gosselink, 2000). Wetland creation differs from restoration in that restoration implies reestablishing hydrology and functions of a former wetland. We define tidal marsh creation in this paper as human-induced conversion of upland or subtidal habitat to tidal salt or brackish water marsh ecosystems characterized by emergent vegetation. Tidal fresh water marshes (<0.5) will not be addressed. Tidal marsh creation is accomplished by manipulating topography, hydrology, and soils, often followed by planting vegetation to create conditions that will, through succession, lead to self-sustaining ecosystems similar in structure and function to natural tidal marshes in the area. Typical objectives of tidal marsh creation include dredged material stabilization (Landin et al., 1989; Streever, 2000), mitigation (make impacts less severe) required by government regulations (Darnell and Smith, 2001; Hough and Robertson, 2008), shoreline erosion control (Maryland Department of the Environment Wetlands and Waterways Program, 2006), accumulation of sediment for sea defense (Moller et al., 2001; Hofstede, 2003), and to create agricultural land (Chung, 2006). In Europe, tidal marshes are created using dredged material (Bernhardt and Handke, 1992), accretion enhancement techniques (Hofstede, 2003), and by realignment of coastal defenses such as clay banks that are used to protect the coast from storm tides and rising sea level (Miren et al., 2001; Crooks et al., 2002; Wolters et al., 2005; Garbutt et al., 2006, Chapter 27). These tidal marsh creation projects are passive in that, rather than planting, vegetation is allowed to naturally
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colonize the site. In the United Kingdom and elsewhere in Europe (Netherlands, Belgium, Germany), managed and unmanaged realignment of coastal defenses has led to reclamation of large areas of tidal marshes, some of which were reclaimed from the sea for agriculture hundreds of years ago (Wolters et al., 2005). Many agricultural sites were accidentally breached during storm surges. Recently, embankments were deliberately breached to re-create salt marshes for conservation and as a first line of defense against the sea (Wolters et al., 2005). There are many examples of successful tidal marsh creation, but methods and techniques vary depending on local conditions. Tidal marsh creation requires careful planning that takes into account site selection, hydrology, vegetation, soil, and construction costs. It should be recognized that there are likely to be unknowns at any particular site, requiring an adaptive approach to define goals and identify potential techniques to achieve those goals (Zedler, 2001).
2.1. Site selection An important consideration in selecting a site is the relative value or importance of the ecosystem that will be replaced. For example, is converting an upland forest or grassland to a tidal marsh a positive gain for the environment? The same question arises if aquatic habitat is destroyed by fill material or sediment deposited at a river mouth or inlet that raises the elevation to create a tidal marsh. Answers to such questions may be determined to some extent by scientific data, or by value judgments usually determined by environmental managers, regulators, or political processes. On the other hand, converting a dredged material disposal site, marginal farmland, an eroding shoreline or flood-prone urban land to tidal marsh habitat would be seen by most as a positive action. Locating tidal marsh creation sites in areas that are protected from erosion is an important consideration in site selection. Waves and currents as well as severe storms are threats to establishment and long-term stability of coastal marshes.
2.2. Conceptual design Natural reference marshes, preferably located in the same general area and landscape setting as the proposed tidal marsh creation, should be identified to serve as models or target ecosystems for each creation project. Observations and studies of reference marshes can be used to determine the dominant plant species, the upper and lower elevation limits of each plant community, the slope and surface topography, and the depth and width of creeks if they are present. Determining the precise elevation requirements of plant communities and creating those elevations along with proper slope and drainage in the newly created marsh is critical to success. Marsh creation is more difficult in areas that have narrow tide ranges because the elevation difference between upper and lower elevation limits of plant communities is small and leaves little room for error in preparing the surface to be planted (Broome, 1990). Creeks or drainage channels should be installed in large creation projects to maximize tidal exchange and utilization by fish and other marine organisms (Zeff, 1999).
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2.3. Hydrology Hydrology is the dominant factor that determines development of the biological and physical characteristics of a wetland (Mitsch and Gosselink, 2000). Tidal flooding is the most obvious hydrologic factor that affects zonation of plant species, plant growth, soil chemical and physical properties, and biological processes in tidal marshes. Hydrology is largely determined by elevation, slope, and tidal regime, which interact to determine the area of the intertidal zone and the depth and duration of flooding that occurs. Subsurface hydrology may also affect these same physical and biological processes. For example, where marshes occur adjacent to uplands, groundwater seepage and runoff may be important hydrologic factors, adding fresh water and nutrients to the landward edge of the marsh (Harvey and Odum, 1990; Nuttle and Harvey, 1995). Rainfall, river flow, and wind effects also influence the hydrology of tidal marshes. The fundamental requirement of tidal marsh creation is establishing a surface for plant growth within the intertidal zone at elevations that support the flora and fauna of the target communities. Natural development of tidal marsh occurs when sediments accumulate to an elevation that can be colonized by pioneer marsh plant species by seed, rhizomes, marsh plants, or whole marsh sods that may wash up on the site (Boorman et al., 2002). Marsh creation by placement of dredged material is similar to this natural process, while marsh creation from upland requires excavation and grading to create a surface within the intertidal zone (Figure1a,b). The surface drainage or creek systems that occur in reference tidal marshes should be duplicated in created marshes with similar spatial distribution, width, and depth to facilitate tidal exchange (Zeff, 1999; Hood, 2007).
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Figure 1 (a) A tidal marsh was created on land previously used for crop production by grading to intertidal elevations, replacing the stockpiled topsoil, and excavating a drainage system that was connected to the estuary after planting. Greenhouse-grown seedlings of J. roemerianus, Spartina patens, and S. alterniflora were planted (60 cm spacing) in June 2006. (photo June 2006). (b) Vegetative cover was nearly complete by July 2007.
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2.4. Soil Soil physical and chemical properties affect construction, plant growth, and functional development of habitat in created marshes (Callaway, 2001). Construction operations such as grading, installing drainage channels, and planting vegetation are generally less difficult on sandy soils than fine textured clayey and silty soils, or organic material. Sand has a greater bearing capacity than finer sediments, which makes operation of equipment, transplanting vegetation, and even walking on created sites much easier. A disadvantage of sandy material is its low content of plant available nutrients, which may result in poor initial growth of vegetation. Over time, tidal flooding will deposit silt, clay, and organic sediments that add nutrients to the developing created marsh. Nitrogen (N) and/or phosphorus (P) are the nutrients that often limit plant growth in created marshes. The other essential plant nutrients are usually supplied in adequate amounts either from the soil or saltwater flooding. Adding N and P fertilizers to enhance plant growth and initiate nutrient cycling is beneficial and may be critical to success on sites that are nutrient deficient (Broome, 1990). Maintenance fertilization may also be necessary until nutrient pools are enhanced and nutrient cycling is self-sustaining. At some locations, natural sources of nutrients from tidal inputs, seepage, runoff from uplands, N fixation, and precipitation, may be more than adequate to maximize plant growth. Pollutants such as sewage effluent, and nonpoint sources from agricultural land and urban areas may also result in N and P enrichment. When upland soils are graded to intertidal elevations to create marshes, the new surface that is exposed may have physical and chemical properties that are unfavorable for plant growth and marsh development (Figure 2). Appropriate soil amendments such as organic matter, lime, and fertilizer may be needed to improve plant
Figure 2 Soil pH values below 3.0 developed in small areas in a tidal wetland created by grading an upland soil in eastern North Carolina to intertidal elevations. The low pH was the result of oxidation of acid sulfate soil material when it was exposed to the surface. Spartina cynosuroides seedlings did not survive in the low pH soils (foreground) until the affected areas were limed and replanted.
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growth. This problem can be avoided by stockpiling and then replacing the topsoil (A horizon) after grading. Care should be taken to insure that the correct final surface elevation to support plant growth is maintained after the topsoil is applied. At most locations, topsoil is likely to have more favorable physical and chemical properties than subsurface layers, and would be expected to produce better plant growth and development of marsh functions. Standard soil testing, as used for agricultural crops and lawns, can be utilized to determine if available plant nutrients are adequate, and if the pH is within an acceptable range in the surface soil to be planted. Moderate amounts of sediments transported by tides, waves, longshore drift, and upland erosion are beneficial to created marshes. Nutrients associated with sediments increase plant productivity, and accretion helps marsh surfaces keep pace with rising sea level. However, excessive sediment accumulation can damage marsh vegetation by burying plants and increasing surface elevation above the limits of intertidal vegetation (Zedler and West, 2008). Sand blowing onto created marshes may be a problem in some locations, particularly dredged material disposal sites. Installing sand fences and establishing dune vegetation on the upland portion of the site will intercept and stabilize blowing sand to prevent abrupt elevation changes in the intertidal zone.
2.5. Establishing vegetation A self-sustaining plant community is the primary goal of tidal marsh creation since vegetation performs many of the biological functions of the ecosystem (Sullivan, 2001). When the appropriate elevation, hydrology, and soil are in place for the created marsh, establishing the dominant emergent plant species that mimic the target reference marsh accelerates development. Vegetation provides structure and creates an environment favorable for recruitment of invertebrates, microbes, and other flora and fauna that are adapted to the tidal marsh environment. Marsh plants also are important primary producers that are a part of the complex food web of marsh/estuary ecosystems (Wainright et al., 2000). Strategies for establishing vegetation in created tidal marshes are site specific, vary with geographic region and with environmental factors that determine the plant species composition. Generally, the greater the salinity at the site the fewer plant species are present. For example, along the Atlantic Coast of the United States in marshes with salinities near seawater strength, the vegetation of the intertidal zone is a monoculture of Spartina alterniflora with distinct upper and lower elevation limits. Tidal marshes in estuaries farther inland, which are flooded by less saline water, have greater plant species diversity with less distinct zonation of species. The key to planning vegetation establishment is to utilize reference marshes to determine the target plant species, plant community assemblages, salinity and nutrient status of the soil, and their upper and lower elevation limits. Based on this information, establishment methods can be developed to achieve optimum results. Methods of establishing vegetation range from waiting for natural recruitment of seeds or other propagules, to planting all the species identified in the reference marsh. Methods selected should take into consideration environmental factors at
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the site, characteristics of the plant species, cost, and practicality. Creation sites in low-energy environments that are protected from erosion by waves and currents may become vegetated naturally in a few years as seed or other propagules are spread by tides, wind, or animals. Sites exposed to waves and currents may remain un-vegetated indefinitely. Planting the target vegetation has the advantage of producing a cover in a shorter period of time to prevent erosion, and accelerating the process of developing a functioning ecosystem. Rapid establishment of the target vegetation gives it a competitive advantage that helps prevent colonization by undesirable or invasive plant species. In some cases, planting a single species that is easy to propagate and grows rapidly could be the best option for initial cover with the expectation that the natural plant assemblage will develop with time. Introducing plants to quickly establish vegetation is logical since the monetary cost and effort in planting at a tidal marsh creation site is usually small in comparison to the cost of earth moving and other construction (Broome, 1990). If a natural marsh in the area is being destroyed due to construction, plants, chunks of sod, and soils, which may contain seed and rhizomes, can be preserved and transferred to the creation site to speed development (Sullivan, 2001). This method also introduces benthic infauna to the new marsh. Introduction of non-native plant species should be avoided since it may have unintended consequences that negatively impact ecosystems. An example of this is the introduction of S. alterniflora to the west coast of the United States both accidentally in Willapa Bay, Washington from packing material used for shipping eastern US oysters (Major et al., 2003), or planting for shoreline stabilization and salt marsh restoration (Global Invasive Species Database, 2007). S. alterniflora grows at both higher and lower elevations than native marsh vegetation. Impacts to the ecosystem include conversion of mud flat habitat to tidal marsh, trapping sediment, raising elevations, and loss of native plant communities through competition in both low and high marsh. In San Francisco Bay, California, S. alterniflora competes with and hybridizes with the native Spartina foliosa (San Francisco Estuary Invasive Spartina Project, 2001). In some circumstances, introducing non-native plant species may be beneficial. In his review of utilizing Spartina for ecological engineering in China, Chung (2006) documented many benefits of planting several non-native plant species to create tidal marshes. Spartina anglica was brought to China in 1963 followed by the introduction of other species of Spartina including S. alterniflora in 1979. Significant economic, social and, ecological benefits derived from Spartina plantations include coastal stabilization, land reclamation, control of siltation, primary production in marine food webs, esthetic value for tourism, amelioration of saline soil, green manure, animal fodder, fish feed, and biofuel. Qin et al. (1998) reported the health benefits of two products, biomineral liquid and total flavonoids, extracted from Spartina plants. If the goal is to create tidal marsh ecosystems similar to local reference marshes, it is important to obtain planting material from near the creation site to avoid introducing non-native plants or ecotypes that may not be adapted to the local environment. For plant species that produce significant amounts of seed, collect seed from or near the reference marsh. Direct seeding of created sites may be an
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option if seed supplies are sufficient and the planting surface is stable. More often, seed are used to grow plants in containers and then planted at created sites. Commercial production of native marsh plants in nurseries and greenhouses is a common practice in some localities. Alternatively, plants can be dug from natural marshes, but there is risk of negatively impacting the donor marsh. If this method of obtaining planting stock is used, newly established stands on dredged material or recently accreted sediments are often good sources of plants. It is difficult to separate viable individual plants from the thick mat of roots and rhizomes in older established marshes. Planting techniques must be adapted to conditions at the site and the labor available. Mechanical transplanters such as those used for vegetable plants may be used on large sites with soil material that is free of debris and will support equipment. Spades or tree-planting dibbles work well for opening planting holes where planting is done manually. Spacing of plants is an important consideration that is a trade-off between the numbers of plants required and how long it takes to achieve complete cover. For example, a plant spacing of 1 1 m requires 10,000 plants per hectare, while a spacing of 60 60 cm requires 27,778 plants per hectare, and 30 30 cm spacing requires 111,111 plants per hectare. The rate of growth of the plant species is a factor that must be considered. S. alterniflora spreads rapidly, and our experience has shown that 60 60 cm plant spacing results in complete cover in one growing season (7 months). Spacing plants 1 m apart is acceptable on sites that are protected from erosion, with complete cover in about 1 year through spread of rhizomes and germination of seed produced at the end of the first growing season. Fertilization at planting with slow release N and P fertilizers enhances growth of vegetation in created marshes (Broome et al., 1983). Ammonium sources of N should be used rather than nitrate-N. When nitrate is applied to reduced soils, it is subject to atmospheric loss as N gas through denitrification and runoff from tides. Plant growth response to fertilization varies from location to location depending on whether the available soil nutrients, and inputs from sediment accretion, runoff, birds, etc. are adequate for optimum plant growth. Growth response to fertilizer also varies with plant species. For example, the growth rates of grasses like Spartina spp. are more likely to be increased by application of fertilizers than Juncus spp. Maintenance applications of fertilizers may be needed in subsequent growing seasons on sites where nutrients are limiting growth. Surface application of ammonium sulfate and concentrated superphosphate at low tide is a good method to supply N and P until nutrient pools and nutrient cycling are sufficient to maintain a healthy marsh ecosystem. Fertilizer should be applied only when necessary since excessive nutrients due to pollution are a problem in many coastal environments. In Europe, several projects have used seeding and planting to accelerate vegetation colonization and succession. A managed realignment at Tollesbury, Essex, UK, employed seeding at low and high densities as well as plugs and turfs of salt marsh vegetation (Garbutt et al., 2006). However, seeding and transplanting were ineffective at the site because the seeds did not germinate (or they washed away) and the transplants experienced near complete (97%) mortality. Although not a marsh creation project, tidal marsh vegetation was planted along the Brittany coast, France, to re-vegetate salt marshes damaged by the Amoco Cadiz oil spill and
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subsequent clean-up effort (Broome et al., 1988). Sprigs and plugs of Puccinellia maritima and Halimione (Atriplex) portulacoides were successfully established.
3. EVALUATING F UNCTIONAL EQUIVALENCE OF C REATED T IDAL MARSHES If the goal of tidal marsh creation is to create self-sustaining ecosystems that develop similar appearances (Figure 3), structures, values, and functions as natural marshes, criteria for success must be defined and measured. The definition of success is controversial and may mean different things to different people, and determining success is challenging (Kentula, 2000). Functional equivalence can be determined by evaluating whether key ecological functions are similar to reference marshes. Success may be defined as replicating all the functions of reference marshes, while a more reasonable approach might consider replacement of some of the functions as acceptable (Kentula et al., 1992). Important functions of tidal marshes that may be assessed include biological productivity, food webs, and biogeochemical cycles.
3.1. Biological productivity and food webs Because they are detritus-based ecosystems, development of ecological functions of tidal marshes requires high levels of net primary production (NPP) that, over time, accumulates as soil organic matter. The development of an organic-rich surface soil
Figure 3 A three-hectare brackish water (10-15 ppt) marsh created in 1983 by excavating an upland soil in a pine woodland. Greenhouse-grown seedlings of Spartina cynosuroides, S. alterniflora, and S. patens were planted 60 cm apart. Photographed June 12, 2008, 25 years after the project was completed. Photo courtesy of Mr. Jeffrey C. Furness, PCS Phosphate Company Inc.
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horizon promotes colonization by detritus feeding benthic invertebrates that supports higher-level consumers, finfish, and wading birds. Soil organic matter also is necessary to support heterotrophic microbial processes, organic matter mineralization, methanogenesis, and denitrification. Development of biological productivity and food webs first requires establishing vegetation with high NPP. In some cases, saltwater marshes achieve aboveground NPP comparable to natural marshes within 3–5 years following creation (Broome et al., 1986, Figure 4a, Table 1). Aboveground biomass develops faster than belowground biomass (Figure 4b). In other cases, the plant community and NPP develop slowly or not at all because of problems associated with recreating tidal hydrology and wetland soil characteristics, especially adequate organic matter and N (Zedler, 2001). NPP of marshes created on graded terrestrial soils may not develop to levels found in natural marshes because of acidity and low fertility of the planting substrate (Broome et al., 1988). On these sites, liming and fertilizing with N and P may be necessary for satisfactory plant growth (Broome et al., 1988). Development of NPP is related to the frequency of tidal inundation, and NPP develops more quickly in salt marshes where tidal inundation is frequent and regular (occurs twice a day) as compared to marshes where tidal inundation is infrequent and irregular. On the marsh levee where inundation occurs frequently, NPP developed to levels similar to natural marshes within 3 years (Broome et al., 1988; Craft et al., 2002). At the highest elevations where inundation occurs only during spring tides and storm tides, NPP never consistently achieved equivalence to natural marshes even after 15 years (Craft et al., 2002). Plant species composition also affects the development of NPP. Biomass of C3 vegetation such as black needlerush (Juncus roemerianus), which grows slowly, takes longer to achieve equivalence to natural Juncus marshes (10 years) than faster growing C4 Spartina spp. (Broome, 1988; Craft et al., 2002). Macrophyte biomass and NPP do not develop quickly on all sites. In southern California, vegetation development is slowed by low N in soil (Langis et al., 1991) and annual additions of N were needed to maintain high levels of biomass production (Boyer and Zedler, 1998). Species richness, which is greater in southern California salt marshes than in Atlantic coast marshes, also was slow to develop in west coast marshes. Whereas Salicornia virginica naturally recruited into created marshes, other native species required seeding or transplanting (Lindig-Cisneros and Zedler, 2002). In southern California, transplanting native species to increase species richness accelerated development of biomass stocks and N accumulation in created marsh soils (Callaway et al., 2003). In Europe, natural colonization of dredge spoil and areas reclaimed by managed realignment occurs by germination of propagules in the seed bank and by dispersal from nearby natural marshes. Pioneer species consist of Sueda maritima, Salicornia sp., and Atriplex sp. that colonize the site within 5 years (Bernhardt and Handke, 1992). During succession (5–10 years), pioneer species are replaced by salt marsh species, Puccinellia maritima, Trigloglin maritima, and Aster sp. Natural colonization of a managed realignment (UK) occurred quickly and within 6 years after breaching, the site contained 15 different species of plants and 6 ha of the 21 ha site was colonized by salt marsh vegetation (Garbutt et al., 2006; Chapter 27). Nineteen
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Figure 4 (a) Aboveground biomass of Spartina alterniflora along a chronosequence of created salt marshes and natural reference marshes. Asterisks ( ) indicate that the created marsh and paired reference marsh are significantly different (P 0.05) according to Student’s t test. Dashed lines represent the range of values measured in the natural marshes. (b) Macro-organic matter (MOM, living and dead root and root/ rhizome mat) biomass of Spartina alterniflora along a chronosequence of created salt marshes and natural reference marshes. Asterisks ( ) indicate that the created marsh and paired reference marsh are significantly different (P 0.05) according to Student’s t test. Dashed lines represent the range of values measured in the natural marshes (from Craft et al., 2003).
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Table 1 Estimated rate of development of wetland dependent functions following saltwater marsh creation Time needed to achieve equivalence to natural marshesa (years) Ecological functions Primary productionb Secondary production Benthic invertebratesc Finfish and shellfishd Avifauna use Waterfowl and wading birdse Songbirdse Outwellingf Biogeochemical functions Sedimentationg N retentionh P retentioni Carbon sequestrationj C mineralizationk Methanogenesisl Denitrificationm
3–5 10–20 2–>15 1–3 10–15 3–5 1–3 3–5 1–3 3–5 5–15 5–15 years 5–15 years
a
See text for an explanation of the rationale for choosing the timescales. Broome et al. (1986, 1988), Craft et al. (2002, 2003). Moy and Levin (1991), Levin et al. (1996), Scatolini and Zedler (1996), Posey et al. (1997), Craft et al. (1999), Craft and Sacco (2003). d Moy and Levin (1991), Rulifson (1991), Minello and Zimmerman (1992, 1997), Havens et al. (1995), Williams and Zedler (1999), Talley (2000). e Havens et al. (1995). f Craft et al. (1989). g Craft (1997), Craft et al. (2003). h Lindau and Hossner (1981), Craft et al. (1988, 1999, 2002, 2003), Craft (1997). i Craft et al. (1988, 1999, 2003), Craft (1997), Poach and Faulkner (1998). j Lindau and Hossner (1981), Craft et al. (1988, 1999, 2002, 2003), Craft (1997). k Craft et al. (2003), Cornell et al. (2007). l C. Craft (unpublished data), Cornell et al. (2007). m Thompson et al. (1995), Currin et al. (1996), Figure 7 of this study. b c
species of benthic invertebrates had also recruited to the site after 6 years, with Hydrobia ulvae identified as the dominant species. Breaching of the embankment also led to sedimentation in the newly reclaimed marsh that facilitated soil accretion during the 6-year monitoring period 1995–2001. The rate of colonization of reclaimed marshes varied depending on size, tide range, and the time since breaching occurred. Species richness was greater in larger sites (>100 ha) than for smaller sites (<30 ha) (Wolters et al., 2005). Also, sites with the largest elevation range within the tidal prism (mean high water neap to mean high water spring tide) contained more species than sites with less elevation range. Comparison of salt marsh vegetation (Puccinellia maritima, H. (Atriplex) portulacoides) planted following cleanup of the Amoco Cadiz oil spill (France) revealed that, after 4 years, natural colonization of the site was very slow and that planting was necessary to accelerate succession and restore the site (Broome et al., 1988). Across Europe, studies indicate that considerable time must pass before reclaimed marshes develop plant communities that are comparable to natural
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marshes. Miren et al. (2001) observed that, whereas 20- and 35-year-old reclaimed marshes in Spain contained comparable numbers of species (16–17), species richness was much lower than the 36 species recorded in a natural salt marsh. Establishment of vegetation and NPP are prerequisite to colonization by marsh consumers. Many created marshes contain significantly fewer benthic invertebrates than comparable natural marshes (Moy and Levin, 1991; Sacco et al., 1994; Levin et al., 1996; Scatolini and Zedler, 1996) although there is a trend toward increasing density with marsh age (Figure 5, LaSalle et al., 1991; Simenstad and Thom, 1996; Craft et al., 1999). Surface deposit feeding invertebrates, crabs (Uca spp., Hemigrapsus spp.) and polychaete worms, colonize created marshes within a few years (Moy and Levin, 1991; Levin et al., 1996; Scatolini and Zedler, 1996; Posey et al., 1997). These organisms feed at the marsh surface and use emergent stems for cover and refuge from predators. However, subsurface deposit feeders such as oligochaete worms and dipteran larvae were slower to colonize (Moy and Levin, 1991; Levin et al., 1996; Scatolini and Zedler, 1996; Posey et al., 1997). Low densities of subsurface deposit feeders may be the result of insufficient soil organic matter (Moy and Levin, 1991; Levin et al., 1996) and up to 20 years may elapse before created marshes develop an organic-rich surface layer that supports densities of subsurface deposit feeders that are comparable to natural marshes (Table 1, Craft et al., 1999; Craft and Sacco, 2003). Studies of shellfish and finfish use of created marshes yield mixed findings. Young (2–5 years old) created salt marshes in Texas contained significantly fewer numbers of grass shrimp (Palaemonetes pugio) and brown shrimp (Penaeus aztecus) than natural marshes (Minello and Zimmerman, 1992). Likewise, a 5-year-old constructed marsh in Virginia contained fewer blue crabs (Callinectes sapidus) than two comparable natural marshes (Havens et al., 1995). Both studies, though, reported no difference in densities of small fish between the constructed and natural marshes. Comparable studies of created and natural marsh creek channels in San Diego Bay and Pamlico River (NC) also revealed no differences in finfish density or species composition (Rulifson, 1991; Williams and Zedler, 1999). Other studies though, report significantly lower density of finfish in created versus natural marshes. In southern California, Talley (2000) measured fewer finfish but similar species richness and dominance in 1- to 3-year-old created marsh creeks relative to natural marsh creeks. A 3-year-old created salt marsh in North Carolina also contained significantly fewer killifish (Fundulus heteroclitus) than natural marshes (Moy and Levin, 1991). Differences in trophic dynamics also were noted. Gut analysis revealed that Fundulus fed mostly on polychaetes and algae in the created marsh whereas, in the natural marsh, gut contents were mostly detritus and insects. Minello and Webb (1997) compared finfish and shellfish populations in five natural marshes and 10 created marshes (3–15 years old) in Texas and found that created marshes contained significantly fewer finfish (gobies and pinfish, Lagodon rhomboides) and crustaceans, including commercially important blue crabs, white shrimp (Penaeus setiferus), and brown shrimp (Penaeus aztecus). In constructed marshes, lower densities
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Figure 5 (a) Density (no./m2 – SE) of benthic infauna along a chronosequence of constructed salt marshes and natural reference marshes. Asterisks ( ) indicate that the created marsh and paired reference marsh are significantly different (P 0.05) according to Student’s t test. Dashed lines represent the range of values measured in the natural marshes. (b) Species richness (no. 7.07 cm ^2 – SE) of benthic infauna along a chronosequence of constructed salt marshes and natural reference marshes. Asterisks ( ) indicate that the created marsh and paired reference marsh are significantly different (P 0.05) according to Student’s t test. Dashed lines represent the range of values measured in the natural marshes (from Craft and Sacco, 2003).
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were attributed to higher surface elevation that reduced tidal flooding and limited entry of estuarine nekton. Proper site preparation and grading to create low elevation marsh habitat was recommended for future projects to increase nekton access to the marsh (Minello and Webb, 1997). Other recommendations to increase finfish use of created marshes include increasing channel heterogeneity to create a variety of habitats (Williams and Zedler, 1999) and maximize vegetated edge to facilitate access to the marsh (Minello et al., 1994). With respect to finfish and shellfish utilization, it appears that created marshes require anywhere from two to more than 15 years to achieve equivalence to natural marshes (Table 1). Use of created marshes by birds has received less attention. A Virginiaconstructed marsh attracted similar numbers of wading birds but fewer numbers of songbirds than nearby natural marshes (Havens et al., 1995). In natural marshes, songbird utilization was attributed to the presence of shrubs (Iva frutescens, Baccharis halimifolia) that provided more perching habitat. Wading bird and waterfowl utilization of created marshes occur relatively quickly, in 1–3 years whereas songbird use of created marshes takes longer since it depends on establishment of woody shrubs in the marsh (Table 1). As nesting habitat, created marshes often lack the structural complexity needed to support breeding requirements of certain species. In Virginia, Marsh Wrens (Cistothorus palustris) exhibited a strong preference for nesting in natural marshes versus a constructed marsh (Havens et al., 1995). In southern California, the federally endangered Light-footed clapper rail (Rallus longirostris levipes) nests only in S. foliosa marshes of southern California (Zedler, 1993). Zedler found that S. foliosa marsh created to mitigate for wetland loss from highway construction activities did not provide nesting habitat for the rail. In the created marsh, the Spartina stems were shorter, 60–90 cm tall, than in the natural marsh where most stems were greater than 90 cm in height. Clapper rails are unable to establish nests in the short Spartina canopy because nests are washed out of the lower marsh by high tides (Zedler, 1993). Onetime N additions were found to increase growth and produce stems (100 stems per m2 with 30 stems per m2 taller than 90 cm) that were sufficiently tall for nesting (Zedler, 1993; Boyer and Zedler, 1998). However, the following year, fertilized Spartina plots produced 30 “tall” stems per m2, indicating that existing marsh soil and plant N pools were inadequate to sustain improved growth and taller canopies over the long term without annual N additions (Boyer and Zedler, 1998). Like natural marshes, created tidal marshes are a source of detritus to estuarine waters. Two created brackish water marshes in North Carolina exported dissolved organic C (DOC) and N (DON) to the tidal creek and imported inorganic N (NH4) and P (PO4) (Craft et al., 1989). Overall, the marshes were a source of organic C to the estuary whereas, with N, they served as transformers by importing NH4, converting it to organic form and exporting it as DON. One marsh, amended with soil organic matter prior to planting, exported much more DOC and DON than the other marsh, which contained little soil organic matter (Craft et al., 1989).
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3.2. Biogeochemical cycles Created tidal marshes, as with natural marshes, trap sediment and nutrients (N, P), sequester organic matter and C and are hotspots for microbial processes such as denitrification, Fe/Mn reduction, and sulfate reduction. Biogeochemical functions develop slowly on wetlands created from upland soils relative to those created on dredged material because of the large amount of oxidized Fe that must be reduced (Craft et al., 1991) before other reducing reactions (e.g., sulfate reduction) will occur. Created marshes are active sites of sediment deposition. Once vegetation becomes established, aboveground stems dramatically reduce wave energy, which facilitates deposition of inorganic sediment. Young created marshes often trap sediment at rates exceeding those of natural marshes (Table 2). During the 6 months following establishment of vegetation, a created marsh trapped 31 kg sediment per m2, increasing marsh elevation by 11 mm (C.B. Craft, unpublished data). In a nearby natural marsh, sediment deposition was much lower, 12 kg/m2, during the same period. Sedimentation also deposits substantial amounts of inorganic nutrients. In the same created marsh, sedimentation contributed 7.7 g P/m2 to the marsh surface during the first 6 months as compared to 3.5 g P/m2 in the nearby natural marsh. Created tidal marsh soils also are sinks for N (Figure 6a). Young created marshes often accumulate N at rates comparable to or exceeding those in natural marshes (Craft et al., 1999). N accumulates at higher rates at more frequently inundated low elevations of the marsh than at higher elevations where inundation is less frequent (Lindau and Hossner, 1981; Craft et al., 2002). Enhanced N accumulation in young created marshes is a result of greater epiphytic (Currin and Paerl, 1998) and benthic (Piehler et al., 1998) N fixation in response to low soil N and N limitation (Craft et al., 1999). In spite of accelerated N accumulation, many created marsh soils still contain less N than natural marshes even after 30 years (Craft et al., 1999). P accumulation in created and natural tidal marshes is less than N, usually less than 1 g/m2/year (Table 2). And, like N, young created marshes often accumulate P at higher rates than older created marshes and natural marshes (Craft, 1997). Sorption of Table 2 Comparison of selected biogeochemical processes in created and natural tidal salt marshes
a
Process (g/m2/year)
Createda
Naturala,b
Sediment deposition (g/m2/year) Soil P accumulation (g/m2/year) Soil N accumulation (g/m2/year) Soil C sequestration (g/m2/year) Organic matter qualityd Denitrification (g/m2/year)c
1,000–11,000 0–0.9 7.1–17 80–125 15–20% lignin <0.1
160–2,700 0–2.3 1.3–23 21–393 >30% lignin 0–3
Cammen (1976), Craft (1997), Craft et al. (1999, Table 7), C.B. Craft (unpublished data). Craft et al. (1999, Table 7). Currin et al. (1996), Craft (2001). d C.B. Craft (unpublished data). b c
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R 2 = 0.69 (p ≤ 0.05)
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Figure 6 (a) Soil N pools along a chronosequence of created salt marshes and natural reference marshes. Asterisks ( ) indicate that the created marsh and paired reference marsh are significantly different (P 0.05) according to Student’s t test. Dashed lines represent the range of values measured in the natural marshes. (b) Soil organic carbon pools along a chronosequence of created salt marshes and natural reference marshes. Asterisks ( ) indicate that the created marsh and paired reference marsh are significantly different (P % 0.05) according to Student’s t test. Dashed lines represent the range of values measured in the natural marshes (from Craft et al., 2003).
P to Al or Ca bearing minerals enhances P accumulation in soil. A brackish water marsh created on a graded terrestrial soil high in Al and Fe sorbed floodwater PO4-P during tidal inundation (Craft et al., 1989). In the Mississippi River Delta, 15- to 20year-old marshes created on dredged material contain more Ca-bound P than younger (1- to 3-year-old) marshes because of increased sorption to Ca in
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floodwaters (Poach and Faulkner, 1998). Salt marshes created on dredged material or shorelines often contain calcareous shell fragments that facilitate P sorption. For example, 10-year-old marsh created on a soil containing numerous shell fragments in North Carolina sorbed 6–8 g P/m2/year (Craft, 1997). In created marshes, organic forms of P also accrue in accumulating soil organic matter, (Poach and Faulkner, 1998) although rates are relatively low, 0.2–1.0 g P m2/year (Craft, 1997). Soil organic matter and C sequestration also increase with created marsh age (Figure 6b). Once vegetation becomes established and NPP develops, created marshes begin to accumulate organic matter and C at rates comparable to natural marshes (Table 2). Like N, soil organic C pools increase with created marsh age but, after 30 years, still have not achieved equivalence to natural marshes (Figure 6b). And, organic C accumulation is greater at low elevations in the marsh that are frequently inundated as compared to higher elevations (Lindau and Hossner, 1981; Craft et al., 2002). An interesting difference between created and natural marshes is that, in young created marshes, the accumulating organic matter is of higher quality. Soil organic matter in young (<10 years old) created marshes contain mostly labile organic compounds like carbohydrates and water soluble compounds and relatively little lignin (Craft et al., 2003, Table 2) whereas, in natural marshes, organic matter contains proportionally more lignin (>30%) that is recalcitrant and not readily decomposable. Development of heterotrophic microbial activity in created tidal marshes is strongly linked to accumulation of soil organic matter. Comparison of microbial processes and soil organic C stocks in created and natural marshes revealed that C mineralization increased with increasing percent soil organic C (R2 = 0.69, p < 0.001) (Craft et al., 2003). Rates of denitrification (R2 = 0.63, p < 0.01) and
Denitrification rate (ng/g dw soil/day)
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R 2 = 0.84 (p = 0.01)
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Figure 7 Potential rates of denitrification along a chronosequence of created salt marshes and natural reference marshes. Asterisks ( ) indicate that the created marsh and paired reference marsh are significantly different (P 0.05) according to Student’s t test. Dashed lines represent the range of values measured in the natural marshes.
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methanogenesis (R2 = 0.64, p < 0.01) also increased with increasing soil organic C (Cornell et al., 2007; C. Craft, unpublished data). In general, denitrification is lower in young created marshes as compared to natural marshes (Table 2). Thompson et al. (1995) measured denitrification rates that were 44 times lower in a 2-year-old created S. alterniflora marsh than a nearby natural marsh and Currin et al. (1996) measured low denitrification rates on the same created marsh. In some created marshes, denitrification is inhibited by coarse (sandy) texture of the soil that provides less surface area for microbial populations, greater tidal flushing of pore water nutrients, and more exposure to oxygen relative to fine-textured natural marsh soils (Thompson et al., 1995). However, as created marshes age and soil organic matter accumulates, denitrification increases to levels comparable to levels measured in natural marshes (Figure 7).
4. SUMMARY Successful creation of tidal marshes, which become functionally equivalent to natural marshes in a reasonable length of time, requires careful attention to establishing the target biotic communities. Important steps are site selection, assessment of a reference marsh, design and construction, obtaining plant propagules, proper planting, plant spacing, providing adequate nutrients from fertilizers if natural nutrient supplies are low, and maintenance. Tidal marsh creation methods and basic ecological principles must be adapted to the unique physical conditions and biological communities that occur at each creation site. There is a need for continued research with careful monitoring, documentation of results, and synthesis of information from tidal marsh creation projects in a variety of coastal environments. Some issues that require further investigation include the following: (1) defining design criteria including geomorphic, hydrologic, and biotic factors; (2) developing models for designing tidal creeks and predicting hydrology; (3) the effects of adjacent uplands on hydrology; (4) the value of creating transitional ecotones; (5) the effects of soils, sediments, and plant-available nutrients on success; (6) defining measures of success and determining the time required to achieve structural and functional equivalency to reference wetlands for different marsh types; (7) comparing the benefits of planting native vegetation with natural recruitment; and (8) developing methods for accelerating development of created marsh ecosystems. When sound principles of ecological engineering are applied, tidal marshes can be created that have the same appearance and, with time, provide many of the functions and values of natural marshes.
REFERENCES Bernhardt, K.G., Handke, P., 1992. Successional dynamics of newly created saline marsh soils. Ekologia 11, 139–152. Boorman, L., Hazelden, J., Boorman, M., 2002. New salt marshes for old – salt marsh creation and management. In: Eurocoast/eucc, Porto – Portugal Ed. Littoral 2002 The Changing Coast. Eurocoast – Portugal, pp. 35–45.
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Boyer, K.E., Zedler, J.B., 1998. Effects of nitrogen additions on the vertical structure of a constructed cordgrass marsh. Ecol. Appl. 8, 692–705. Broome, S.W., 1990. Creation and restoration of wetlands of the southeastern United States. In: Kusler, J.A., Kentula, M.E. (Eds.), Wetland Creation and Restoration. Island Press, Washington, DC, pp. 37–72. Broome, S.W., Seneca, E.D., Woodhouse Jr., W.W., 1983. The effects of source rate and placement of nitrogen and phosphorus fertilizers on growth of Spartina alterniflora transplants in North Carolina. Estuaries 6, 212–226. Broome, S.W., Seneca, E.D., Woodhouse Jr., W.W., 1986. Long-term growth and development of transplants of the salt marsh grass Spartina alterniflora. Estuaries 9, 63–74. Broome, S.W., Seneca, E.D., Woodhouse Jr., W.W., 1988. Tidal salt marsh restoration. Aquat. Bot. 32, 1–22. Callaway, J.C., 2001. Hydrology and substrate. In: Zedler, J.B. (Ed.), Handbook for Restoring Tidal Wetlands. CRC Press, Boca Raton, FL. Callaway, J.C., Sullivan G., Zedler, J.B., 2003. Species-rich plantings increase biomass and nitrogen accumulation in a wetland restoration experiment. Ecol. Appl. 13, 1626–1639. Cammen, L.M., 1976. Accumulation rate and turnover time of organic carbon in a salt marsh sediment. Limnol. Oceanogr. 20, 1012–1015. Chung, C.H., 2006. Forty years of ecological engineering with Spartina plantations in China. Ecol. Eng. 27, 49–57. Cornell, J.A., Craft, C., Megonigal, J.P., 2007. Ecosystem gas exchange across a created salt marsh chronosequence. Wetlands 27, 240–250. Craft, C.B., Broome, S.W., Seneca, E.D., 1988. Nitrogen, phosphorus and organic carbon pools in natural and transplanted marsh soils. Estuaries 11, 272–280. Craft, C.B., 1997. Dynamics of nitrogen and phosphorus retention during wetland ecosystem succession. Wetl. Ecol. Manage. 4, 177–187. Craft, C.B., 2001. Biology of wetland soils. In: Richardson, J.L., Vepraskas, M.J. (Eds.), Wetland Soils: Their Genesis Hydrology, Landscape and Separation into Hydric and Nonhydric Soils. CRC Press, Boca Raton, FL, pp. 107–135. Craft, C.B., Broome, S.W., Campbell, C.L., 2002. Fifteen years of vegetation and soil development following brackish-water marsh creation. Restor. Ecol. 10, 248–258. Craft, C.B., Broome, S.W., Seneca, E.D., 1989. Exchange of nitrogen, phosphorus and organic carbon between transplanted marshes and estuarine waters. J. Environ. Qual. 18, 206–211. Craft. C.B., Sacco, J.N., 2003. Long-term succession of benthic infauna communities on constructed Spartina alterniflora marshes. Mar. Ecol. Prog. Ser. 257, 45–58. Craft, C.B., Megonigal, J.P., Broome, S.W., Cornell, J., Freese, R., Stevenson, R.J., Zheng, L., Sacco, J., 2003. The pace of ecosystem development of constructed Spartina alterniflora marshes. Ecol. Appl. 13, 1417–1432. Craft, C.B., Seneca, E.D., Broome, S.W., 1991. Porewater chemistry of natural and created marsh soils. J. Exp. Mar. Biol. Ecol. 152, 187–200. Craft, C.B., Reader, J.M., Sacco, J.N., Broome, S.W., 1999. Twenty-five years of ecosystem development of constructed Spartina alterniflora (Loisel) marshes. Ecol. Appl. 9, 1405–1419. Crooks, S., Schutten J., Sceern, G.D., Pye, K., Davy, A.J., 2002. Drainage and elevation as factors in the restoration of salt marsh in Britain. Restor. Ecol. 10, 591–602. Currin, C.A., Joye, S.B., Paerl, H.W., 1996. Diel rates of N2-fixation and denitrification in a transplanted Spartina alterniflora marsh: Implications for N-flux dynamics. Estuar. Coast. Shelf Sci. 42, 597–616. Currin, C.A., Paerl, H.W., 1998. Epiphytic nitrogen fixation associated with standing dead shoots of smooth cordgrass, Spartina alterniflora. Estuaries 21, 108–117. Darnell, T.M., Smith, E.H., 2001. Recommended design for more accurate duplication of natural conditions in salt marsh creation. Environ. Manage. 29, 813–823. Garbutt, R.A., Reading, C.J., Wolters, M., Gray, A.J., Rothery, P., 2006. Monitoring the development of intertidal habitats on former agricultural land after the managed realignment of coastal defences at Tollesbury, Essex, UK. Mar. Pollut. Bull. 53, 155–164. Global Invasive Species Database, 2007. Spartina alterniflora (grass). http://www.issg.org. Accessed date May, 2008.
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Harvey, J.W., Odum, W.E., 1990. Groundwater inputs to coastal waters. Biogeochemistry 10, 217– 236. Havens, K.J., Varnell, L.M., Bradshaw, J.G., 1995. An assessment of ecological conditions in a constructed tidal marsh and two natural reference tidal marshes in coastal Virginia. Ecol. Eng. 4, 117–141. Hofstede, J.L.A., 2003. Integrated management of artificially created salt marshes in the Wadden Sea of Schleswig-Holstein, Germany. Wetl. Ecol. Manage. 11, 183–194. Hood, G.W., 2007. Scaling tidal channel geometry with marsh island area: A tool for habitat restoration, linked to channel formation process. Water Resour. Res. 43, W03409. doi:10.1029/2006WR005083. Hough, P., Robertson, M., 2008. Mitigation under Section 404 of the Clean Water Act: where it comes from, what it means. Wetl. Ecol. Manage (in press). doi:10.1007/s11273-008-9093-7. http://www.springerlink.com/content/ag615v755494325v. Accessed date May, 2008. Kentula, M.E., 2000. Perspectives on setting success criteria for wetland restoration. Ecol. Eng. 15, 199–209. Kentula, M.E., Brooks, R.P., Gwin, S.E., Holland, C.C., Sherman, A.D., Sifneas, J.C., 1992. An Approach to Improving Decision Making in Wetland Creation and Restoration. Island Press, Washington, DC. Landin, M.C., Webb, J.W., Knutson, P.L., 1989. Long-term monitoring of eleven Corps of Engineeers habitat development field sites built of dredged material, 1974–1987. Technical Report K-89-1, Department of the Army, Waterways Experiment Station, Vicksburg, MS. Langis, R., Zalejko, M., Zedler, J.B., 1991. Nitrogen assessment in a constructed and natural salt marsh of San Diego Bay. Ecol. Appl. 1, 40–51. LaSalle, M.W., Landin, M.C., Sims, J.G., 1991. Evaluation of the flora and fauna of a Spartina alterniflora marsh established on dredged material in Winyah Bay, South Carolina. Wetlands 11, 191–208. Lewis III, R.R., 1990. Wetlands restoration/creation/enhancement terminology: suggestions for standardization. In: Kusler, J.A., Kentula, M.E. (Eds.), Wetland Creation and Restoration: The Status of the Science. Island Press, Washington, DC, pp. 417–422. Levin, L.A., Talley, D., Thayer, G., 1996. Succession of macrobenthos in a created salt marsh. Mar. Ecol. Prog. Ser. 141, 67–82. Lindau, C.W., Hossner, L.R., 1981. Substrate characterization of an experimental marsh and three natural marshes. Soil Sci. Soc. Am. J. 45, 1171–1176. Lindig-Cisneros, R., Zedler, J.B., 2002. Halophyte recruitment in a salt marsh restoration site. Estuaries 25, 1174–1183. Major III, W.W., Grue, C.E., Grassley, J.M., Conquest, L.L., 2003. Mechanical and chemical control of smooth cordgrass in Willapa Bay, Washington. J. Aquat. Plant Manage. 41, 6–12. Maryland Department of the Environment Wetlands and Waterways Program, 2006. Shore Erosion Control Guidelines Marsh Creation. http://www.mde.state.md.us/Programs/WaterPrograms/ Wetlands_Waterways. Accessed date May, 2008. Mendelssohn, I.A., Kuhn, N.L., 2003. Sediment subsidy: effects on soil-plant responses in a rapidly submerging coastal salt marsh. Ecol. Eng. 21, 115–128. Minello, T.J., Able, K.W., Weinstein, M.P., Hays, C.G., 2003. Salt marshes as nurseries for nekton: testing hypotheses on density, growth and survival through meta-analysis. Mar. Ecol. Prog. Ser. 246, 9–59. Minello, T.J., Matthews, G.A., Caldwell, P.A., Rozas, L.P., 2008. Population and production estimates for decapod crustaceans in wetlands of Galveston Bay, Texas. Trans. Am. Fish. Soc. 137, 129–146. Minello, T.J., Webb Jr., J.W., 1997. Use of natural and created Spartina alterniflora salt marshes by fishery species and other aquatic fauna in Galveston Bay, Texas, USA. Mar. Ecol. Prog. Ser. 151, 165–179. Minello, T.J., Zimmerman, R.J., 1992. Utilization of natural and transplanted Texas marshes by fish and decapod crustaceans. Mar. Ecol. Prog. Ser. 90, 273–285. Minello, T.J., Zimmerman, R.J., Medina, R., 1994. The importance of edge for natant macrofauna in a created salt marsh. Wetlands 14, 184–198. Miren, O., Albizu, I., Amezaga, I., 2001. Effect of time on the natural regeneration of salt marsh. Appl. Veg. Sci. 4, 247–256.
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Mitsch, W.J., Gosselink, J.G., 2000. Wetlands, third ed. John Wiley, New York. Moller, M.A., Phil, M. Spencer, T., French, J.R., Dixon, M., 2001. The sea-defence value of salt marshes: field evidence from North Norfolk. Water Environ. J. 15, 109–116. Moy, L.D., Levin, L.A., 1991. Are Spartina marshes a replaceable resource? A functional approach to evaluation of marsh creation efforts. Estuaries 14, 1–16. Nuttle, W.K., Harvey, J.W.1995. Fluxes of water and solute in a coastal wetland sediment. 1. The contribution of regional groundwater discharge. J. Hydrol. 164, 89–107. Piehler, M.F., Currin, C.A., Cassanova, R., Paerl, H.W., 1998. Development and N2 fixing activity of the benthic microbial community in transplanted Spartina alterniflora marshes in North Carolina. Restor. Ecol. 6, 290–296. Poach, M.E., Faulkner, S.P., 1998. Soil phosphorus characteristics of created and natural wetlands in the Atchafalaya Delta, Louisiana. Estuar. Coast. Shelf Sci. 46, 195–203. Posey, M.H., Alpin, T.D., Powell, C.M., 1997. Plant and infaunal communities associated with a created marsh. Estuaries 20, 42–47. Qin, P., Xie, M., Jiang, Y.1998. Spartina green food ecological engineering. Ecol. Eng. 11, 147–156. Rulifson, R.A., 1991. Finfish utilization of man-initiated and adjacent natural creeks of South Creek estuary, North Carolina using multiple gear types. Estuaries 14, 447–464. Sacco, J.N., Seneca, E.D., Wentworth, T.R., 1994. Infaunal community development of artificially established salt marshes in North Carolina. Estuaries 17, 489–500. San Francisco Estuary Invasive Spartina Project, 2001. Introduced Spartina alterniflora/hybrids. Coastal Conservancy. http://www.spartina.org. Accessed date May, 2008. Scatolini, S.R., Zedler, J.B., 1996. Epibenthic invertebrates of natural and constructed marshes of San Diego Bay. Wetlands 16, 24–37. Simenstad, C.A., Thom, R.M., 1996. Functional equivalency trajectories of the restored Gog-Le-HiTe estuarine wetland. Ecol. Appl. 6, 38–56. Streever, W.J., 1999. An International Perspective on Wetland Rehabilitation. Kluwer Academic Publishers, Dorecht, The Netherlands. Streever, B., 2000. Dredged material marshes: summary of three research projects. Wetl. Res. Bull. Vol CRWRP-2 No.1. pp. 1–4. Sullivan, G., 2001. Establishing vegetation in restored and created coastal wetlands. In: Zedler, J.B. (Ed.), Handbook for Restoring Tidal Wetlands. CRC Press, Boca Raton, FL. Talley, D.M., 2000. Ichthyofaunal utilization of newly-created versus natural salt marsh creeks in Mission Bay, California. Wetl. Ecol. Manage. 8, 117–132. Thompson, S.P., Paerl, H.W., Go, M.C., 1995. Seasonal patterns of nitrification and denitrification in a natural and a restored salt marsh. Estuaries 18, 399–408. Wainright, S.C., Weinstein, M.P., Able, K.W., Currin, C.A., 2000. Relative importance of benthic mcroalgae, phytoplankton and the detritus of smooth cordgrass Spartina alterniflora and the common reed Phragmites australis to brackish-marsh food webs. Mar. Ecol. Prog. Ser., 200, 77–91. Williams, G.D., Zedler, J.B., 1999. Fish assemblage composition in constructed and natural tidal marshes of San Diego Bay: relative influence of channel morphology and restoration history. Estuaries 22, 702–716. Wolters, M., Garbutt, A., Bakker, J.P., 2005. Salt-marsh restoration: evaluating the success of deembankments in north-west Europe. Biol. Conserv. 123, 249–268. Zedler, J.B., 1993. Canopy architecture of natural and planted cordgrass marshes: selecting habitat evaluation criteria. Ecol. Appl. 3, 123–138. Zedler, J.B. (Ed.), 2001. Handbook for Restoring Tidal Wetlands. CRC Press, Boca Raton, Florida. Zedler, J.B., West, J.M., 2008. Declining diversity in natural and restored salt marshes: A 30-year study of Tijuana estuary. Rest. Ecol. 16, 249–262. Zeff, M.L., 1999. Salt marsh tidal channel morphometry: applications for wetland creation and restoration. Restor. Ecol. 7, 205–211.
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Contents 1. 2. 3. 4. 5.
Introduction Setting Objectives Planning for the Future Addressing Causes and not Symptoms Managing Disturbance 5.1. Restoring hydrology 5.2. Managing weeds 5.3. Introduced fauna 5.4. Grazing 5.5. Pollution 6. Conflicting Priorities 7. Discussion 8. Conclusions References
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1. INTRODUCTION There is a very substantial literature on salt marsh restoration and re-creation, which, as well as numerous site-specific accounts, includes reviews by Lewis (1982), Zedler (1995, 2001), Bakker et al. (1997), Niering (1997), Atkinson et al. (2001), Bakker et al. (2002), Zedler and Adam (2002), Boorman (2003), Callaway (2005), Nottage and Robertson (2005), Bakker and Piersma (2006), and Wolters (2006). The considerable investment in salt marsh restoration reflects the high degree of public awareness of the ecological values of healthy salt marsh ecosystems and increasing recognition of the extent of absolute loss or degradation of many sites. Estuaries have been a focus for human settlement for centuries (Adam, 1990, 2002; Marshall, 2004) resulting in loss of large areas of salt marsh through reclamation, and been subject to degradation through past industrial and waste disposal practices. Sites which have restoration include those in urban areas, both small (Stricker, 1995) and large (Marshall, 2004; US Army Corps of Engineers, 2005), Coastal Wetlands: An Integrated Ecosystem Approach
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and some of which, while affected by human activity, are outside urban areas (Weinstein et al., 2000; Nottage and Robertson, 2005). Rather than reviewing the extensive case studies, I will address a number of issues which have not always been fully considered in project planning. The issues which need to be considered are also relevant in freshwater tidal wetlands, and the discussion by Baldwin et al. (2009) provides a complement to this account. The majority of published studies on salt marsh restoration are from North America and Europe, but degraded salt marshes which could be rehabilitated occur widely. In general, lessons which can be learnt from projects at specific sites are probably broadly applicable, although restoration of high-latitude marshes might be more difficult than at lower latitudes with more optimal conditions. Even when the object of restoration is to benefit a component of the fauna (such as birds), projects normally consist of manipulation of the physical environment and vegetation, rather than involving direct handling of animals. Invertebrate and microbial assemblages are important components of salt marsh biota, with major ecosystem process roles (Christian et al., 1981; Montague et al., 1981; Wiebe et al., 1981). Microbial assemblages in disturbed polluted salt marshes differ from those in unpolluted marshes (Cao et al., 2006). The development of microbial communities in transplanted Spartina alterniflora marshes was studied by Piehler et al. (1998): and after 6 years the community was still not identical to the control site. Development of benthic infauna assemblages has been studied in a number of restored wetlands (e.g., see Levin et al., 1996; Craft, 2000; Craft and Sacco, 2003; Moseman et al., 2004). Although initial colonization may be rapid, convergence to reference assemblages may take many years, (25 years in some cases), well beyond the period normally specified for postconstruction monitoring (Craft and Sacco, 2003). Levin et al. (1996) suggested that for species with poor dispersal “seeding” was an option to speed up community development. While a great deal is known about species composition and ecosystem processes in salt marshes in North America, in other parts of the world there is often far less information available for potential reference and degraded sites.
2. SETTING OBJECTIVES A great many restoration projects have been conducted in an ad hoc fashion with lack of goals and with no commitment to evaluation and reporting. Unless the objectives of the project are clearly articulated, success or failure cannot be determined. Without evaluation, wetland restoration will not contribute to the accumulation of basic knowledge required for restoration ecology to become a predictive science, and despite the occasional apparent success there will also be numerous, perhaps expensive, failures. While there is reluctance from both government agencies and private developers to fund long term monitoring, without adequate documentation of past experience there will also be reluctance to invest heavily in new projects if the prospects of success are uncertain. Chapman and Underwood (2000) provide a critique of past practice and provide guidance for improvement in setting objectives and protocols for the design and conduct of rigorous monitoring programs. What is monitored will need to be relevant to the objectives. If the requirement is to enhance populations of
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particular species (e.g., rare or endangered taxa) then assessment of success must involve monitoring of these species, both at restored sites and elsewhere. Success will need to be determined by reference to the starting position at the restoration sites and also against population trends at regional and larger scales. If a species is in general decline, then establishing a sustainable population may be unlikely. The need to sample external reference sites will add to the cost of monitoring, but is essential for evaluation of restoration. Reference sites need to be ecologically equivalent to the experimental sites (Chapman and Underwood, 1997); comparing upper estuarine sites to those in more seaward positions will not help assessment. If the objectives include restoration of ecosystem functions and processes then not only do the relevant functions and processes need to be defined, but monitoring needs to incorporate direct or indirect measurement of them. Ecosystem process studies are expensive and again require reference sites. In many cases the appropriate methodologies for assessing functions and processes in particular salt marsh types would need to be developed. It is frequently the case that there are a number of objectives, and their assessment will require different sampling designs, both spatially and temporally. There is little background knowledge on the scales of variation in many of the components and processes in salt marshes (Zedler, 1996a) so that pilot studies will frequently be necessary prior to embarking on monitoring. Attempting to include all components within a single sampling design is unlikely to measure any of them adequately (Chapman and Underwood, 2000). While the object of restoration is to create fully functional ecosystems, it is unrealistic to require monitoring of ecosystem processes in every restored site. Rather, what is needed is research to establish reliable surrogates of ecosystem processes in different salt marsh types (surrogates which are appropriate for S. alterniflora marshes will not necessarily be valid for marshes dominated by other genera in different geographic regions). The results of monitoring need to feed back into adaptive management. This may lead to modification of the management techniques through an iterative process, but it may also require reassessment of the objectives. If the objective cannot be achieved because of continuing disturbance, and the disturbance factors cannot be controlled, then this needs to be recognized and new goals set. This is not to say that the objectives should be simplified and generalized to triviality, but rather that restoration should be a realistic proposition. Restoration of salt marshes frequently occurs with the intention of mitigating the damage caused by development or as part of a more general program to repair past damage. In these circumstances objectives are often set in very general or idealistic terms. The objective may be stated simply as the production of a functional (or healthy) salt marsh without “functional” being defined or the type of salt marsh specified. There is a tendency to assume that it is necessary to maximize species and habitat-type diversity, which is justified as increasing biodiversity. However, this is based on a misunderstanding of the concept of biodiversity and its quantification (Adam, 1998); species poor communities are still a component of biodiversity. Zedler (1995, 1996b) has argued that objectives for individual sites need to be set within a regional context and that overall the goal should be “to maintain the natural diversity of species and community types.” Salt marsh vascular plant communities are often species poor (Adam, 1990), although the total species
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richness, including microalgae and bacteria, of the communities is not known. These species poor plant communities may appear monotonous, even “boring,” leading to demand for “more interesting” more species rich communities. Many of the ecological functions for which salt marshes are valued, such as provision of fish nursery habitats or absorption of wave energy are primarily features of the low marsh zone, which support a small number of vascular plants. In some circumstances creation of more diverse communities may be possible, even if the resulting assemblage of species is, within its local context, “unnatural,” although whether these created communities will be self-sustaining in the longer term is uncertain. Zedler (1995) points out that more is not necessarily better, and that in the particular case of Californian salt marshes, monotypic stands of Salicornia bigelovii are important habitat for Belding’s Savannah Sparrows (Passerculus sandwichensis), protected under the US Endangered Species Act. Encouraging the establishment of other species in Salicornia habitat reduces its value to the sparrow and could constitute a breach of the legislation. Those setting objectives must be cognizant of the fact that salt marshes are dynamic ecosystems. At individual sites succession may be occurring, but the present state of a marsh reflects seral changes over the developmental history of the site. Adam (1990) has warned against simplistic space–time substitution interpretations of zonation to predict future succession. Creation of “new” salt marshes by establishment of early seral stages and leaving succession to take its course will not necessarily result in mid and upper marsh communities identical to those previously existing on local sites. Succession can be rapid (Adam, 2000) but at many sites the rate of change appears to be slow; it may be decades before upper marsh develops – probably well after the end of any monitoring programmes. Salt marshes may also be dynamic at larger spatial scales. Adam (2000) documented changes in the distribution and extent of salt marshes in Morecambe Bay in NW England over a quarter of a century. While some marshes were little changed, others had expanded considerably and yet others had undergone extensive erosion as a result of changes in the position of channels (see also Pringle, 1995). Similar dynamic changes to estuarine wetlands have been reported from the Ebro Delta in Spain by Valdemoro et al. (2007). Restoration projects in regions where dynamic changes are likely present particular problems. Concentration of effort at the local, individual marsh scale may be doomed to failure if the site concerned is destined to erode. If a wider view is adopted, for example in the Ebro, the whole delta (Valdemoro et al., 2007), a strategy might be developed whereby the loss of some marshes could be compensated by the formation or creation of new ones. However, while this is conceptually straight forward, implementation would be difficult given the uncertainty of predicting where and when erosion and colonization will occur.
3. PLANNING FOR THE FUTURE Restoration projects are undertaken with the intention of providing for longterm survival of functional ecosystems. Projects should therefore be designed with the future in mind. At the most elementary level, this requires consideration of
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possible developments which may affect the viability of the restoration. The Eve Street Wetland restoration (Stricker, 1995) resulted in a salt marsh which was of heritage significance, an example of what had been, prior to suburbanization, a very extensive community, and which continued to be a significant habitat for migratory wading birds. Despite its conservation value, a few years after the restoration a major motorway was constructed across part of the site. While some of the vegetation has survived, the value of the site as bird habitat has been substantially diminished. Climate and associated environmental changes have affected salt marshes throughout geological history. Global warming after the last Ice Age was accompanied by rapid change in the distribution of both individual species and communities in response to both rising sea level and higher temperatures. Sea level reached its present position roughly 6,000 years ago, although at high latitudes isostatic rebound following deglaciation continues, raising the level of the land relative to the sea and creating new surface for colonization by salt marsh. Salt marshes have obviously accommodated past fluctuations by migration, but the ability to respond to future environmental change is severely compromised by the concentration of the world’s human population in the coastal zone. Urban, industrial, and agriculture development in the coastal zone limits the opportunities for formation of new habitat for salt marsh colonization and the ability of species to migrate. In planning for salt marsh restoration, the viability of sites under the changed conditions predicted to prevail in decades ahead will need to be assessed. While the threat posed by climate change-induced sea-level rise has been recognized for many years (Committee on Engineering Implications of Changes in Relative Mean Sea Level, 1987; Titus, 2000) there is a relative reluctance in many parts of the world from all levels of government to respond in a coordinated and timely manner. Climate change induced increase in the volume of the sea will not translate into a uniform rise in sea level, relative to the land. Relative sea level is influenced by other factors at the regional and local scale. For example, the rise in sea level may be balanced or exceeded by continuing isostasy; regional subsidence, as is occurring in southeast England, might increase the effects of global sea-level rise; human activities such as groundwater or hydrocarbon extraction might result in local land subsidence. Very rapid changes in land level may result from tectonic activity, which has, in recent history, elevated salt marshes above tidal influence in Alaska (Crow, 1971) and lowered forests in Chile into the intertidal so they have been replaced by salt marsh (Valiela, 2006). A rise in relative sea level might be expected to cause loss of the seaward margin of salt marshes; historically any losses might have been mitigated by retreat of the landward margin and establishment of new salt marsh by colonization of what were previously upland communities. The ability in the future of salt marsh to retreat into terrestrial vegetation will be impaired by barriers such as seawalls and urban development. A rise in sea level could also be countered by accretion of the marsh surface, both from inwashed sediment and plant production incorporated into the soil, at a rate greater than the sea-level rise. While knowing where and when earthquakes will occur is not possible, other factors affecting relative sea level are known, or data can be obtained, so that it is
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possible to predict the responses of salt marshes at the local scale (Morris et al., 2002). Likely changes in relative sea level should be assessed before restoration. If the site is likely to be overtaken within decades then the appropriateness of the project needs to be considered. If deficiency in sediment supply is identified as the problem, augmentation of effective sedimentation can be considered (Bakker et al., 2002) through sediment recharge, promotion of sediment accretion through construction of a sedimentation field (as in the Schleswig–Holstein method, where a network of brushwood fences is constructed), or by construction of offshore breakwaters to limit erosion (Nottage and Robertson, 2005). These techniques can also be employed to protect and enhance existing marshes so that loss and degradation does not occur. The methods are relatively expensive and labor intensive, and in the case of sedimentation fields do not offer instant success but may require years of operation and maintenance to achieve the desired result (Nottage and Robertson, 2005). Sediment recharge requires a source of supply, which is often derived from dredging operations. However, the sediment available may not have appropriate physical characteristics for salt marsh development, and, if derived from an industrialized estuary, may be polluted. Extraction may have environmental impacts which would need to be fully assessed before a project could be approved (Nottage and Robertson, 2005). Once deposited at the “new” site the sediment would need to be dewatered, stabilized, and protected from erosion before vegetated marsh could be developed. In many parts of the world, extensive areas of former marsh have been reclaimed to form low-lying agricultural land, protected from the sea by embankments and sluices. With rising relative sea level the expense of maintaining sea defenses may not be justifiable; managed realignment by which existing sea defenses are breached, allows for the creation of new landward defenses and the reestablishment of new salt marsh to replace that lost to erosion and rising sea level outside the original sea walls (Dixon et al., 1998; Nottage and Robinson, 2005; Baldwin et al., 2009). Managed realignment has been practiced in northern Europe (Atkinson et al., 2001, Nottage and Robinson, 2005; Weaver, 2006). The approach of restoring salt marshes after breaching embankments has also been applied on a large scale in San Francisco Bay (Williams and Faber, 2001, 2004; Williams and Orr, 2002), although in these cases the primary object was to regenerate salt marsh on former salt production ponds or agriculture lands rather than to mitigate against the effects of rising sea level. Changing sea level is only one of the consequences of an increased greenhouse effect which could impact on salt marsh restoration projects. Salt marsh plant species in general have wide geographical distributions. Nevertheless variation in floristic composition with latitude is observable (Adam, 1990) and temperature is likely to be the major determinant. Global warming is likely to result in changes to species distributions, although different species are likely to respond at different rates so that assemblages of species (communities) will be resorted. If restoration projects involve planting, should the object be to plant what are the current regional assemblages or should future change be anticipated so that the species planted are sourced from slightly lower latitudes? Even if such bold experiments are not conducted the provenance of planting material would still need to be
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considered. Widespread salt marsh species are often made up of numerous genotypes (Adam, 1990; Zedler and Adam, 2002) and genetic variation is expressed at a range of scales both within and between sites. In restoration projects, use of local provenance material in plantings is often encouraged (Zedler and Adam, 2002), but does climate change require a relaxation of the rules? If environmental conditions are likely to change, should genotypes more adapted to predicted future environments be planted now, rather than geographically local forms which were selected for under previous environmental conditions? There are no definitive answers to these questions, but it will be important to treat every project as an experiment, with full documentation of all stages. In terms of planting material it will be necessary to collect and propagate from known provenances, rather than using “off the shelf” stock of unknown origin from commercial nurseries. While the majority of salt marsh plant species have wide distributions there is a small number of geographically and ecologically restricted taxa. These “rare” species may be particularly vulnerable to climate change. Should these species be given particular attention in restoration projects and be planted outside their current geographical range to anticipate change? The majority of botanists and conservation authorities would probably urge caution, and while supporting ex situ conservation, would not support planting in the wild outside existing distributions. Such a position would be strengthened by the experience of species which are rare in their natural distribution but which have elsewhere become major weeds (e.g., Juncus acutus). Nevertheless, Ranwell (1981) argued that there was a need to give more consideration to introduction of species, particularly rare species, outside their current natural range. With appropriate checks and monitoring, restoration sites (but not “pristine” natural sites) could provide an opportunity for experimentation with rare species. Although temperature is the component of future climate change which has been given the most attention, and for which it is believed the more reliable predictions are available, a number of other aspects of climate are also likely to change. These include the frequency and intensity of extreme climatic events and rainfall seasonal distribution, intensity, and quantity. These changes will have effects on salt marshes – changed seasonality and quantity of rainfall coupled with temperature increases could result in changes in soil salinity patterns, particularly in the upper salt marsh zones. There is too much uncertainty associated with predictions of these parameters for them to be assumed as definitive in planning restoration, but the likelihood of change further strengthens the need for long-term monitoring and adaptive management strategies. Changed climatic conditions may also affect fire regimes. Communities dominated by Phragmites or Juncus spp. can burn fiercely. Although the likelihood of fire affecting natural or restored salt marshes will remain small it would be precautionary to build capacity to respond to fires into the management regimes of those restored marshes with flammable vegetation. The Eve Street Wetland (Stricker, 1995) is one of a chain of wetlands in Sydney’s suburbs. Following concerns expressed by neighboring residents that their properties were at risk from fire in the wetland vegetation, a local conservation group (the Rockdale Wetlands Preservation Society) negotiated an action plan with the fire brigade so that any fire would be fought from the perimeter with minimal disturbance to the wetlands (Rayner, personal communication).
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The underlying driver of anthropogenic climate change is increased atmospheric concentrations of greenhouse gases, particularly carbon dioxide. Changes in carbon dioxide concentration will have effects on vegetation independent of any climatic effects. Increased carbon dioxide will differentially favor species utilizing the C3 photosynthetic pathway over C4 plants. The consequences of this are potentially greater for salt marshes than most other ecosystems. Salt marsh vegetation is often made up of mixtures of C3 and C4 species, so that as the carbon dioxide concentration increases there could be a shift to increased abundance of C3 species. Field experiments involving increasing carbon dioxide in chambers in brackish marshes have shown increased productivity in C3 Cyperaceae but not in C4 Spartina (Arp and Drake, 1991; Drake, 1992; Drake et al., 1996), and an increase in abundance of the C3 species over C4 (Arp et al., 1993). In both C3 and C4 species increased carbon dioxide resulted in decreased transpiration (Drake, 1992), which could limit increases in summer soil salinity. Under increased carbon dioxide C3 plants are likely to have lower nitrogen contents, this could result in lower damage by pathogens (Thompson and Drake, 1994), but also lower attractiveness to herbivores which could result in complex flow-on effects through the food chain. The strong probability that changes in carbon dioxide concentration will result in changes in community species composition suggests that setting objectives for restoration to exactly replicate current communities will be unrealistic.
4. ADDRESSING C AUSES AND NOT SYMPTOMS Many of the more immediate causes of salt marsh degradation are themselves symptoms of wider problems. Restoration which focuses on ameliorating symptoms without addressing the underlying cause is unlikely to be effective in the long term. Kelleway (2005) showed that many salt marshes in southern Sydney (Australia) had been extensively damaged by unauthorized vehicle access, a phenomenon repeated in many other urban areas. Restoration of this damage could require ripping of heavily compacted areas, filling of eroded wheel tracks, reinstatement of original drainage patterns, and replanting. However, such efforts would be to no avail if vehicle access still occurs. Blocking of access will be essential but not sufficient. Barriers and fences can soon be destroyed or circumvented. Enforcement, education, and possibly provision of alternate recreational facilities will be required if recurrence of damage to salt marshes is to be prevented. Some forms of degradation can be addressed at the site-specific scale – access by vehicles, changes due to too high or too low grazing pressure, changed fire regimes, restoration of tidal flushing and drainage patterns, some forms of pollution, some weed problems – but in many cases action at a much larger spatial scale will be required. If degradation is due to agricultural runoff the problem would need to be addressed at the catchment scale, for example, the reduction of phosphorus inputs into the Peel–Harvey estuary in Western Australia (Brearley, 2005). Some
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problems will require action at continental scales. In both Eurasia and North America there have been substantial increases in the populations of migratory geese in recent decades. A major factor in the increase has been greater winter food availability in temperate latitudes because of changes in agricultural practices, combined with significant reduction in hunting pressure. The increased grazing pressure and grubbing of rhizomes has caused disturbance to temperate marshes in winter (Smith and Odum, 1981) but the most extensive damage has been to breeding grounds on Arctic salt marshes, where loss of vegetation and increased soil salinity will be difficult to counter (Jefferies, 1997). Management of the goose populations, which will be an essential prerequisite for any rehabilitation of the marshes, will require international cooperation. If successful rehabilitation will, in most cases, require actions which go beyond the site specific, are single site projects a waste of resources? Zedler (1995, 1996b) has argued that, ideally, restoration should be undertaken in the context of at least regional objectives and strategies, but it is likely that most projects will continue to be site based and often ad hoc. Many projects are initiated, and strongly supported, by the local community. Failure will not only cause disappointment but it may jeopardize future support for conservation management. It is extremely important that there is community endorsement of objectives and strategy, and recognition that some outcomes are only temporary, and that further remediation to address the same problems will be required. If the need for continuing maintenance is stressed at the outset, it is possible that some projects would not be attempted, but it is likely that often the community will still support rehabilitation of sites in circumstances where the ultimate cause of degradation cannot be addressed but where treatment of symptoms on a recurrent basis will still lead to maintenance of desired features and ecological processes.
5. MANAGING D ISTURBANCE There are many factors which could be managed or modified in the restoration projects. Every site is unique, but there are a number of issues which arise frequently at restoration sites, and these are discussed below.
5.1. Restoring hydrology Salt marshes are often topographically complex with numerous creeks and pans. The topography creates habitat diversity which is reflected in the distribution of individual species and communities (Adam, 1990). The creeks provide the major pathway for exchange of nutrients, detritus, and biota between marshes and estuaries and coastal seas. There is considerable variation in the nature of creek and pan systems between marshes, which can be partly related to substrate characteristics, tidal amplitude, and vegetation (Chapman, 1974). In temperate Australia creeks and pans may be virtually absent from some large salt marshes (Adam, 1997). The basic distribution patterns of creeks and pans are laid down from the earliest
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stages of marsh colonization, and altered by both erosion and infilling during the course of marsh development. In some mature marshes the creek system may be thousands of years old. In existing, but degraded, marshes, parts of the creek system may have been obstructed through construction of causeways and culverts, canalized to improve discharge of run off from the catchments, or augmented by additional ditches to improve drainage. Given the importance of hydrology to the nature, development, and function of salt marshes, high priority is given to hydrological issues in restoration (Callaway, 2001). Where new wetlands are created it will be necessary to design and construct drainage channels. The process of creation compresses the establishment of processes which in a naturally developed marsh may have taken centuries (Callaway, 2001) and questions arise as to whether the desired complexity can be created instantly or whether a framework should be established with adjustment by natural processes to take place over a period of years. Callaway (2001) reviewed a large number of restoration projects and identified a number of common problems including low density of creeks, lack of small firstorder creeks, generally straight creeks lacking sinuosity, creek bank instability, lack of steep bank slopes, and failure to replicate the natural topography of the creek bank–marsh interface. Many of these problems arise because of the difficulty of constructing replicates of natural geomorphology with bulldozers and excavators. Further issues arise if the marsh being restored has preexisting fill or where fill is required to bring it to the level needed to support the required community type. If the fill has different physical properties from that found at local reference sites the ability to create creek banks of the required steepness and stability may be impaired, while the chemical properties of fill will influence the development of both vegetation and faunal communities (reviewed by Callaway, 2001). The option of specifying the nature of fill may not be open; often marsh restoration proceeds on the basis of using what is already in place, or moving fill from a predetermined site elsewhere. Modification of hydrology is one of the major components of restoration which can take place within the confines of a single site, but issues of freshwater input, either from groundwater or as surface flow, will require consideration of the broader catchment.
5.2. Managing weeds The spread of weeds in all habitats is one of the major threats to biodiversity (Mooney and Hobbs, 2000). Salt marshes occupy an environment which is inimical to many plants, and the number of species which can tolerate various combinations of salinity and water logging is relatively small. Additionally, many salt marsh species have very wide geographical distributions (Adam, 1990) and potential vectors for long distance transport of propagules (currents, tides, and migratory birds) have operated over a far longer time period than humans have been transporting species around the globe. It might be expected that most salt marsh species will already have spread to the tolerance limits of their range and that the incidence of more recent exotic invaders would be low. However, there are a number of
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examples of invasive introduced species, although the naturally wide distribution of many salt marsh species can give rise to uncertainty as to the status of some (e.g., is Spartina maritima native in southern Africa? – Pierce, 1982, on which continent is Lampranthus tegens native? – Adam, 1996), but in most cases the time of introduction and subsequent pattern of spread are reasonably well known. Removal of introduced species may be part of the restoration of salt marshes but can be an expensive and difficult operation. While there is often a knee-jerk response by managers to the presence of introduced species which assumes that control is essential, this needs to be questioned before expensive programs are initiated. Firstly, what is the evidence that the introduced species are responsible for “harm”? Introduced invasive weeds could have a variety of impacts including; competitive exclusion of native species, changes to the structure and composition of both floral and faunal communities, alteration of physical and chemical characteristics of the environment (such as modification to drainage characteristics, sedimentation rates, or soil properties), or changes in genetic structure of native populations through hybridization (Spartina townsendii, and subsequently Spartina anglica, arose in southern England following spontaneous hybridization between the European S. maritima and the introduced North American S. alterniflora, hybridization between introduced S. alterniflora and Spartina foliosa in California is regarded as a threat to the survival of the native species – Daehler and Strong, 1994) and alteration to ecosystem processes. These impacts will have affects on biodiversity, although evidence that these extend to hastening extinction of any species is inconclusive (Ranwell, 1981; Valiela, 2006). Yet there is at least circumstantial evidence for some local extinction of species, and loss in components of biodiversity may render other species more vulnerable to total extinction. In addition to impacts on existing marshes, Spartina species may colonize mudflats below the previous lower limit of salt marsh reducing the extent of mudflat habitat (Goss-Custard and Moser, 1988; Adam, 1990; Ayres et al., 2004; Valiela, 2006). The second issue which needs to be addressed is whether control is likely to succeed. Control will be unsuccessful unless the factors promoting weed invasion are also addressed. These may include alterations or disturbance to the environment (e.g., stormwater discharge promoting growth of Typha or Phragmites – Zedler et al., 1990) or abundant regional supply of new propagules. There may not be any practical measures for the control of many species because there are considerable constraints on the use of chemical controls, physical removal is labor intensive and can be onerous in soft mud, and the research and development necessary to obtain regulatory approval for biological control is time consuming, expensive, and not guaranteed of success. Invasive species are often closely related to species indigenous to the marshes being invaded, so that biological control, which requires that the control agent attack only the invasive species, may not be appropriate. Nevertheless, the potential for use of biological methods for the control of S. alterniflora on the west coast of the United States (Grevstad et al., 2003) and for Phragmites australis (Blossey, 2003) has been suggested. Some of the issues created by invasive species may, in the longer term, be self correcting. The original dense monocultures of S. anglica will, during the course
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of succession, become more species rich (Ranwell, 1981), sometimes relatively quickly (Adam, 2000). The outcome may not be similar in composition to preinvasion communities, but returning to the past may be an impossible dream. In other cases of invasion such as by J. acutus or P. australis there is little indication so far that there will be a decline in dominance and increase in species diversity. Invasion by weeds undoubtedly has impacts on some components of biodiversity but in many cases there are insufficient data to determine the effects on ecosystem processes and services. Does weed invasion change the productivity of salt marshes; what are the impacts of this on adjacent estuarine and coastal water ecosystems? How important is plant species composition and vegetation structure for the maintenance of the faunal community? While there is a very strong argument for being vigilant to detect and control new weeds and new infestations before they become established, reduction of major long-standing populations of weeds may not be possible, necessitating changes to the objectives of restoration projects. There are a large number of weeds reported from salt marshes, but many are currently relatively local in their impact. The species of most concern, and the subject of active management, are tall growing competitive dominants, capable of spread by clonal growth. These species are, or are assumed to be, ecosystem engineers. Species in this category include Spartina spp., P. australis, and J. acutus. Spartina spp. differ from most other salt marsh invaders in that they were deliberately planted at many locations around the world (Ranwell, 1967; Phillips, 1975; Boston, 1981) and have subsequently spread from the point of introduction. Spartina planting is still promoted in China (Chung et al., 2004) despite being on the official list of harmful invasive species (Wang et al., 2006). P. australis, the common weed is an almost cosmopolitan species in fresh and brackish wetland. Change in environmental conditions (such as grazing, drainage or nutrient regimes) can promote the spread of Phragmites (Esselink et al., 2002; Silliman and Bertness, 2004), but the extensive spread in the coastal marshes in the United States in recent decades, has been of an European genotype acting as a cryptic alien (Saltonstall, 2002). J. acutus, a species of conservation value in its native northern hemisphere habitats, has become a very aggressive invader in south east Australia. A range of techniques have been employed as control measures, these are often expensive and labor intensive. They include physical removal (DPIWE, 2002; Lacambra et al., 2004; Paul and Young, 2006) and herbicide use (Truscott, 1984; DPIWE, 2002; Lacambra et al., 2004), although the effects of herbicide or other estuarine flora and fauna are poorly documented. As well as large, obviously aggressive invaders there are many other nonindigenous species recorded from salt marshes. Some of these are currently known from only a few sites, although amongst these may be sleepers – species which after a latency period can become aggressive invaders. Others may be widespread and abundant, but because they are of relatively small stature there has been a perception that they do not constitute a problem. Introduced species are particularly abundant in marshes in Mediterranean climatic zones (Bridgewater and Kaeshagen, 1979; Sullivan and Noe, 2001),
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particularly in disturbed areas (such as footpaths) and in the high marsh transition to supralittoral vegetation. This latter zone is where the majority of rare or geographically restricted salt marsh plants occur (Ranwell, 1981). One such species is the endangered hemiparasite Cordylanthus maritimus ssp. maritimus in upper salt marshes in California. Amongst the invasive alien species in Cordylanthus habitat is the annual European grass Parapholis incurva. Fellows and Zedler (2005) have shown that Parapholis in this circumstance acts as a “pseudo host” in that haustoria formation occurs between the two species, but flowering of Cordylanthus is substantially reduced (average one flower per plant) compared with the native host, the perennial grass Distichlis spicata (average 13 flowers per plant). This reduction, if it results in similarly reduced seed production, could over time lead to population decline. This finding warns us of the possibility of subtle adverse consequences of any introduced species on salt marsh. Another annual grass of European Mediterranean origin which has become established in both California and Australian salt marshes is Polypogon monspeliensis. Callaway and Zedler (1998) showed that increased cover of Polypogon compared with that of the native succulent perennial chenopod Sarcocornia perennis was favored by decreased soil salinity and increased freshwater inputs. In southern California many salt marshes experience increased freshwater inputs from either agricultural and urban runoff (Callaway and Zedler, 1998) and the impacts are often exacerbated by anthropogenic reductions in tidal exchange (Boland and Zedler, 1996). Spread of annual grasses at the expense of evergreen perennials could alter the vegetation structure and food availability of the salt marsh, with consequent affects up the food chain (Callaway and Zedler, 1998). Kuhn and Zedler (1997) showed that application of salt was an effective means of eradicating small patches of annual grasses but while this might be an appropriate control technique in the early stages of invasion it is unlikely to be practical at large scales with a well established population (application of salt, in conjunction with other control techniques, was also shown by Paul and Young (2006) to be effective against J. acutus). Restoration of tidal access and reduction in freshwater run off entering marshes would be likely to control not just Polypogon but a whole suite of other species with similar ecological characteristics. We should take all possible measures to prevent the further spread of introduced species, but the control and eradication of existing infestations of many species as part of rehabilitation is likely to be difficult if not impossible.
5.3. Introduced fauna Ports and estuaries are hot spots for introduced fauna (Valiela, 2006) but relatively little is known about introduced fauna on salt marshes. The Australian burrowing isopod Sphaeroma quoyanum is an invasive species in Californian salt marshes (Talley et al., 2001), where it has a serious impact by increasing the erosion of creek banks weakened by its burrows. There is no control method, but the potential for invasion of new sites may be a constraint for future restoration projects. In 1998 the mosquito Ochlerotatus camptorhynchus was discovered in New Zealand and is presumed to be an introduction. As the species is a potential vector
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of the Ross River fever virus the discovery prompted an intensive control program using an insect growth regulator and insecticide (Stark, 2005a,b). The damage to nontarget organisms is believed to have been slight (Stark, 2005b), but the event highlights the risk to human health posed by mosquitoes. The distribution of mosquito species, and the incidence of a number of diseases are expected to increase as a result of global warming. Historically, attempts to control salt marsh mosquitoes through drainage and insecticide use have been an important factor in marsh degradation. However, increased community concern about possible health risks may become a constraint on the initiation of restoration projects, and greater scrutiny of projects will occur to ensure minimization of breeding habitat through appropriate drainage. 5.3.1. Vertebrates Waterfowl and mammals are significant herbivores in salt marshes. While provision of habitat may be one of the long-term objectives of restoration, it may be necessary to limit their access in early stages of development to allow vegetation to become established. Timing of restoration activities will also need to take into account the requirements of fauna. For activities which may cause disturbance to birds such as sediment recharge or construction of banks and drainage systems there may be only limited windows of opportunity (taking migratory patterns and breeding seasons into consideration) when disturbance will be minimal and conditions are appropriate for establishment of vegetation. Considerable lead time may be required to ensure availability of machinery and other resources during these critical intervals.
5.4. Grazing In their aboriginal state salt marshes would be grazed by a diversity of vertebrate fauna including (at least seasonally) waterfowl as well as mammals both large (e.g., deer and kangaroos) and small (e.g., voles and hares). This pattern of grazing still continues in many regions, but at many sites natural grazing has been supplemented, or replaced, by grazing of domestic livestock. Agricultural exploitation of salt marshes has also included hay making in both Europe and North America (Valiela, 2006), although currently it is a very minor use. Both livestock grazing and haymaking result in export of productivity to the hinterland, against the natural flow into estuaries. Agricultural grazing is still an important widespread use of salt marshes in northern Europe and South America but is more local on other continents. Grazing can affect the composition of vegetation through adverse impacts on sensitive species, and the consequential promotion of grazing tolerant species. Additionally, trampling by livestock may result in loss of vegetation cover and modification of local drainage patterns. The effects of grazing have been extensively studied in Europe where grazed salt marshes differ considerably in species composition, structure and vegetation type from ungrazed marshes (Adam, 1978, 1981; Bakker, 1989). Changes in grazing pressure can result in rapid changes to the composition of salt marsh vegetation
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(Adam, 2000). An assessment of the condition of salt marshes in Britain and Northern Ireland designated as being of high conservation value (Williams, 2006) concluded that a number were threatened by either overgrazing or undergrazing. The question therefore arises as to whether restoration of salt marsh might, in some circumstances, involve reestablishment of an appropriate grazing regime. Whether grazing is a desirable management strategy will depend initially on the local context. For very small sites, agricultural grazing would be neither practical nor economically viable. In other cases changes in local agricultural practice, such as from mixed farming to cropping, may mean that there are no adjacent farmers able to take up salt marsh grazing; if grazing is thought desirable then conservation agencies would have to be responsible for management of livestock. If grazing could be practically and ecologically viable, under what circumstances might it be appropriate? The answer will depend both on the desired end condition of a site, and its prior history. If the site has never been grazed then introduction of grazing will result in a change in species composition through loss of sensitive species, although total species diversity might be maximized under moderate grazing pressure (Bakker, 1989). If the site has been grazed, relaxation or cessation of grazing pressure is unlikely to result in a reversion to anything resembling a never grazed state. Rather than reestablishment of sensitive species, the most likely outcome would be a loss of species diversity and the development of rank species poor plant communities (Adam, 1978, 1981, 1990, 2000; Bakker, 1989) dominated in Northern Europe for example by Festuca rubra or Elytrigia atherica. If grazing pressure has only recently been reduced then reintroduction of the former grazing may prevent the reduction in diversity but once species poor stands have developed, grazing alone is unlikely to restore the former vegetation. Manipulation of grazing pressure is one of the tools available to conservation managers to alter vegetation composition and structure. Whether and how this tool should be used at particular sites will require a wider regional perspective to maximize conservation outcomes. If the objective, for example, is to maintain the maximum diversity of salt and brackish marsh birds then sufficiently large (so as to maintain viable populations) stands of vegetation with different structural characteristics will need to be maintained, restored or created over a range of sites. Changes comparable to those shown in response to grazing in northern Europe have not been reported from elsewhere. Although marshes on the Atlantic coast of North America were grazed and used for hay making from early in European settlement (Teal and Teal, 1969; Valiela, 2006), the overwhelming dominance of Spartina spp. may be the reason for the apparent lack of impact from grazing. Outside Europe, grazing is unlikely to come into consideration as a tool in restoration.
5.5. Pollution Many salt marshes have been degraded by pollution. Pollution may be chronic or episodic, local, regional, or global, and persistent or ephemeral. In setting objectives for restoration, consideration needs to be given to the impacts of
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pollution on the ecosystem and whether they can be addressed at a scale appropriate to the restoration project. Site-specific pollution includes point source discharges and spills. In many urban salt marshes stormwater discharge has a major impact. The lowering of soil salinity by freshwater inputs can facilitate the invasion of tall growing fresh and brackish water marsh plants which can form dense, often monospecific, stands out competing the original halophytic vegetation (Zedler et al., 1990). Diversion of flow will permit the normal marsh salinity regime to reestablish, but restoration of the vegetation may require manual removal of the invaders and replanting. Urban stormwater introduces other pollutants into marshes; petrochemicals and heavy metals from road surfaces, nutrients, sediments, weed propagules, and rubbish such as plastic and paper. In new urban development the drainage system should be designed either not to discharge into wetlands, or, if this is not possible, to be treated to an appropriate standard before discharge. Retrofitting existing drainage systems is, however, expensive and may not be practically possible as, for example, space for installation of features such as gross pollutant traps may not be available. In industralized estuaries salt marsh sediments may have been long-term sinks for heavy metals; high levels of metals have been reported from both sediments and biota (Beeftink and Nieuwenhuize, 1986; Chenhall et al., 1992; Ohmsen et al., 1995). Although concerns have been expressed about the possible transfer of metals up the food chain (including to humans – Beeftink et al., 1982), and adverse physiological impacts on test species are known from toxicological studies, there have been few studies of the effects of pollution on ecosystem processes in salt marshes. Valiela (2006) reviewed available literature and suggested that despite the clear experimental evidence of physiological effects there was little discernible evidence for population and community level impacts. Chemical loads within organisms, and species composition of assemblages (for example bacterial assemblages differ in composition in polluted salt marsh sediment – Cao et al., 2006) can be used to monitor pollution, but do not necessarily provide surrogate measures of impacts on functions or processes. Nevertheless, a precautionary approach would suggest that pollution discharges should be reduced and if possible eliminated when planning restoration, but there is little that can be done to address existing pollution loads. Excavation of polluted sediment could mobilize contaminants back into the estuary and create a problem of disposal of contaminated sediment. For these reasons environmental protection authorities would be unlikely to grant approval for restoration projects which involve disturbance to known contaminated sites. Even if approval were granted there would be a need to source the same quantity of clean fill with appropriate physico-chemical properties. The fact that a site is contaminated with heavy metals does not rule it out from being restored, but there would be a need to prevent disturbance of the sediment and possibly long-term restrictions on access and usage, to prevent, for example, human consumption of Salicornia spp. Oil pollution is an ever present threat in the marine environment, with large numbers of spills, large and small, being recorded every year (European Environment Agency, 2006). When spills impact on the coast there is major public and political concern, and pressure for cleanup to occur. If salt marshes are oiled should
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there be active rehabilitation or should natural recovery be allowed to take its course? A number of major oil spills have affected salt marshes in different parts of the world, and a review of impacts and treatment methods is provided by Baker et al. (1994). A number of techniques can be used to treat surface deposits of oil, but remediating subsurface oil is more difficult. Methods include low pressure water flushing, although care has to be taken not to cause erosion; rapid deployment of sorbents could reduce penetration into the substrate but collection and disposal of oiled sorbent material may be a problem. Burning of oiled vegetation may be an option in some circumstances, but burning may increase the hydrocarbon content in the underlying sediment. Cutting of oiled vegetation, particularly if leaving oil might be a threat to wildlife, is an option but the logistics of operating in soft sediment and of removing and disposing of oiled material would need to be considered. The most extreme treatment approach is combined vegetation and sediment removal. After stripping of sediment transplanting and/or seeding is necessary to reestablish habitat and prevent erosion. The approach has been successful in smallscale trials but serious problems occurred when it was used on a large scale following the “Amoco Cadiz” spill in 1978. This spill deposited large amounts of oil on salt marshes in Brittany (France). Heavy machinery was used to remove as much as 50 cm of sediment and marsh creeks were widened and straightened. Although vegetation of sites not subject to treatment recovered within a decade, large parts of the treated marshes remained unvegetated in 1990, primarily because the sediment removal had created a surface too low in the tidal range for marsh establishment (Baker et al., 1994). The immediate impact of an oil spill on salt marshes is dramatic with smothered vegetation and dead or dying fauna very visible. Once the surface oil has weathered or been removed the marsh surface may appear “normal” to casual observation, but very little is known about the long-term effects of exposure to residual hydrocarbons. Oil residues can be detected in sediments long after pollution incidents but impacts on the ecosystems from these residues have been harder to establish. In 1969 the barge “Florida” ran aground in Buzzards Bay, Massachusetts (USA) and spilled oil spread over adjacent salt marshes. Nearly 40 years later the salt marsh vegetation at the spill site appears to be the same as that at nearby marshes which were not affected by the spill, but high concentrations of hydrocarbons can be detected in the anoxic underlying sediment (Culbertson et al., 2007). One of the major faunal species on the marsh, the fiddler crab Uca pugnax, burrows into the sediment. Comparison between the behavior of U. pugnax at the contaminated and control sites have shown behavioral differences where crabs at the oiled site avoided burrowing into the contaminated layers, had delayed escape responses, lowered feeding rate and occurred at lower densities (Culbertson et al., 2007). The oil is thus still biologically active and despite superficial appearance, marsh recovery has not occurred. These results highlight the importance of initial prevention of oiling, but in the context of restoration emphasize the importance of choice of matters to include in monitoring programs. Study of only the vegetation could well have supported on erroneous conclusion that the marsh had fully recovered.
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The ecological importance of salt marsh has been recognized in national oil spill contingency plans (see, e.g., Carter, 1994) and priority is given to preventing oil reaching marsh surfaces. In designing restoration projects for sites where there is a high probability of oil spills (such as near ports or refineries) it may be appropriate to install permanent anchoring points for the rapid deployment of exclusion booms in the event of spills. Although shipping accidents are the source of major oil spills, oil and other chemicals may be spilt into creeks and drains following road accidents so that contingency plans to protect salt marshes (both restored and natural) from spills would also need to consider drainage from the terrestrial catchment. One component of salt marsh biota, which is vital for the restoration of ecosystem functions, is the microalgal film on the sediment surface. These films play a major role in stabilizing the surface, permitting higher plant colonization, and preventing erosion (Coles, 1979; Mason et al., 2003), are important food resources for a diversity of fauna, and play a considerable role in nutrient fluxes including through nitrogen fixation. Formation of microalgal films in restored marshes has been demonstrated by Underwood (1997) and Janousek et al. (2007). However, Mason et al. (2003) have shown that sublethal concentrations of agricultural triazine herbicides (such as simazine and atrazine) adversely affect growth and photosynthetic efficiency in diatoms. Given the very widespread use of these chemicals, and their presence in runoff into estuaries, they may be affecting colonization success in marsh restoration more widely then has been appreciated.
6. C ONFLICTING P RIORITIES Given the high conservation values ascribed to salt marshes, restoration and re-creation of degraded salt marshes would be expected to have high priority. Nevertheless, there are circumstances where former salt marsh has developed into habitats of high conservation value in their own right. In northern Europe embankments and the complex of grazing marshes and brackish ditches behind them may in some instances be hundreds of years old. Embankments may support rich floras, include a number of rare species (Gray, 1977; Ranwell, 1981). Natural unimpounded brackish marshes are now very limited in occurrence, and the grazing marshes and their ditches are now the major stronghold for a number of species and communities (Gray, 1977; Doody, 2001). The habitat values of the grazing marshes are themselves now threatened by changes in agricultural practices and policies, and are afforded protected status. Conflicts may arise if proposals for restoration or managed realignment involve loss of grazing marsh (Pethick, 2002). The regulatory and environmental assessment regime required for salt marsh restoration project approval may be complex (Nottage and Robertson, 2005). But ultimately decisions as to whether salt marsh restoration prevails over conservation of some other habitat will be made on the basis of policies, which while they may be informed by science, are made within political fora. There is no absolute criterion for deciding that salt marsh is “better” and thus should be favored over conservation of grazing marsh.
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However policy and legislation may provide a basis for addressing issues consistently rather than an in erratic ad hoc manner. While supporters of salt marsh may press the preeminence of salt marsh restoration (Pethick, 2002) circumstances may arise when protection of other habitat may be given priority. The invasion of salt marshes in southern Australia by mangroves raises similar dilemmas (Saintilan and Williams, 1999; see Saintilan et al., 2009). In the state of New South Wales salt marsh is legally recognized as an endangered ecological community, but mangroves are afforded special protection because of their contribution to the maintenance of fisheries. If mangroves invade a restored wetland should this be regarded as an inevitable ecological change or should active removal of mangrove seedlings be undertaken, even though this would require long term commitments? Under a number of jurisdictions around the world, salt marsh is protected under either ecological community or habitat protection legislation. Additionally, individual species may be formally recognized as threatened. Legislation may set constraints on the determination of objectives for particular restoration projects in that actions which might potentially have adverse impacts on particular species or communities may be forbidden, or there may be additional approval, monitoring and reporting requirements. Zedler (1995) discusses how, in California, favoring mixed communities over dominance of Salicornia bigelovii might adversely affect Belding’s Savannah Sparrow).
7. D ISCUSSION The global trend to increasing urbanization [for the first time in human history more than half of the global population now lives in cities, and the shift from rural to urban living continues (United Nations, 2004)] leading to further expansion of the world’s major coastal cities, and the related need to further develop transport and industrial facilities on estuaries, will inevitably place pressure on intertidal wetlands continuing the long term history of coastal wetland being a major focus for development activities (Bromberg and Bertness, 2005). Despite increasing recognition of the importance of salt marsh there will continue to be losses of sites where the social and economic benefits of development are deemed to outweigh those of conservation. However, in return for development approval it is increasingly likely that obligations for mitigation, either in the form of restoration of degraded salt marsh or the creation of new habitat, will be imposed. The overall intent will be expressed as no net loss, but whether this will be achievable either in terms of area, or more importantly, function remains uncertain. To give greater certainty to future restoration proposals, there will need to be more rigor in setting objectives and in monitoring and evaluating the performance of projects against the objectives. As well as providing for mitigation of past and planned damage to natural salt marshes, restoration, and re-creation projects present other opportunities. They are appropriate venues for public education campaigns, through events, boardwalks, observation sites, and interpretative signs. These can be designed into projects, so
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their provision should not entail the disturbance which might result from attempting to provide facilities at natural sites. Additionally, many restoration sites will be in, or close to, urban areas, thereby increasing accessibility to the public. Restoration projects themselves need to be treated as experiments, but they provide additional opportunities for manipulative experimentation, which as natural marsh is given increased protection under species and habitat conservation legislation will become difficult to conduct elsewhere. Not every restoration project will become an experimental research site, but it is an opportune time to identify a range of sites in different geographic regions and with different marsh types which can form a network for long term ecological research. This has already occurred with sites in southern California (Zedler, 2001) but in many parts of the world there is little interaction between ecological researchers and habitat restoration projects (Chapman and Underwood, 2000), and the opportunities for research are ignored and lost.
8. C ONCLUSIONS Large numbers of salt marsh restoration projections have been conducted, although the majority have been in the United States, and have been applied to a relatively restricted range of the world’s salt marsh types. Although there is a substantial literature on salt marsh restoration many projects have not been assessed (Ambrose, 2000) and for those that have been described, inappropriate setting of objectives and conduct of monitoring limits our ability to draw general conclusions (Chapman and Underwood, 2000). If the challenge of meeting a goal of no net loss of salt marsh is to be met then restoration and re-creation will have to play a major part, but for success there needs to be a much firmer scientific framework. The task will be made more complicated by the uncertainties associated with global climate change and sea level rise. Nevertheless, many of the factors which have caused salt marsh degradation can be identified, and addressing these will be a major component of restoration. However, in some cases, such as the invasion of some exotic species there is little prospect of effective control, necessitating a recasting of objectives for restoration. Re-creating the past will often not be an appropriate option, but restoring ecological functions and habitat for particular species in self sustaining salt marshes may still be possible. Salt marsh restoration affords unrivalled opportunities for experimental manipulation, adding to our fundamental knowledge of salt marsh ecosystems and permitting improvement in restoration techniques.
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Nottage, A., Robertson, P., 2005. The salt marsh creation handbook: a project manager’s guide to the creation of salt marsh and intertidal mudflat. The RSPB, Sandy and CIWEM, London, 128pp. Ohmsen, G.S., Chenhall, B.E., Jones, B.G., 1995. Trace metal distributions in two salt marsh substrates, Illawara region, New South Wales, Australia. Wetlands (Australia) 14, 19–31. Paul, S., Young, R., 2006. Experimental control of exotic spiny rush, Juncus acutus, from Sydney Olympic Park: I. Juncus mortality and re-growth. Wetlands (Australia) 23 (2), 1–13. Pethick, J., 2002. Estuarine and tidal wetland restoration in the United Kingdom: policy versus practice. Restor. Ecol. 10, 431–437. Phillips, A.W., 1975. The establishment of Spartina in the Tamar estuary, Tasmania. Pap. Proc. R. Soc. Tasman. 109, 66–76. Pierce, S.M., 1982. What is Spartina doing in our estuaries? S. Afr. J. Sci. 78, 229–230. Piehler, M.F., Currin, C.A., Casanova, R., Paerl, H.W., 1998. Development and N2-fixing activity of the benthic microalgal community in transplanted Spartina alterniflora marshes in North Carolina. Restor. Ecol. 6 (3), 290–296. Pringle, A.W., 1995. Erosion of a cyclic salt marsh in Morecambe Bay, north-west England. Earth Surf. Process. Landf. 20, 387–405. Ranwell, D.S., 1967. World resources of Spartina townsendii (sensu lato) and economic use of Spartina marshland. J. Appl. Ecol. 4, 239–256. Ranwell, D.S., 1981. Introduced coastal plants and rare species in Britain. In: Synge, H. (Ed.), The Biological Aspects of Rare Plant Conservation. Wiley, Chichester, pp. 413–420. Saintilan, N., Rogers, K., McKee, K., 2009. Salt marsh–mangrove interactions in Australasia and the Americas. In: Perillo, G.M.E., Wolanski, E., Cahoon, D.R., Brinson, M.M. (Eds.), Coastal Wetlands: An Integrated Ecosystem Approach. Elsevier Science, Amsterdam, pp. 855–884. Saintilan, N., Williams, R.J., 1999. Mangrove transgression into salt marsh environments in South East Australia. Global Ecol. Biogeogr. Lett. 8, 117–124. Saltonstall, K., 2002. Cryptic invasion by a non-native genotype of the common reed, Phragmites australis, in North America. Proc. Natl. Acad. Sci. 99, 2445–2449. Silliman, B.R., Bertness, M.D., 2004. Shoreline development drives invasion of Phragmites australis and the loss of New England salt marsh plant diversity. Conserv. Biol. 18, 1424–1434. Smith, T.J., Odum, W.E., 1981. The effects of grazing by snow geese on coastal salt marshes. Ecology 62, 98–106. Stark, J.D., 2005a. A review and update of the report “Environmental and health impacts of the insect juvenile hormone analogue, S-methoprene” 1999 by Travis R.Glare and Maureen O’Callaghan. Report for New Zealand Ministry of Health, Wellington, 32pp. Stark, J.D., 2005b. Recommendations for estimating pesticide effects on non target organisms during mosquito eradication programmes in New Zealand. Report for New Zealand Ministry of Health, Wellington, 26pp. Stricker, J., 1995. Reviving wetlands. Wetlands (Australia) 14, 20–25. Sullivan, G., Noe, G.B., 2001. Coastal wetland plant species in Southern California. Appendix 2. In: Zedler, J.B. (Ed.), Handbook for Restoring Tidal Wetlands. CRC Press, Boca Raton, pp. 369–400. Talley, T.S., Crooks, J.A., Levin, L.A., 2001. Habitat utilization and alteration by the invasive burrowing isopod, Sphaeroma quoyanum in California salt marshes. Mar. Biol. 138, 561–573. Teal, J., Teal, M., 1969. Life and Death of the Salt Marsh. Little Brown and Co., Boston, 278pp. Thompson, G.B., Drake, B.G., 1994. Insect and fungi on a C3 sedge and a C4 grass exposed to elevated atmospheric CO2 concentrations in open-top chambers in the field. Plant Cell Environ. 17, 1161–1167. Titus, J.G., 2000. Does the U.S. government realize that the sea is rising? How to restructure federal programs so that wetlands can survive. Gold. Gate Univ. Law Rev. 30 (4), 717–778. Truscott, A., 1984. Control of Spartina anglica on the amenity beaches of Southport. In: Doody, P. (Ed.), Spartina anglica in Great Britain. Nature Conservancy Council, Attingham Park, pp. 64–69. Underwood, G.J.C., 1997. Microalgal colonization in a salt marsh restoration scheme. Estuar. Coast. Shelf Sci. 44, 471–481. United Nations, 2004. World Urbanization Prospects; the 2003 Revision. United Nations, New York, pp. xþ323 (including Annexes).
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US Army Corps of Engineers, 2005. Jamaica Bay Marsh Islands, Jamaica Bay, New York. Draft integrated ecosystem restoration report and environmental assessment. US Army Corps of Engineers, New York District, 140pp. Valdemoro, H.I., Sa´nchez-Arcilla, A., Jime´nez, J.A., 2007. Coastal dynamics and wetlands stability. The Ebro delta case. Hydrobiologia 577, 57–29. Valiela, I., 2006. Global Coastal Change. Blackwell Publishing, Oxford, 368pp. Wang, Q., An, S.Q., Ma, Z-J, Zhao, B., Chen, J.K., Li, B., 2006. Invasive Spartina alterniflora: biology, ecology and management. Acta Phytotaxon. Sin. 44, 559–588. Weaver, M., 2006. Flood scheme recreates ancient Essex wetlands. Guardian unlimited. http:// www.guardian.co.uk/conservation/story/0.1812368.00 html (accessed 06.07.06.). Weinstein, M.P., Philipp, K.R., Goodwin, P., 2000. Catastrophes, near-catastrophes and the bounds of expectation: success criteria for macroscale marsh restoration. In: Weinstein, M.P., Kreeger, D.A. (Eds.), Concepts and Controversies in Tidal Marsh Ecology. Kluwer, Dordrecht, pp. 777–825. Wiebe, W.J., Christian, R.R., Hanson, R.B., King, G., Sherr, B., Skyring, G., 1981. Anaerobic respiration and fermentation. In: Pomeroy, L.R., Wiegert, R.G. (Eds.), The Ecology of a Salt Marsh. Springer-Verlag, New York, pp. 136–159. Williams, J.M. (Ed.), 2006. Common Standards Monitoring for designated sites: first six year report. Habitats, JNCC, Peterborough, 72pp. Williams, P. and Associates, Faber, P.M., 2004. Design guidelines for tidal wetland restoration in San Francisco Bay. The Bay Institute and California State Coastal Conservancy, Oakland, 83pp. Williams, P.B., Faber, P.B., 2001. Salt marsh restoration experience in San Francisco Bay. J. Coast. Res. 27, 203–211. Williams, P.B., Orr, M.K., 2002. Physical evolution of restored breached levee salt marshes in the San Francisco Bay estuary. Restor. Ecol. 10, 527–542. Wolters, H.S., 2006. Restoration of salt marshes. PhD Thesis, Groningen University. Zedler, J.B., 1995. Salt marsh restoration: lessons from California. In: Cairns, J. (Ed.), Rehabilitating Damaged Ecosystems, second ed.. CRC Press, Boca Raton, FL, pp. 75–95. Zedler, J.B., 1996a. Ecological issues in wetland mitigation: an introduction to the forum. Ecol. Appl. 6, 33–37. Zedler, J.B., 1996b. Coastal mitigation in Southern California: the need for a regional restoration strategy. Ecol. Appl. 6, 84–93. Zedler, J.B. (Ed.), 2001. Handbook for Restoring Tidal Wetlands. CRC Press, Boca Raton, 439pp. Zedler, J.B., Adam, P., 2002. Salt marshes. In: Perrow, M.R., Davy, A.J. (Eds.), Handbook of Ecological Restoration, vol. 2. Restoration in Practice. Cambridge University Press, Cambridge, pp. 209–235. Zedler, J.B., Paling, E., McComb, A., 1990. Differential salinity responses to help explain the replacement of native Juncus kraussii by Typha orientalis in Western Australian salt marshes. Aust. J. Ecol. 15, 57–72.
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C H A P T E R
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M ANAGED R EALIGNMENT : R E - CREATING I NTERTIDAL H ABITATS ON F ORMERLY R ECLAIMED L AND Angus Garbutt and Laurence A. Boorman
Contents 1. Introduction 2. Location, Drivers, and Constraints to Managed Realignment 3. Site Evolution 3.1. Sediments 3.2. Creeks 3.3. Soils 3.4. Nutrient fluxes 3.5. Vegetation 3.6. Fishes 3.7. Spiders 3.8. Benthic invertebrates 3.9. Birds 4. Challenges in Managed Realignment Research Acknowledgments References
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1. INTRODUCTION Coastal wetlands have for many centuries been subject to alteration through management, land claim, and flood defense. It is now increasingly recognized that embankments, constructed to protect rural and urban land and infrastructure from tidal inundation, are becoming unsustainable (Turner et al., 2007). Rising sea levels, and an increase in the frequency of storm events, have added to the cost of improvement and maintenance of flood embankments and led policy makers to look for alternative strategies for coastal management. Maintaining the traditional “hold-the-line” policy is no longer regarded as a realistic long-term option for coastal management in many countries in northwest Europe (French, 2006). The current rigid coast line is not optimally designed to accommodate environmental Coastal Wetlands: An Integrated Ecosystem Approach
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change and continually improving defense standards are no longer economically viable, in some cases. In recent years the focus has moved toward managing coasts in a more dynamic way, protecting areas of high value whilst allowing natural coastal processes to proceed relatively unhindered elsewhere. Managed realignment, where coastal defenses are breached allowing previously reclaimed land to be reexposed to tidal inundation, has since the 1990s become increasingly used as a sustainable and cost-effective response to changing environmental conditions. Managed realignment involves moving coastal embankments landward toward a new line of defense or to higher ground under controlled conditions (Figure 1). One or several breaches are made in the original sea defense in most cases, rather than entire removal of the embankment, partly to reduce costs and partly to protect newly flooded land from high wave energy and current velocity. In this way, salt marshes that were embanked, drained, and used for agriculture are re-exposed to regular tidal inundation, many for the first time in decades or even centuries. In doing so, the intertidal area of the estuary is increased, which can help offset the impacts of sea-level rise. In addition, reinstating tidal inundation to reclaimed land recreates intertidal habitats, offsetting losses elsewhere, and acts as a more natural, “soft” defense against tidal flooding of the hinterland. Managed realignment is not without its problems, however, both in the availability of land to be realigned and
Figure 1 The Freiston Shore managed realignment scheme,The Wash (UK). Artificial creeks can be seen draining the site through three breaches made in the original flood embankment. In addition to the creation of salt marsh, a saline lagoon was created (top center of the picture) with gravel islands built for breeding birds. Image courtesy of the Environment Agency.
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uncertainties about the implications for the wider estuary. There is a requirement for a multifaceted approach to the implementation of schemes with the need to consider former land use, soils, estuarine processes, and changing environmental conditions such as sea-level rise, ecology, and socioeconomics. Much has been written about the concept of managed realignment (French, 2001), identifying possible sites (CEFAS, 2004) and site design (Legget et al., 2004). Because of the relative newness of the method, however, there are few data available on the outcomes of managed realignment schemes in Europe, nor is the theory well developed or practiced. Compare this to the wealth of literature from the United States where the science of salt marsh and intertidal habitat creation is well advanced (see Zedler, 2001, for an overview). The drivers, impacts, and subsequent development of managed realignment schemes in northwest Europe can differ from intertidal creation efforts elsewhere. The majority of studies on the restoration and functioning of salt marsh ecosystems comes from the United States and the distinctive Spartina alterniflora marshes, in particular (Craft et al., 1999; Morgan and Short, 2002). It is unclear whether generalizations about species reassembly, productivity, and ecosystem function in US marshes can be extrapolated to European systems where species, soils, and tidal range differ (Adam, 1990). What is not in doubt, however, are the advances in the development of a theoretical framework for the restoration of coastal wetland habitats made by US researchers. Restoration models have been developed which set out a framework that allows detailed description of the desired physical, chemical, and biological attributes of a restored ecosystem; could enhance biological diversity and functioning; identifies the constraints to restoration; and develops appropriate success criteria (Vivian-Smith, 2001). In Europe, monitoring is variable and ad hoc in uptake. Because of the relative newness of the technique, monitoring schemes have, to date, focused on describing the physical and biological development of individual sites.
2. LOCATION , D RIVERS , AND CONSTRAINTS TO MANAGED R EALIGNMENT The first deliberate managed realignment scheme was carried out on Northey Island, Blackwater Estuary, in southeast England. Here, in 1991, tidal flooding was reinstated to 0.8 ha of grassland for the first time since the area was embanked in 1873. Between one and seven schemes have been implemented throughout northwest Europe each year since 1991 from the north coast of France through to the German Baltic Sea (Wolters et al., 2005b; Figure 2). In mainland Europe, habitat re-creation driven by European legislation (C.E.G., 1992) has been the primary goal of managed realignment, particularly in The Netherlands and Germany. In the United Kingdom, coastal defense has been the primary reason for the implementation of most schemes, especially in the south and east of England where accelerated sea-level rise has increased the risk of flooding to coastal areas. The United Kingdom has the highest number of schemes proposed and implemented (RuppArmstrong and Nicholls, 2007), and as a result most available literature is of UK
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Figure 2 Location of managed realignment schemes in northwest Europe (adapted from Wolters et al., 2005b; Goeldner-Gianella, 2007; Rupp-Armstrong and Nicholls, 2007).
origin. UK sites rarely exceed 100 ha in size, whereas those on mainland Europe are usually over 100 ha. The most advanced estuarine management project to date is for the Scheldt Estuary that runs through The Netherlands and Belgium (van den Bergh et al., 2005). In a plan for the long-term management of the estuary, measures for navigation, safety against floods, and ecological functioning were adopted by both the Dutch and Flemish governments in 2004. The resulting “Sigmaplan” incorporated the strengthening and heightening of flood defenses and the construction of managed realignment sites and flood control areas, where flood waters are allowed to over-top embankments and subsequently retained in storage areas until water levels have receded. In total, this project is expected to create 1,150 ha of intertidal marsh and mudflat by managed realignment and 3,250 ha of additional wetlands through flood control areas and other measures. The concept of managed realignment is one that many coastal residents find hard to accept. Giving up land that has been hard won back to the sea creates strong feelings and most realignment schemes have been met with a degree of resistance from local people. When the concept was first introduced it was referred to as “set back” or “managed retreat” and whilst both terms are descriptive they portrayed a negative image of the method. Managed realignment has been adopted as a more positive term to describe the process of relocating the primary line of coastal defense to a more physically and economically viable position. For the first time, the dilemma of flood defense and coastal erosion has not been interpreted as a threat but as an opportunity for change. On the heavily populated coastal areas of northwest Europe, public acceptability is as important to the implementation of managed
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realignment as the financial and nature conservation considerations. The public reaction to managed realignment has been mixed (Myatt et al., 2003a–c; GoeldnerGianella, 2007). While some can accept the economic and conservation arguments, others are vehemently opposed to giving the land back to the sea, in some cases to such an extent that certain schemes have had to be abandoned due to local political and public pressure. In other cases, land for realignment has to be bought at a premium. A site in the United Kingdom was purchased at £13,750 per hectare in 2000, well above market rates, normally below £8,000 per hectare. In contrast to the planning and building of the Wallasea managed realignment site (UK), which was relatively rapid (30 months), the process of site selection and procurement took over 7 years (Scott, 2007; Figure 3). Whilst the method is still controversial, there remains a strong case for its use as a strategic tool for sustainable estuarine management on both economic and nature conservation grounds. Cost–benefit analysis shows that managed realignment can be more economically efficient than holding the current line of defense over a period greater than 25 years (Turner et al., 2007). Land claim alone has accounted for an estimated 25% loss of intertidal land in estuaries world wide (French, 1997) and some estuaries in Europe such as the Humber Estuary (UK) have lost up to 80% of their original area. Many European estuaries are canalized and embanked, leaving no room for morphological adjustment or rollover to a new equilibrium. As sea
Figure 3 The Wallasea managed realignment site on the Crouch Estuary (UK). A raised strip of mudflat can be seen on the landward edge of the site where dredged sediment was used to raise the elevation to that suitable for the colonization of salt marsh plants. Image courtesy of Defra.
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levels rise and global weather patterns lead to increased storminess and changes in wind direction, natural “soft” defenses are being eroded making the coast line more vulnerable to flooding. Moving the primary line of coastal defense inland and recreating mudflat or salt marsh in the intervening area has the dual benefits of habitat creation and increased flood defense with reduced costs. Salt marshes are extremely effective in dissipating wave energy, four times more efficient than sand flats (Mo¨ller et al., 1999). As a wave moves across the vegetated surface the energy levels and wave height decay exponentially. The presence of salt marsh in front of a sea defense can, therefore, be extremely valuable for coastal protection where, as the width of salt marsh increases, the height of the embankment required decreases (King and Lester, 1995).
3. S ITE EVOLUTION 3.1. Sediments Major embankment of salt marshes took place in northwest Europe between the 15th and 19th centuries, cutting off tidal inundation and subsequently sediment supply. The effect of compaction and dewatering on embanked marsh sediments coupled with relative sea-level rise has accentuated the differences between the embanked land and nearby salt marshes. In many cases a difference of 1.0–1.5 m exists between the surface of the realignment site and adjacent marshes. Therefore, unless the surface elevation of low-lying realignment sites is raised over time by natural sedimentary processes or raised artificially using imported sediment, intertidal marshes are not likely to be re-established at these sites. The rate of sedimentation within managed realignment sites can be extremely rapid in the early phases of implementation, with up to 50 mm per month being recorded at Paull Holme, the Humber Estuary (UK) (Boyes and Mazik, 2004). These rates are short lived, however, with low-lying areas building up sediment to find equilibrium profiles. Mathematical models and field evidence predicts an elevation–time curve for mudflat/marsh growth where sediment build-up rises steeply during the early stages but thereafter flattens off to an equilibrium level around that of relative sea-level rise (Temmerman et al., 2004). The detailed form of the curve depends on the balance between several parameters (sediment supply, tidal regime, and compaction under sediment load; Allen, 1990). The rate of minerogenic sediment accretion is determined chiefly by the tidal and finesediment regimes and is expected to be a decreasing function of mudflat elevation. This function can be seen at the Tollesbury managed realignment site where sediments were measured over a 12-year period (Figure 4). Mean sedimentation rates were 31 mm/year in the first year falling to 11 mm/year by year 11 as the overall elevation of the site increased. The long-term rates for sediment accumulation on the adjacent natural salt marshes were around 4 mm/year equivalent to predicted mean sea-level rise in the region (van der Wal and Pye, 2004). French (1999) warns that failure to allow for local fetch when planning such schemes may seriously impede sedimentation and vegetation establishment. By leaving the
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Figure 4 Mean change in the bed level (n = 20) of the Tollesbury managed realignment site (UK). Line shows fitted quadratic model used to account for the simple curvature in the data (y = ^79.32 þ 0.08927 day ^0.000007 day2; R2 = 99.4%, p-value for quadratic effect <0.001). Error bars show one standard error on either side of the mean.
hedges that were growing on the site prior to inundation to act as baffles and the original sea wall intact, erosion within the Tollesbury site has been restricted to the area around the breach. When restoring marshes with extremely low elevations, where high tidal inundations lead to restricted vegetation growth, Pethick and Burd (1995) recommended the provision of a creek system to reduce tidal flow velocities or a series of low bunds to reduce wave generation.
3.2. Creeks Creeks are an important and integral part of salt marshes, distributing sediments to the interior of the marsh, allowing aquatic organisms’ access to habitat, and providing drainage following tidal inundation (Desmond et al., 2000; Temmerman et al., 2005; Williams and Zedler, 1999; Wood and Hine, 2007). Salt marsh creek networks show a variety of morphological characteristics: parallel, meandering, braided, and interconnecting (Eisma, 1997), though most are commonly dendritic, with channels of varying sizes. Creeks develop through a variety of processes including headwater retreat, downward cutting, and lateral migration (French and Stoddart, 1992). Initial flow of water over intertidal flats occurs as sheet flow until subtle changes in topography channel focus water to depressions in the land surface, causing erosion and the formation of creeks (Whitehouse et al., 2000).
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Managed realignment projects are implemented on land that was formerly salt marsh prior to embankment, and creek design has usually tried to replicate the original marsh drainage system. The extent of the excavation is usually limited and, in most cases, only a main creek dug with perhaps one “first-order” tributary. In the Wadden Sea, embankments for both agricultural purposes and coastal protection have caused a major loss of mainland salt marshes. To offset these losses, some 1,100 ha of summer polders in the Dutch Noard-Fryslaˆn Buˆtendyks region are being restored to salt marsh (van Duin et al., 2007). Summer polders are areas of coastal grazing marsh in front of primary coastal defenses are protected from tidal flooding during all but the highest tides by earth embankments. In September 2001, three breaches were made in the embankments of the first polder, introducing tidal flooding to 135 ha of land. The creeks that were dug connecting the interior of the site to the breaches were dug deeper (1.5 m) and wider (6 m) than was thought to be necessary to transport efficiently water on and off the marsh. After a rapid period of sediment infilling, these channels took on a more natural meandering form, albeit within the confines of the excavated channel. The development of natural creek systems within managed realignment schemes varies. Creek development within the realignment site at Tollesbury (UK) only occurred in newly accreted marine sediments, rather than in the original agricultural substrate, once a critical depth between 20 and 30 cm was reached (Watts et al., 2003). There was no evidence of lateral erosion into the preinundation land surface, where low water velocity is thought to have been insufficient to cut down through the sediment. Once formed, the banks and adjacent margins of the newly developed creeks drained faster between tides and increased in stability and shear strength up to 30 times that of the surrounding sediments. Salicornia colonized the edges of the embryo creeks, whereas this plant does not occur on the adjacent sediments, emphasizing the functional role of creek systems. After a study of optimal creek design, French (1995) concluded that if existing or new nonoptimal lines of drainage are left in place following restoration of tidal flooding, they may compromise the efficiency with which surfaces are raised by sedimentation and lead to undesirable erosional adjustments. There are several examples of “accidental” realignments around the southeast coast of the United Kingdom where embankments have been breached during storm events and subsequently abandoned (French et al., 2000; Garbutt and Wolters, 2008). As a result, salt marsh has developed over the previously cultivated soils. One striking feature of a number of these sites is the persistence of old field drainage patterns. Some of these sites are over 100 years old, where the embankments were breached in a storm in 1897. The current creek network still resembles that of the field drains prior to the reinstatement of tidal flooding, with an orderly grid system of channels. This is with 1–2 m of marine sediment overlying the old agricultural surface. Managed realignment has the effect of increasing the tidal prism of an estuary and there is evidence that this causes morphological adjustment of existing creek systems. Symonds and Collins (2007) recorded the effect of managed realignment on an existing creek system on intertidal flats adjacent to the Freiston Shore site (UK). Here, the managed realignment site acted as a reservoir for water storage, following high water on spring tides. The 700,000 m3 of water that had entered the site at high
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water was continually released through the breaches in the embankments until the following high water. This “unnatural” discharge condition meant that existing creeks were at bank full conditions throughout the tidal cycle and forced excess water to escape as sheet flow. The excess water concentrated in existing creeks and depressions which caused accelerated creek development. Rapid annual landward extension rates of 400 m/year were recorded for existing creeks in the first 2 years after realignment. These continuous, high volumes of water leaving the site occurred over high tides for 2 years until the breaches in the embankment had eroded to an extent that all the water contained within the site could be discharged within one tidal cycle. During this period the creeks that had formed to accommodate the increased discharges from the site were several meters deep.
3.3. Soils The key difference between salt marsh creation through managed realignment and natural salt marsh development lies in the former being based on soils which, although marine in origin, have over many years become completely terrestrial in their chemical and physical characteristics. The changes that occur during this process of desalination are many and varied, but the nonreversible changes are critical. These changes include the consolidation of the soil through irreversible drying and the loss of organic matter through oxidation. Although prolonged flooding will increase the moisture content, the soil will never return to its original state. A sea wall failure over a century ago led to the development of secondary marshes at North Fambridge, southeast England, and the old soil surface, now more than a meter under the present marsh surface, is still visible as a distinct horizon forming a dense, hard layer (Hazelden and Boorman, 2001). Following 6 years of sediment build-up at the Tollesbury site the underlying agricultural sediments were both very strong and highly resistant to erosion (Watts et al., 2003). This “over consolidation” may have been due to the formation within the original agricultural sediments of an over-consolidated horizon with low hydraulic conductivity, forming an aquaclude, or barrier to water, that restricts subsurface drainage within the developing marsh sediments. Crooks (1999) found that the presence of suitable quantities of detrital calcium carbonate is important in creating regional differences in the formation of these horizons. The calcium-deficient alluvium of the Thames Estuary region appears to be particularly sensitive to the effects of a lowered saline water table (on reclamation), inducing the deflocculation of clays and the formation of a dense horizon. In contrast, the sediments in the carbonate-bearing alluvium of the Severn Estuary are insensitive to pore water salinity changes on reclamation giving an absence of over-consolidated horizons. Reintroduction of tidal flooding to reclaimed land leads to major changes in the chemistry and biogeochemistry of the sediment–soil system. Conditions change from an oxidizing to a reducing environment, with changes in pH and nutrient levels and increases in the extractability of heavy metal contaminants (Emmerson et al., 2001; Blackwell et al., 2004). Additional work is needed to understand how future managed realignment sites will develop in terms of the redistribution of sediments and the transport of contaminants (Chang et al., 2001).
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3.4. Nutrient fluxes Recent studies have shown that nutrient cycling is an integral part of the internal functioning of salt marshes and that salt marshes provide a functional link between the nutrient resources of land and sea (Boorman, 2000). It is unclear, however, whether these observations apply equally to salt marshes created by managed realignment. Relatively little is known about the timescales needed for the development of mature salt marsh, but the vegetation communities of natural marshes have been shown to possess a degree of stability that lasts at least for several decades (Adam, 1990). Assessing the success of salt marsh creation schemes could best be judged not by the assemblage of plant or animal communities but by assessing whether the normal regime of salt marsh nutrient fluxes has been restored (Craft et al., 1999). The creation of large areas of new marshes through managed realignment could markedly increase annual sinks for the uptake of nitrogen, organic carbon, and reactive phosphorus, albeit at the expense of enhanced greenhouse gas emissions by new marshes in less saline areas of the estuary. Recently cost–benefit analyses of managed realignment in the Blackwater Estuary have shown a positive benefit from enhanced burial of carbon, nitrogen, and phosphorus as well as enhanced metabolism of dissolved nitrate (Shepherd et al., 2005). It seems likely from the monitoring of created salt marshes that the development of mature salt marshes with the full range of associated plant species is likely to be a much longer and slower process than that of the initial plant colonization. However, a recent French study showed that the fluxes normally associated with an immature marsh community were restored within 14 years of reconnecting the area to the sea (Dausse et al., 2005). It has been suggested that sediment input and accumulation as a result of marsh creation could enhance the storage of pollutant heavy metals (Andrews et al., 2006). However, there is the risk that reworking of sediments could result in significant quantities of these pollutants again being released into the environment (Chang et al., 2001; Boorman, 2003). The same could also apply to any reworking of existing marine sediments during the initial processes of salt marsh creation. The presence of salt marsh pollutants in intertidal sediments could, through sediment redistribution, be a long-term threat to salt marsh survival (Leggett et al., 1995). Studies on agricultural soils from the managed realignment site at Tollesbury, however, showed that this is unlikely to lead to the release of damaging levels of heavy metals to the environment (Emmerson et al., 2001). The salt marsh plants involved in marsh creation themselves require nutrients for their growth and both nitrogen and phosphorus levels are essential. Studies in a range of natural western European marshes showed little correlation between soil nitrogen levels and plant productivity although Dutch studies showed that the addition of nitrogen stimulated the growth of salt marsh plants (Kiehl et al., 1997). However, marked declines in levels of available phosphorus during the growing season and reductions in production of plant material were attributed to low phosphorus levels in the soil and it was noted that high productivity in marshes in the Western Scheldt and Portugal were correlated with raised phosphate levels (Boorman, 2000). It should also be noted that excessive levels of either nitrogen or
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phosphorus can have damaging environmental effects. The eutrophication of coastal waters can result in the rapid growth of mat-forming algae and these mats can delay colonization by higher plants by smothering the growth of pioneer salt marsh species such as Salicornia spp. (Underwood, 1997).
3.5. Vegetation Self-sustaining plant communities are often a primary goal of habitat creation efforts within managed realignment sites as they can perform some of the biological and economically desirable functions of wetland ecosystems (Zedler and Lindig-Cisneros, 2000). Specific targets are often lacking, however, with the extent of salt marsh or mudflat frequently used as the only measure of success. In an effort to provide a generic measure of success for salt marsh restoration projects in northwest Europe, Wolters et al. (2005b) devised a saturation index where the presence of target species in a restoration site was expressed as a percentage of a regional target species pool. The species pool was based on the three biogeographical regions of salt marsh vegetation covering northwest Europe and target species identified within these zones. The saturation index was based on an ideal “best-case scenario” where all species had an equal chance of colonization from pioneer through to transitional communities. Of the European countries examined, the restoration sites in the United Kingdom were the least diverse with the majority of sites having saturation indices of fewer than 30%. Many of the high marsh species, and those associated with transition to terrestrial communities, were absent including Poa subcaerulea, Puccinellia fasciculata, Carex serotina, Blysmus rufus, Oenanthe lachenalii, and Ononis repens. This absence is likely to be a function of the restricted elevation of many sites, rarity in the species pool leading to limited dispersal opportunities, or biotic or abiotic constrains with sites. The most common species were Salicornia spp., Suaeda maritima, Aster tripolium, and Puccinellia maritima, which were recorded from over 80% of sites. These species are all typical of pioneer or low marsh communities reflecting the low elevation of most sites. Planting of salt marsh vegetation, an established technique in North America (Sullivan, 2001), is rarely practiced in managed realignment sites in northwest Europe. Experimental planting, where it has occurred, has not been a great success (Garbutt et al., 2006). Sites in low-energy environments, located near natural marshes, regenerate relatively rapidly. After removal of the flood defense embankment in 1993, the Karrendorfer Wiesen, Baltic Sea, Germany, was re-exposed to tidal inundation for the first time since 1850 (Bernhardt and Koch, 2003). Five years after removal of the embankment, nearly 75% of the 350 ha site, which was formerly improved grassland, was covered by salt marsh plant species as a result of natural colonization. Several factors influence the establishment and development of vegetation within managed realignment sites. Former land use, the physical properties of the formerly embanked soils and marine sediments, a viable seed bank, and the dispersal of plant propagules into new sites can all affect the colonization process and the future development of the salt marsh (Wolters et al., 2005c). However, elevation in relation to the tidal frame is the single most important factor determining the eventual outcome of intertidal habitat creation
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efforts. Until the mudflat surface reaches an appropriate elevation, salt marsh will not develop. Although the reliability of managed realignment as an appropriate technique for salt marsh restoration has been under debate (Hughes and Paramor, 2004; Morris et al., 2004; Wolters et al., 2005a), experience so far has shown that, where the elevation has been suitable, vegetation has colonized. The majority of salt marsh species do not form a persistent seed bank (Thompson et al., 1997; Wolters and Bakker, 2002), and in embanked areas the seed bank of former marshes is likely to decline rapidly. Agricultural practices such as ploughing would further deplete the seed bank either by destroying the seeds or stimulating germination. The regeneration of salt marsh vegetation after managed realignment therefore relies on regular tidal inundation as the key agent for the dispersal of diaspores into the site. Salt marsh seeds can float in sea water for varying amounts of time from a few hours to several months. Most seeds retain their viability in salt water (Koutstaal et al., 1987) and germinate when exposed to freshwater conditions (Woodell, 1985). After realignment, relatively rapid colonization of pioneer and low marsh species occurs, provided the plants are present in a nearby source area and the elevation is suitable. Land use prior to inundation can have an affect on the initial development of salt marsh vegetation. In an experiment to test the effect of preinundation land use on the colonization of plant species, plots that were covered in grasses or cereal stubble prior to flooding proved the most effective for colonization by pioneer species (Garbutt et al., 2006). Shoot densities of Salicornia spp. were about 40% higher than plots left unvegetated (both bare earth or ploughed ridge and furrow). The remnants of the terrestrial vegetation, even when killed off by saltwater inundation, provided the best surface for initially trapping propagules and subsequently retaining seedlings. In sites where sediment accumulation is rapid, waterlogging may affect salt marsh evolution. Watts et al. (2003) found that after 6 years, the newly accreted sediments within the Tollesbury realignment site on the Blackwater Estuary (UK) were characterized by high water content, low dry bulk density, and low resistance to erosion, presumably because of the lack of time to consolidate. This has considerable implications for the development of vegetation within realignment areas as waterlogged sediments are important in determining the lower limits of plants in the tidal frame and the course of vegetation succession. Within the Tollesbury site Salicornia europaea agg. grew at elevations down to its expected lower limit fulfilling the predicted extent of salt marsh development. However, Garbutt et al. (2006) found that the lower elevation of other salt marsh plants were on average 0.19 m higher than those of the adjacent natural marshes after 6 years of tidal inundation. If an adequate drainage system fails to develop, the site may produce a long-standing waterlogged mudflat fringed by relatively static salt marsh vegetation. Where there has been very little sedimentation, the growth of salt marsh plants appears unrestricted on the formerly embanked soils. Colonization of the high areas of the Sieperda Marsh (The Netherlands) where sedimentation was less than 5 mm/year and Freiston Shore (UK) was rapid (Eertman et al., 2002; Badley and Allcorn, 2006). Plant species diversity comparable to adjacent marshes was restored within 3 years at Freiston Shore (UK) although abundance differed significantly.
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Several studies have examined salt marshes that have regenerated on reclaimed land following breaches in embankments due to storm events as analogues for managed realignment. These reports show that in some cases it can take decades for plant communities to reach equilibrium with reference conditions, and the regenerated plant communities can lack certain species (Burd, 1992; Onaindia et al., 2001). Garbutt and Wolters (2008) surveyed the plant communities of regenerated salt marshes within managed realignment sites and sites of abandoned reclamations up to 107 years old in southeast England. Plant diversity within regenerated marshes was lower than in adjacent natural marshes and correlated with a greater abundance of Spartina anglica. The plant communities were, in general, representative of less diverse communities or those typical of lower elevations, with a higher tidal inundation frequency. Vegetation height within the regenerated marshes was also significantly greater as a result of the presence of Spartina. The development of a complete flora may have been constrained by limited dispersal of plant propagules or lack of appropriate physical conditions. Although the propagules of salt marsh plants have the potential to travel great distances on tidal currents, the evidence is that this does not happen (Wolters et al., 2005c). Given the potential lack of long distance dispersal for salt marsh plant propagules, the location of target communities close to restoration sites could be fundamental in determining the eventual composition of the vegetation. Whilst S. anglica is now currently accepted as an integral part of the flora of European salt marshes (Lacambra et al., 2004), the presence of this species in the early stages of salt marsh restoration may be less desirable. The successful establishment and spread of this neo-endemic species throughout Europe during the 20th century has been well documented and was largely due to the species perennial life history and the existence of a zone of mudflat formerly unoccupied by salt marsh plants (Gray et al., 1990). The sparsely vegetated mudflats that typify the early phases of salt marsh development within realignment sites provide ideal conditions for the establishment of Spartina. Figure 5 shows the expansion of plant species in the Tollesbury realignment site (UK) during the first 8 years of Restoration time m OD 2.8 2.3 2.0 1.8
2
3
4
5
6
7
8 Bare soil Salicornia spp. Suaeda maritima Puccinellia maritima Spartina anglica
1.7 20 m
1.6
Figure 5 Colonization of pioneer species within the Tollesbury managed realignment site between 1997 (year 2) and 2003 (year 8). Transects are 20 m wide and 125 m long running from high ground down to low mudflat. Elevation is shown as meters above UK Ordnance Datum (m OD). Mean high water neap tides for the site are at 1.50 m OD (Source: MinekeWolters).
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development. There was a 30% expansion in frequency from year 4 when Spartina was first recorded within the site to year 8 (Wolters et al., 2006). Spartina anglica can drastically alter the sedimentary and drainage characteristics of its surroundings leading to the creation of waterlogged and anoxic soils (Doody, 1984). The establishment of Spartina early in the development of salt marsh within managed realignment sites gives the regenerate marshes a very different starting point to that of natural marshes and may affect the eventual outcome of creation efforts.
3.6. Fishes It is well recognized that intertidal marshes provide important nursery and rich feeding areas for young fish (Laffaille et al., 2000; Mathieson et al., 2000). During the crucial early development stage in the life history of fishes, the availability of such habitats plays an important role in early growth and survival of some species, and consequently an important role in the recruitment process (Laffaille et al., 2001). Much work has been done in North America on fish usage of restored wetlands (Williams and Desmond, 2001), whilst in Northern Europe the importance of managed realignment for fishes has only recently begun to be recognized. Managed realignment sites can provide valuable habitat for fish species, both for foraging and as refugia. The majority of fish caught within realignment sites are juveniles, dominated by the 0þ year class (Colclough et al., 2005; Pinder et al., 2007). The depth of water is often not sufficient for larger fish to risk entering, particularly during daylight for risk of stranding or predation by birds. The abundance of pioneer vegetation, at least in the early years of a sites development, is thought to be advantageous to the young fish both as cover and in foraging around the stems of plants. Large shoals of fry can be observed within sites moving with the tide, often on the edge of the tide/land interface. Standing water at low tide, either in the form of old drainage channels or ponds acting as brackish lake systems, can contain large numbers of fish acting as brackish lake systems, offering continuous refuge and food cover over each tidal cycle. At Abbotts Hall, Blackwater Estuary (UK), Colclough et al. (2005) captured around 2000 0þ year group herring (Clupea harengus) at low tide in one seine net haul from an old freshwater ditch. Similar results were found at Freiston Shore (UK) (Table 1), where the site acted as a nursery area for a range of different fish species, including economically important bass (Dicentrarchus labrax), sprat (Sprattus sprattus), and herring (Pinder et al., 2007). Gut analysis of juvenile fish using Freiston Shore showed the site provided a nursery habitat throughout the entire tidal cycle with the continuous utilization of permanently flooded channels and food resources within these waterbodies. This study suggests that constructing additional areas of standing water within realignment sites would enhance the quality of this habitat for juvenile fishes. Creating additional water bodies would increase available habitat beyond the period of spring tide inundation at a site, thus decreasing competition for food resources and promoting enhanced growth rates and survival.
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Table 1 Number of fish caught in the Freiston Shore managed realignment site, The Wash (UK) between 2003 and 2006 (after Pinder et al., 2007) Common name
Scientific name
Sand goby Three-spined stickleback Sprat Herring Smelt Flounder Bass Poor cod Sand smelt Mullet Eel Nilssons pipefish
Pomatoschistus minutus Gasterosteus aculeatus Sprattus sprattus Clupea harengus Osmerus eperlanus Platichthys flesus Dicentrarchus labrax Trisopterus minutus Atherina presbyter Mugilidae sp. Anguilla anguilla Syngnathus rostellatus
Total fish caught
2003
2004
2005
2006
835 697 99
3,240 45 413
3,742 16 169
1,767 165 158
49 84
1 20
1 1 1
1
4 3 1 1 1
1,641
5
3,834
21
2
28
3,951
2,144
Numbers of sprat and herring are grouped.
3.7. Spiders Spiders are known to react strongly to changes in microhabitat conditions and are consequently often used as indicators of the effects of habitat changes (Marc et al., 1999; Bell et al., 2001). In a study aimed at evaluating the response of arthropod communities to the restoration of biodiversity, Pe´tillon and Garbutt (2008) found differences between the abundance and composition of spiders in recent managed realignment schemes and natural salt marshes. Ground-dwelling spiders were compared between three managed realignment sites, aged between 3 and 14 years and adjacent natural salt marshes in the Blackwater Estuary (UK). A total of 27 species of spider was caught in pitfall traps during the survey (Table 2). The natural salt marshes were characterized by low species richness, the dominance of late successional stage species such as Pirata piraticus, and the exclusive presence of large species, preferring close habitats like the rare Arctosa fulvolineata that lives in the litter and debris of salt marsh habitats (Pe´tillon et al., 2005). The regenerated salt marshes of the realignment sites had greater species richness. The halophilic wolf spider (Pardosa purbeckensis) was the most abundant species present with other halophiles (Enoplognatha mordax and Erigone longipalpis) also recorded. The greater abundance and diversity of spiders within the regenerating marshes reflected greater habitat diversity and areas of bare ground, favoring pioneer species (mainly linyphiid). The natural marshes of the study area are classic examples of the type affected by “coastal squeeze,” truncated by coastal defenses at the landward edge with salt marsh cliffs at the seaward edge, leaving salt marshes with only one or two zones (typically mid-marsh). The lack of transitional communities (mudflat to low marsh and high marsh to terrestrial vegetation) undoubtedly limits the available habitat for many arthropod species. The implementation of managed realignment
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Table 2 Species richness, number of individuals, and abundances of ground-dwelling spiders (more than five individuals) collected in pitfall traps from the Abbotts Hall, Tollesbury and Northey Island managed realignment sites (UK) and adjacent natural salt marshes Parameter
Habitat
F-ratio
Natural
Reclaimed
Total species richness Total number of individuals Mean species richness Mean number of individuals Abundance of Pardosa purbeckensis Pirata piraticus Oedothorax apicatus Oedothorax fuscus Pardosa prativaga
15 61 1.3 + 0.2
21 230 2.4 + 0.3
6.65
2.0 + 0.4
7.7 + 1.8
9.10
0.5 + 0.2 0.8 + 0.3 0.0 + 0.0 0.03 + 0.03 0.2 + 0.1
3.3 + 1.0 0.3 + 0.1 1.5 + 0.5 1.6 + 0.5 0.1 + 0.1
7.84 1.78 10.80 9.53 1.08
Tenuiphantes tenuis
0.1 + 0.1
0.2 + 0.1
1.44
p
n.s.
n.s. n.s.
Mean parameters are compared between reclaimed and natural sites by ANOVA (59 d.f.). p < 0.05. p < 0.01. n.s., nonsignificant.
schemes in the estuary increased the available habitat for ground-dwelling spiders, particularly, where there was a transition from mudflat to terrestrial vegetation.
3.8. Benthic invertebrates Studies of benthic invertebrate colonization into regenerated mudflats in Europe are limited, largely due to a limited number of schemes that have been implemented. Where results are available, they indicate relatively rapid colonization of sites, although this relies on an adequate supply of sediment in the system, as colonization is limited in the former agricultural soils. Evans et al. (1998) found that colonization of a created area of intertidal land by invertebrates was slower than expected due to compaction caused by earth-moving equipment and the low organic content of the mud. Garbutt et al. (2006) observed that invertebrate colonization only occurred within the newly accreted sediments of the Tollesbury site (UK) and benthic organisms were absent from the original, consolidated agricultural substrate. The time taken for species to colonize a site will be affected by their life history and availability in the local species pool. Species that are mobile as adults are recorded in the early stages of colonization. For example, the crustacean Corophium volutator regularly occurs in large numbers at several sites (Atkinson et al., 2001). Mazik et al. (2007) recorded a species with poor microinvertebrate assemblage, 1 year after the creation of the Paull Holme managed realignment site, composed predominantly of early colonizing species of Oligochaeta and Nematoda, although a total of 20 species of invertebrate were recorded. The oligochaete Paranais litoralis accounted for 53% of the invertebrate abundance in the first year the site was
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created (2004) and only 2% 3 years later. More sedentary species, such as bivalves, rely on planktonic stages to colonize new sites and it may be several years before a stable population of these relatively long-lived species becomes established. Mazik et al. (2007) found that by year 3 with in the Paull Holme site, Hediste diversicolor and Hydrobia ulvae were the most abundant species with low numbers of Macoma baltica and Streblospio shrubsolii. Fourteen species of intertidal invertebrate were recorded within the Tollesbury site after 2 months of tidal inundation with size and density comparing favorably with a nearby reference site after 6 years (Garbutt et al., 2006). After a 5-year period, high densities of invertebrates had colonized the newly developed intertidal areas of the Sieperda Marsh in the Scheldt Estuary where Corophium volutator and Nereis diversicolor dominated the biomass (Eertman et al., 2002). Colonization of newly created mudflats is not always as rabid as most available data suggest. At the Orplands site (UK), bivalve mollusc species, present in substantial numbers in the nearby mudflats, failed to colonize the sediments of the site in the first 4 years, possibly as a result of anoxic sediments (Environment Agency, 1999).
3.9. Birds Studies of waterbird assemblages on created intertidal wetlands are limited (Atkinson et al., 2001, 2004; Badley and Allcorn, 2006; Mander et al., 2007). Studies that have been undertaken found that birds are quick to colonize, either for roosting or to feed. In the first few years after realignment, waterbird assemblages are generally variable and undergo large changes adjusting to the biological and physical evolution of the site (Atkinson et al., 2004). The first few years are characterized by a dominance of passerine species such as sky lark (Alauda arvensis), meadow pipit (Anthus pratensis), and reed bunting (Emberiza schoeniclus), which either exploit the abundance of seed produced by early colonizing annual plant species or amphipods that live in the detritus produced by the dieback of terrestrial vegetation (Badley and Allcorn, 2006). Mander et al. (2007) found that within 3 years of creation, the Paull Holme site (UK) was capable of supporting a functional waterbird assemblage of similar composition to that of adjacent existing intertidal areas at low water. Here, mallard (Anas platyrhynochos) and teal (Anas cracca) were abundant in the first year after the creation of the site, where they were found to be feeding on seeds retained within the site. In contrast, numbers of wading birds were low in the first year after creation but increased thereafter, reflecting the development of the mudflats within the site. The subsequent development and establishment of stable bird communities will be determined by the evolution of individual sites and the availability of suitable prey species or safe areas to roost. If mudflat develops, waterbirds are quick to exploit the resource, with species that forage on small polychaetes that are likely to colonize the area before those that feed on larger bivalves. Managed realignment tends to create high-level mud in the early phases of development, and observations from several sites suggest that some species of wading bird such as dunlin (Calidris alpina) and redshank (Tringa tetanus) use these areas to provide additional feeding time, either side of high tide, when preferred
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feeding areas are inundated. This is particularly important for birds in poor condition or during severe winter weather when access to extra prey resources could increase chances of survival. Where salt marsh develops with a sparse cover of vegetation, large numbers of golden plover (Pluvialis apricaria) and lapwing (Vanellus vanellus) typically use these areas for roost sites (Badley and Allcorn, 2006; Mander et al., 2007).
4. C HALLENGES IN M ANAGED R EALIGNMENT R ESEARCH Managed realignment has become an increasingly used and widely accepted method of coastal management, and is seen as a way to manage estuaries more naturally and economically as they adapt to changing climatic conditions. The method is attractive to policy makers and coastal managers alike. The potential scale, and the design and siting of particular schemes remain fairly contentious issues, however, with some stakeholder disputes unresolved and disagreement about the wider impact. Pethick (2002) calculated that a relatively small site of 100 ha may increase the tidal prism by as much as 1 106 m3. In a small estuary this could give rise to a significant increase in discharge and tidal current velocity leading to erosion of the outer estuary channel, further threatening flood defense embankments and intertidal habitat. In view of this, any realignment of a flood embankment within an estuary should be part of a large-scale design in which the morphodynamics of the whole estuary are incorporated. Monitoring hydrological, physical, and biological processes on an estuary on a wide scale are necessary if the reaction of the estuary to the realignment of its boundaries is to be better understood. Long-term, strategic monitoring of realignment sites is essential if eventual outcomes are to be recognized. Whilst the permissions and consents needed to implement a managed realignment scheme can be difficult to obtain and take several years to complete, site creation is relatively straightforward and possible with minimal management by allowing tidal ingress through a simple, relatively small breach in the sea defenses. There are now several studies that have shown that the landward realignment of sea defenses will quickly produce intertidal mudflats on low-lying agricultural land which are then colonized by invertebrates, and if the elevation is appropriate, salt marsh plants. There are still large gaps in knowledge about the restoration of intertidal habitats in northwest Europe, however (French, 2006). The potential constraints of the underlying agricultural substrate, seed dispersal, and the addition of S. anglica to the European flora, mean that it is uncertain whether regenerated salt marshes will ever achieve equivalent biological diversity and structure to that of natural marshes. Whether differences in diversity and structure mean that regenerated salt marshes perform different functional roles to natural marshes is also unclear. This does not mean the inevitable failure of managed realignment, rather that restored habitats rarely replicate all the structural and functional attributes of reference communities. Re-created habitats should ideally function within the normal variation found in “natural” systems, thereby retaining key attributes such
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as diversity, vegetation structure, and ecological processes. The challenge now is to identify the processes that determine such outcomes and to define criteria by which the success of managed realignment schemes in northwest Europe can be measured.
ACKNOWLEDGMENTS The authors would like to thank Erica van den Burgh, Willem van Duin, Colin Scott, Julien Pe´tillon, and Adrian Pinder for valuable discussion and literature provided for this article. We are grateful for the constructive comments of Jon French and Laurence Rozas on the draft manuscript.
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King, S.E., Lester, J.N., 1995. Pollution economics. The value of salt marshes as a sea defence. Mar. Pollut. Bull. 30, 180–189. Koutstaal, B.P., Markusse, M.M., De Muncke, W., 1987. Aspects of seed dispersal by tidal movements. In: Huiskes, A.H.L., Blom, C.W.P.M., Rozema, J. (Eds.), Vegetation Between Land and Sea: Structure and Processes. W. Junk Publishers, Dordrecht, pp. 226–233. Lacambra, C., Cutts, N., Allen, F., Burd, F., Elliott, M., 2004. Spartina anglica: a review of its status, dynamics and management. English Nature Research Reports 527, Peterborough. Laffaille, P., Feunteun, E., Lefeuvre, J.C., 2000. Composition of fish communities in a European macrotidal salt marsh (the Mont Saint-Michel Bay, France). Estuar. Coast. Shelf Sci. 51, 429–438. Laffaille, P., Lefeuvre, J.C, Schricke, M.T., Feunteun, E., 2001. Feeding ecology of 0-group sea bass, Dicentrarchus labrax, in salt marshes of Mont Saint-Michel Bay, France. Estuaries 24, 116–125. Leggett, D.J., Bubb, J.M., Lester, J.N., 1995. The role of pollutants and sedimentary processes in flood defence. A case study: salt marshes of the Essex coast, UK. Environ. Technol. 16, 457–466. Leggett, D.J., Cooper, N.J., Harvey, R., 2004. Coastal and Estuarine Managed Realignment – Design Issues. CIRIA, London. Mander, L., Nicholas, D.C., Allen, J., Mazik, K., 2007. Assessing the development of newly created habitat for wintering estuarine birds. Estuar. Coast. Shelf Sci. 75, 163–174. Marc, P., Canard, A., Ysnel, F., 1999. Spiders (Araneae) useful for pest limitation and bioindication. Agric. Ecosyst. Environ. 74, 1–46. Mathieson, S., Cattrijsse, A., Costa, M.J., Drake, P., Elliot, M., Gardner, J., Marchand, J., 2000. Fish assemblages of European tidal marshes: a comparison based on species, families and functional guilds. Mar. Ecol. Prog. Ser. 204, 225–242. Mazik, K., Smith, J.E., Leighton, A., Elliott, M., 2007. Physical and biological development of a newly breached managed realignment site, Humber estuary, UK. Mar. Pollut. Bull. 55, 564–578. Mo¨ller, I., Spencer, T., French, J.R., Leggett, D.J., Dixon, A.M., 1999. Wave transformation over saltmarshes: a field and numerical modelling study from North Norfolk, England. Estuar. Coast. Shelf Sci. 49, 411–426. Morgan, P.A., Short, F.T., 2002. Using functional trajectories to track constructed salt marsh development in the Great Bay Estuary, Maine/New Hampshire, USA. Restor. Ecol. 10, 461–473. Morris, R.K.A., Reach, I.S., Duffy, M.J., Collins, T.S., Leafe, R.N., 2004. On the loss of saltmarshes in south-east England and the relationship with Nereis diversicolor. J. Appl. Ecol. 41, 787–791. Myatt, L.B., Scrimshaw, M.D., Lester, J.N., 2003a. Public perceptions and attitudes towards a current managed realignment scheme: Brancaster West Marsh, North Norfolk, UK. J. Coast. Res. 19, 278–286. Myatt, L.B., Scrimshaw, M.D., Lester, J.N., 2003b. Public perceptions and attitudes towards a forthcoming managed realignment scheme: Freiston Shore, Lincolnshire, UK. Ocean Coast. Manage. 46, 565–582. Myatt, L.B., Scrimshaw, M.D., Lester, J.N., 2003c. Public perceptions and attitudes towards an established managed realignment scheme: Orplands, Essex, UK. J. Environ. Manage. 68, 173–181. Onaindia, M., Albizu, I., Amezaga, I., 2001. Effect of time on the natural regeneration of salt marsh. Appl. Veg. Sci. 4, 247–256. Pethick, J., 2002. Estuarine and tidal wetland restoration in the United Kingdom: policy versus practice. Restor. Ecol. 10, 431–437. Pethick, J., Burd, F., 1995. Sedimentary processes under managed retreat. Saltmarsh Management for Flood Defence, Research Seminar Proceedings, November 1995. National Rivers Authority, Project Record 480/1/SW, pp. 14–25. Pe´tillon, J., Garbutt, A., 2008. Success of managed realignment for the restoration of salt-marsh biodiversity: preliminary results on ground-active spiders. J. Arachnol. 36, 388–393. Pe´tillon, J., Ysnel, F., Lefeuvre, J.-C., Canard, A., 2005. Are salt marsh invasions by the grass Elymus athericus a threat for two dominant halophilic wolf spiders? J. Arachnol. 33, 236–242. Pinder, A.C., Scott, L.J., Bass, J.A.B., 2007. Fish utilisation of Freiston Shore realignment: 2003–2006 surveys. Final report to the Environment Agency. Centre for Ecology and Hydrology, Winfrith, Dorset, UK. Rupp-Armstrong, S., Nicholls, R.J., 2007. Coastal and estuarine retreat: a comparison of the application of managed realignment in England and Germany. J. Coast. Res. 23, 1418–1430.
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Scott, C.R., 2007. Wallasea wetland creation scheme: lessons learned. Proceedings of the CIWEM Rivers and Coastal Group Winter Meeting: From Directive to Detail: A Joined-Up Response to Flooding?, 26 January 2007, London. Shepherd, D., Jickells, T., Andrews, J., Cave, R., Ledoux, L., Turner, K., Watkinson, A., Aldridge, J., Malcolm, S., Parker, R., Young, E., 2005. Integrated modelling of an estuarine environment: an assessment of managed realignment options. Technical Report 21, Tyndall Centre for Climate Change. Sullivan, G., 2001. Establishing vegetation in restored and created wetlands. In: Zedler, J.B. (Ed.), Handbook for Restoring Tidal Wetlands. CRC Press, New York, pp. 119–155. Symonds, A.M., Collins, M.B., 2007. The establishment and degeneration of a temporary creek system in response to managed coastal realignment: The Wash, UK. Earth Surf. Proc. Land. 32, 1783–1796. Temmerman, S., Bouma, T.J., Govers, G., Lauwaet, D., 2005. Flow paths of water and sediment in a tidal marsh: relations with marsh developmental stage and tidal inundation height. Estuaries 28, 338–352. Temmerman, S., Govers, G., Wartel, S., Meire, P., 2004. Modelling estuarine variations in tidal marsh sedimentation: response to changing sea level and suspended sediment concentrations. Mar. Geol. 212, 1–19. Thompson, K., Bakker, J.P., Bekker, R.M., 1997. Seed Banks of North West Europe: Methodology, Density and Longevity. Cambridge University Press, Cambridge. Turner, R.K., Burgess, D., Hadley, D., Combes, E., Jackson, N., 2007. A cost-benefit appraisal of coastal managed realignment policy. Glob. Environ. Change 17, 397–407. Underwood, G.J.C., 1997. Microalgal colonisation in a salt marsh restoration scheme. Estuar. Coast. Shelf Sci. 44, 471–481. van den Bergh, E., van Damme, S., Graveland, J., de Jong, D., Baten, I., Meire, P., 2005. Ecological rehabilitation of the Scheldt Estuary (The Netherlands-Belgium; north-west Europe): linking ecology, safety against floods, and accessibility for port development. Restor. Ecol. 13, 204–214. van der Wal, D., Pye, K., 2004. Patterns, rates and possible causes of saltmarsh erosion in the Greater Thames area (UK). Geomorphology 61, 373–391. van Duin, W.E., Esselink, P., Bos, D., Klaver, R., Verweij, G., van Leeuwen, P.-W., 2007. Proefverkweldering Noard-Fryslaˆn Buˆtendyks, Evaluatie kwelderherstel 2000–2005. Wageningen IMARES, Texel, Koeman en Bijkerk b.v. rapport 2006-045, Haren, Altenburg and Wymenga rapport 840, Veenwouden. Vivian-Smith, G., 2001. Developing a framework for restoration. In: Zedler, J.B. (Ed.), Handbook for Restoring Tidal Wetlands. CRC Press, New York, pp. 39–88. Watts, C.W., Tolhurst, T.J., Black, K.S., Whitmore, A.P., 2003. In situ measurements of erosion shear stress and geotechnical shear strength of the intertidal sediments of the experimental managed realignment scheme at Tollesbury, Essex, UK. Estuar. Coast. Shelf Sci. 58, 611–620. Whitehouse, R.J.S., Bassoullet, P., Dyer, K.R., Mitchener, H.J., Roberts, W., 2000. The influence of bedforms on flow and sediment transport over intertidal mudflats. Cont. Shelf Res. 20, 1099–1124. Williams, G.D., Desmond, J.S., 2001. Restoring assemblages of invertebrates and fishes. In: Zedler, J.B. (Ed.), Handbook for Restoring Tidal Wetlands. CRC Press, New York, pp. 235–270. Williams, G.D., Zedler, J.B., 1999. Fish assemblage composition in constructed and natural tidal marshes of San Diago Bay: relative influence of channel morphology and restoration history. Estuaries 22, 702–716. Wolters, M., Bakker, J.P., 2002. Soil seed bank and driftline composition along a successional gradient on a temperate salt marsh. Appl. Veg. Sci. 5, 55–62. Wolters, M., Bakker, J.P., Bertness, M.D., Jefferies, R.L., Moller, I., 2005a. Saltmarsh erosion and restoration in south-east England: squeezing the evidence requires realignment. J. Appl. Ecol. 42, 844–851. Wolters, M., Garbutt, A., Bakker, J.P., 2005b. Salt-marsh restoration: evaluating the success of de-embankments in north-west Europe. Biol. Conserv. 123, 249–268. Wolters, M., Garbutt, A., Bakker, J.P., 2005c. Plant colonization after managed realignment: the relative importance of diaspore dispersal. J. Appl. Ecol. 42, 770–777.
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Wolters, M., Garbutt, A., Bakker, R., Bakker, J.P., Carey, P.D., 2006. Restoration of salt marsh vegetation in relation to site suitability, species pools and dispersal traits. In: Wolters, M. (Ed.), Restoration of Salt Marshes. Ph.D. Thesis, University of Groningen. Wood, N., Hine, A.C., 2007. Spatial trends in marsh sediment deposition within a microtidal creek system, Waccasassa Bay, Florida. J. Coast. Res. 23, 823–833. Woodell, S.R.J., 1985. Salinity and seed germination patterns in coastal plants. Vegetatio 61, 223–229. Zedler, J.B. (Ed.), 2001. Handbook for Restoring Tidal Wetlands. CRC Press, New York. Zedler, J.B., Lindig-Cisneros, R., 2000. Functional equivalency of restored and natural salt marshes. In: Zedler, J.B. (Ed.), Concepts and Controversies in Tidal Marsh Ecology. Kluwer Academic Publications, Dordrecht, pp. 565–582.
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C H A P T E R
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M ETHODS AND C RITERIA FOR S UCCESSFUL M ANGROVE F OREST R ESTORATION Roy R. Lewis III
Contents 1. Introduction 2. General Site Selection for Restoration 3. Specific Site Selection for Restoration 4. Establishing Success Criteria 5. Monitoring and Reporting Success 6. Functionality of Restored Mangrove Forests 7. Summary References
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1. INTRODUCTION Mangrove forest ecosystems currently cover 14.7 million hectares of the tropical shorelines of the world (Wilkie and Fortuna, 2003), which represents a decline from 19.8 million hectares in 1980 and 15.9 million hectares in 1990. These losses represent about 2% per year from 1980 to 1990 and 1% per year from 1990 to 2000. Therefore achieving no-net-loss of mangroves worldwide would require the successful restoration of approximately 150,000 ha/year, unless all major losses of mangroves ceased. Increasing the total area of mangroves worldwide would require an even larger scale effort. An example of documented losses includes combined losses in the Philippines, Thailand, Vietnam, and Malaysia of 7.4 million hectares of mangroves (Spalding, 1997). These figures emphasize the magnitude of the loss, and the magnitude of the opportunities that exist to restore areas like mosquito control impoundments in Florida (Brockmeyer et al., 1997) (several tens of thousands hectares) and abandoned shrimp aquaculture ponds in Southeast Asia (Stevenson et al., 1999) (several hundreds of thousands hectares), back to functional mangrove ecosystems. While great potential exists to reverse the loss of mangrove forests worldwide, most attempts to restore mangroves often fail completely, or fail to achieve the stated goals (Erftemeijer and Lewis, 2000; Lewis, 2000, 2005). Previously documented attempts to restore mangroves (Field, 1996, 1999) where successful, have Coastal Wetlands: An Integrated Ecosystem Approach
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largely concentrated on creation of plantations of mangroves consisting of just a few species, and targeted for harvesting as wood products (Kairo, 2002), or temporarily used to collect eroded soil and raise intertidal areas to usable terrestrial agricultural elevations (Saenger and Siddiqi, 1993). Successful mangrove forest restoration requires careful analyses of a number of factors in advance of attempting actual restoration. First for a given area of mangroves or former mangroves, the existing watershed needs to be defined, and any changes to the coastal plain hydrology that may have impacted the mangroves documented. Second, careful site selection must take place factoring into account the history of the site. Third, clearly stated goals and achievable and measurable success criteria need to be defined and incorporated into a proposed monitoring program. Fourth, the restoration methodology must reflect an acknowledgement of the history of routine failure in attempts at mangrove restoration and proposed use of proven successful techniques. Finally, after the initial restoration activities are complete, the proposed monitoring program must be initiated and used to determine if the project is achieving interim measurable success to indicate whether any mid-course corrections are needed.
2. GENERAL SITE SELECTION FOR R ESTORATION Previous research has documented the general principle that mangrove forests worldwide exist at the down slope end of coastal drainage basins (Kjerfve, 1990). At this down slope location adjacent to the sea, mangroves typically are established on a raised and sloped platform above mean sea level, and inundated approximately 30% or less of the time by tidal waters (Lewis, 2005). Kjerfve (1990) reported inundation times as short as 9% of the time for Klong Ngao on the west coast of Thailand. More frequent flooding causes stress and death of mangroves. Kjerfve (1990) suggests six key data needs prior to proceeding with looking at the hydrology of the basin and associated mangroves: 1. Size and extent of drainage basin 2. Extent and area of mangroves at the downslope (i.e., toward the sea) end of the basin 3. Topography and bathymetry of the mangrove areas 4. Hypsometric characteristics to calculate the current tidal prism of the mangrove areas 5. Rates of terrestrial input of water, sediment, and nutrients 6. Climatic water balance These analyses will yield one or more distinct areas of mangroves with varying characteristic hydrographic patterns, including more or less natural systems, some with natural damage like recent hurricanes, typhoons, or tsunamis; and those impacted by development activities, such as dredging and filling, channelized basin flows, road construction, and restrictions to tidal exchange particularly in lagoonal mangroves.
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3. S PECIFIC S ITE S ELECTION FOR R ESTORATION The general site selection process may be applied to more than one coastal drainage basin to yield a list of individual mangrove forest areas, and general characteristics of these areas. From this list, those areas showing either some current damage or significant declines in total area of mangroves from historical conditions need to be identified from a reconnaissance level examination of available maps and aerial or satellite photography. From this effort a short list of sites warranting further investigation is produced. Each of these potential restoration sites requires a field level investigation with maps and aerial or satellite photography in hand to verify vegetation signatures, including areas of stressed, dead, or lost mangroves. There may also be areas of damaged mangroves showing secondary succession or recovery from a previous damage event. The time frame since the damage event needs to be known in order to answer the key question, which is, “does this site need management to support further recovery, or accelerate recovery, or is it likely to recover over time by itself without intervention?” Or as Saenger (2002) emphasizes, “what is the history of the site or area, or more specifically, what prior activities have led to the present conditions?” Lewis (2005) has introduced the term “propagule limitation” to define a condition in which natural recovery is slowed or stalled due to a lack of natural mangrove propagules being available to volunteer at a damage site. Propagule limitation may be caused by a large loss of adult trees capable of producing propagules or by hydrologic restrictions or blockages (i.e., dikes), which prevent natural waterborne transport of mangrove propagules to a restoration site. Since propagules are produced at different times of the year by different species in different locations (Tomlinson, 1986), more than one site visit may be necessary in order to correctly identify a propagule limited site. Lack of propagules at a single time of year does not necessarily define a propagule limited site, and therefore careful evaluation of this parameter is important. If a damaged forest is going to recover on its own within an acceptable time frame, any attempt to introduce propagules or plant propagules or plant nursery grown mangroves is likely to be a waste of time and money. Recovery is here defined as the recolonization or planting of a restoration site and that the site’s growth of plant materials to some predefined numerical level (e.g., percent cover, total basal area).Priority should be given to sites that would indeed benefit from intervention by man given always limited time and money to devote to any restoration project. These suggestions may seem obvious, but the documentation of successful mangrove forest restoration is very limited, and more commonly former nonmangrove areas like mudflats or seagrass meadows seaward of natural mangroves or damaged areas without a properly documented history are the primary targets of well intentioned, but often faulty, mangrove restoration efforts (Field, 1996; Erftemeijer and Lewis, 2000; Lewis, 2005). The result of unsound evaluations of restoration opportunities has, unfortunately, emphasized first establishing a mangrove nursery, then planting mangroves at a casually selected site, as the primary
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tool in restoration, rather than first assessing the reasons for the loss of mangroves in an area and working with the natural recovery processes. Both Brockmeyer et al. (1997) and Stevenson et al. (1999) present examples of successful mangrove restoration following reestablishment of historical tidal connections to adjacent estuaries. This is termed “hydrologic restoration” (see discussion by Turner and Lewis, 1997). In the examples discussed, volunteer mangrove and mangrove nurse-plant propagules were sufficient to allow for rapid establishment of plant cover. No planting of mangroves was required. Lewis and Marshall (1997) suggest five critical steps are necessary to achieve ecological mangrove restoration (EMR), and these are discussed in more detail in Stevenson et al. (1999). The general approach is to emphasize hydrologic restoration opportunities without first emphasizing planting of mangroves (Turner and Lewis, 1997). These steps have been tested in training courses on mangrove restoration in the United States, Nigeria, Indonesia, Thailand, Vietnam, Sri Lanka, and India, and have been further modified with review and input by both teachers and students to include one added step as follows: 1. Understand the autecology (individual species ecology) of the mangrove species at the site, in particular the patterns of reproduction, propagule distribution, and successful seedling establishment. 2. Understand the normal hydrologic patterns that control the distribution and successful establishment and growth of targeted mangrove species. 3. Assess the modifications of the previous mangrove environment that occurred that currently prevent natural secondary succession. 4. Select appropriate mangrove restoration sites through application of Steps 1–3 above that are both likely to succeed in restoring a sustainable mangrove forest ecosystem, and are cost-effective given the available funds and manpower to carry out the projects, including adequate monitoring of their progress towards meeting quantitative goals established prior to restoration. This step includes resolving land ownership/use issues necessary for ensuring long-term access to and conservation of the site. 5. Design the restoration program at appropriate sites selected in Step 4 above to initially restore the appropriate hydrology and utilize natural volunteer mangrove propagule recruitment for plant establishment. 6. Only utilize actual planting of propagules, collected seedlings, or cultivated seedlings after determining through Steps 1–5 above that natural recruitment will not provide the quantity of successfully established seedlings, rate of stabilization, or rate of growth of saplings established as quantitative goals for the restoration project. Step number 6 is still the most controversial step of EMR. If natural recruitment fails, that may mean the site has not been adequately rehabilitated to facilitate volunteer mangrove recruitment where propagule limitation does not exist. For example, if the hydrology has not been adequately restored, or at an excavated site, the final topographic grade may be too high or too low. Under these circumstances, planting will not overcome these physical limitations on plant establishment, but planting does often occur and the plants then die.
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Local communities plant seedlings even after having undertaken successful EMR for a combination of three reasons: (1) lack of patience, (2) protection of the restoration site since planted areas appear to outsiders (not aware of the project) as an intentional action and provide a measure of protection for that area as it is obvious that there is human activity in the area, and (3) promotion of growth of preferred species such as Rhizophora over colonizers such as Avicennia or Sonneratia. Even with adequate local mangrove recruitment after EMR, planting may serve as an educational process for local communities and encourage local support for the project. It is however still important to document natural recruitment during monitoring and report the differential contribution of volunteer- and humanplanted mangroves to the final estimate of plant cover. Through the application of these six simple steps, and basic principles of ecological restoration using ecological engineering approaches, including careful cost evaluations prior to final selection of a restoration site and design of a restoration program, the opportunity for a cost-effective and successful restoration effort is maximized. Most of the largest attempts to restore mangroves are currently taking place in Southeast Asia. While the six-step process described above has been taught as several short courses and is widely publicized in English in this part of the world, more education needs to be done in native languages, and published in native languages. The lack of large-scale translations of scientific papers about mangrove restoration into local languages is hampering adoption of the six-step process. Also, the lack of general application of the rule of law in several of these countries limits attempts to protect existing mangroves. Large-scale conversion of existing mangrove forests to aquaculture ponds is still taking place, and more recently, conversion of mangroves to what are perceived as more valuable oil palm plantations has accelerated. Ecologists and engineers have not understood mangrove hydrology, as Kjerfve (1990) points out. Although a number of papers discuss the science of mangrove hydrology (Kjerfve, 1990; Wolanski et al., 1992; Furukawa et al., 1997, see also the review in Mazda et al., 2007), their focus has been on tidal and freshwater flows within the forests, and not the critical periods of inundation and dryness that govern the health of the forest. Kjerfve (1990) does discuss the importance of topography and argues that “. . . micro-topography controls the distribution of mangroves, and physical processes play a dominant role in formation and functional maintenance of mangrove ecosystems . . .” Hypersalinty due to year-to-year variations in rainfall can produce natural mangrove diebacks (Cintron et al., 1978), and disruption of normal freshwater flows that dilute seawater in more arid areas can kill mangroves (Perdomo et al., 1998; Medina et al., 2001), or stress mangroves to the point that they may not be able to keep up with sea-level rise through root production and the laying down of a peat layer (Snedaker, 1993). The point of all of this is that flooding depth, duration, and frequency are critical factors in the survival of both mangrove seedlings and mature trees. Once established, mangroves can be further stressed if the tidal hydrology is changed, for example by diking (Brockmeyer et al., 1997). Both increased salinity due to reductions in freshwater availability, and flooding stress, increased anaerobic
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conditions, and free sulfide availability can kill or stress existing stands of mangroves, and mangroves at restoration sites (McKee, 1993). For these reasons, any engineering works constructed near mangrove forests, or in the watershed that drains to mangrove forests, must be designed to allow for sufficient free exchange of seawater with the adjacent ocean or estuary, and not interrupt essential upland or riverine drainage into the mangrove forest. Failure to properly account for these essential inputs and exchange of water will result in stress and possible death of the forest. Engineering works such as dikes created to make shrimp aquaculture ponds disrupt the natural hydrology and produce conditions that prevent natural recovery once these ponds are abandoned due to disease (Stevenson et al., 1999). Use a reference mangrove site for examining normal hydrology for mangroves in your particular area. Either install tide gauges and measure the tidal hydrology of a reference mangrove forest, or use the surveyed elevation of a reference mangrove forest floor as a surrogate for hydrology, and establish those same ranges of elevations at your restoration site, or restore the same hydrology to an impounded mangrove by breaching the dikes in the right places. The “right places” are usually the mouths of historic tidal creeks. These are often visible in vertical (preferred) or oblique aerial photographs.
4. ESTABLISHING SUCCESS C RITERIA Once a site is finally chosen for restoration, and a design developed, quantifiable success criteria should be established. Establishing such criteria is important to actually measure progress toward successful restoration. The first step in establishing numeric criteria for success is to prepare a brief narrative goal or set an objective (Saenger, 2002) for the project. This will define the next steps. A goal may be to establish a monotypic plantation of Rhizophora apiculata to be harvested after 12 years as poles. This is a typical goal of many mangrove restoration projects. It may be an acceptable goal to local stakeholders in the project such as local villages and fishermen, and harvest of wood products from locally managed forests is a typical goal (see discussion of timber production in the Matang Forest, Malaysia, by Saenger, 2002, pp. 231–234). If on the other hand the goal is to provide fish and invertebrate habitat to restore local fisheries, a different approach to establishing success criteria is dictated. Maximizing such habitat use usually means maximizing biodiversity of the plant species, and therefore a monotypic stand of mangroves in an area that normally supports 20 or more mangrove tree species is not a logical goal. Establishment and persistence of tidal creeks to assist with entry and exit of juvenile and adult fish and invertebrates may also be needed criteria. Once narrative success criteria are agreed upon, quantitative criteria need to then be established. For the first example above, a certain number of pole size trees per hectare could be established. Such a success criteria would also likely dictate an immediate planting program of collected propagule or nursery grown plants, rather
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than waiting for volunteer propagules. For the second example, to maximize biodiversity, the restoration site might be left alone and not planted immediately to allow for volunteer colonization of the largest number of different species of mangroves from propagules produced by trees adjacent to a restoration site. The next step is to look at available information on both plantation and natural recruitment indices of success. Saenger (2002, pp. 256–270) discusses in great detail what is to be expected in terms of biomass and stem density for example from typical plantation projects. There has been much work on plantation projects where just a few species of mangroves are managed for, and thus there is a wealth of data to examine. In contrast with this, the availability of data on natural recruitment within a mixed forest is generally not available. McKee and Faulkner (2000) report on the results of sampling for density and basal area within two restored mangrove forests in Florida, USA, and compared these to two adjacent control areas. Their data show density and basal area for volunteer mangroves in the restoration areas exceeded that for planted mangroves. Proffitt and Devlin (2005) report similar results from one of the same sites sampled by McKee and Faulkner (2000) but in later years as the system matured. Lewis et al. (2005) report on the results of cover sampling over a period of 5 years within a restored mangrove forest in another location in Florida, USA. These studies help define parameters that need to be sampled and sampling methodologies, but provide limited data to apply to local situations in other parts of the world.
5. MONITORING AND REPORTING SUCCESS Once success criteria have been established, and the site restored through hydrologic restoration with or without planting, monitoring, and reporting should begin. A typical monitoring schedule would consist of the following 10 reports: 1. 2. 3. 4. 5. 6. 7. 8. 9. 10.
Time Zero 0 þ 3 months 0 þ 6 months 0 þ 9 months 0 þ 12 months 0 þ 18 months 0 þ 24 months 0 þ 36 months 0 þ 48 months 0 þ 60 months
A Time Zero report is prepared after all the site restoration changes have been accomplished and any proposed planting completed. It should include photographs taken from fixed stations where future photography will also be taken. The shorter intervals in the early years of monitoring are designed to insure that any corrective actions necessary due to problems encountered during monitoring are quickly corrected. These are termed “mid-course corrections.”
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This is a typical schedule of reporting as required by wetland mitigation projects in the United States. As noted by the data of Proffitt and Devlin (2005), however, changes in the height, density, and species composition of mangroves on a restoration site will continue over time. Eighteen years have passed since the restoration of the site described in Proffitt and Devlin (2005), and changes are still being observed. For example, the dominant species present 7 years after restoration, Laguncularia racemosa, experienced “reduced recruitment and apparent density-dependent mortality” through the last sampling 18 years after restoration. Longer term monitoring in forested restoration areas is recommended where resources are available. The recommended time interval after the last regular monitoring would be 5-year intervals until the time of maturity of the restored forest. Based on the work of Lewis et al. (2005) total cover by mangroves can be expected to occur rapidly (within 3–5 years), but basal area equivalency will take much longer. Ten years after hydrologic restoration Stevenson et al. (1999) report that the total basal area for all species of mangroves at hydrologically restored site in Costa Rica (abandoned shrimp aquaculture pond) was 64.2% of that of the control site. Eighteen years after restoration, Proffitt and Devlin (2005) report the total basal area of all mangrove species (42.7 m2/ha) exceeded that of the mean of the two control areas (17.9 m2/ha) by a factor of 2.4. They make an important note however in that they encountered a large number of saplings exceeding the 1.3 m height requirement for counting in basal area measurements, but they were less than the minimum of 2.0 cm in DBH (diameter at breast height). Normally these trees are not counted in basal area calculations (Cintron and Novelli, 1984). We believe ignoring them produces a total basal area calculation that does not represent the true value for all trees on the restoration site and recommend they do be counted and DBHs measured, but that category of trees less than 2 cm in DBH be reported separately from the other normally reported classes (i.e., 2, <10, and 10 cm) to allow for direct comparisons to other data sets like that of Stevenson et al. (1999) and McKee and Faulkner (2000). We also recommend the establishment of either permanent, haphazard or random plots of 5 m 5 m, at a density within the restoration area that allows for stratified sampling over the elevation gradient present within the restoration site. The Point Centered Quarter Method (Cintron and Novelli, 1984) can also be used.
6. F UNCTIONALITY OF R ESTORED MANGROVE FORESTS Lewis (2005) notes that ecological restoration of mangroves, where restored ecosystem functions are the goal, is rarely a prime goal of restoration projects, and thus is rarely monitored. Lewis (1992) notes that research on the use by fish of both restored tidal marshes and mangroves in the United States shows that fish populations in these restored plant communities are equivalent in both numbers and species composition within 3–5 years of restoration. McKee and Faulkner (2000) examined the biogeochemical functions of two restored mangrove forests in Florida (6 and 14 years old). Soil Eh was lower at the
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restoration sites and pore water sulfide concentrations were significantly higher. Soil carbon and nitrogen levels were greater overall in natural soils and were correlated with soil organic matter content. They concluded that site-specific parameters such as rates of tidal flushing, topography, and salinity played a larger role in primary production and turnover rates of organic material than site age. Bosire et al. (2008) provide a thorough review of faunal use of restored mangrove forests, including data for restoration projects in six countries (USA, Thailand, Kenya, Philippines, Qatar, and Malaysia). The data collected for these sites was very variable however, and no uniform sampling methodology or target faunal groups were the focus of the scattered research efforts. As would be expected, the conclusion of the authors was that functionality depended on what parameters were measured at a given location and generalities were very difficult to make. For example in Thailand, crab diversity at some of the restored sites was higher than at an upper shore natural mangrove site, and both biomass and crab numbers were consistently higher in the restored sites (Macintosh et al., 2002). However, the natural site was characterized by large numbers of sesarmid crabs. Differences in the crab diversity in Thailand were thought to relate to, among other things, inundation zone and differences in the mangrove species present in the restored sites (Macintosh et al., 2002). However, in Qatar, Al-Khayat and Jones (1999) found lower species richness of crabs at restored sites compared to natural habitats of Avicennia marina. In Kenya, reforested stands of Rhizophora mucronata and A. marina had higher crab densities than their natural references (Bosire et al., 2004) but with similar species diversity and crab species composition compared to respective bare control with similar site history. In the Philippines the relative abundance of mud crab (Scylla olivacea) to two other noncommercial species was used to separate the effects of habitat from fishing pressure and recruitment limitation. A comparison of mud crab (Scylla olivacea) populations in restored, natural, and degraded sites in the Philippines suggested that 16-year-old restored Rhizophora spp. can support densities of mud crabs equivalent to that of natural mixed species mangroves (Walton et al., 2007). Mollusk diversity showed similar patterns to that of crabs in both previously mentioned studies in Qatar and Thailand, while in Kenya, no mollusks were observed in a bare site while the restored site and natural reference site within a Sonneratia alba forest had similar species composition, density, and diversity. Infauna communities (i.e., polycheate worms, amphipods) showed patterns similar to those described above. Lower diversity of taxa was observed in restored versus natural sites in Qatar with the data being less clear in Thailand. Studies in Malaysia suggested that 2-year-old restored mangroves had the greatest biomass and species number followed by the control and 15-year-old stand, although species diversity was highest in the control site and lowest in the 2-year-old site (Sasekumar and Chong, 1998). In Kenya, bare sites had the lowest infauna densities and taxa richness compared to restored sites with natural reference sites having the highest densities. Taxa richness and composition were similar among respective restored and natural sites (Bosire et al., 2004), suggesting successful fauna recolonization following mangrove restoration. Studies of mobile fauna in restored mangroves of varying ages and species composition showed variable patterns. In Qatar, lower diversity of both juvenile and adult fish was observed in restored sites compared to natural stands of A. marina
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(Al-Khayat and Jones, 1999). Studies comparing fish and shrimp density between natural stands of R. apiculata, Avicennia officinalis, A. marina, and a single restored R. apiculata stand (5–6 years old) in the Philippines indicated that density and biomass were primarily influenced by tidal height and mangrove species (Ro¨nnba¨ck et al., 1999). In S. alba plantations in Kenya, there were strong seasonal fluctuations of juvenile fish, showing temporal patterns to be a potentially stronger influence on fish assemblages than type of restoration site or presence of fringing mangroves (Crona and Ro¨nnba¨ck, 2007). However, the spatial scale of observation is likely a much stronger factor affecting biodiversity studies of more mobile fauna compared to less mobile animal communities described above. Since most studied restoration sites are small in size, site-specific effects on fish distribution patterns remain largely unknown. The same is true for juvenile shrimp. In Kenya one species, Penaeus japonicus, dominated the community, and lower species richness was observed in a restored area of S. alba, than in adjacent clear felled areas (Crona and Ro¨nnba¨ck, 2005). Natural forests had higher root complexity and also higher abundances and more even distribution of shrimp species in terms of species composition. Similarly, in the Philippines, higher abundances of juvenile shrimp in a restored R. apiculata site were seemingly related to higher structural root complexity, although more inland stands of mature Avicennia spp. and Rhizophora spp. showed no such differences (Ro¨nnba¨ck et al., 1999). Lewis and Gilmore (2007) discuss fish use of both natural and restored mangrove forests and report specifically about monitoring of a successful 500 ha mangrove restoration project in Hollywood, Florida, USA, where sampled fish populations in both reference and restored sites were statistically indistinguishable within 3–5 years of restoration. They emphasize three restoration and design goals to ensure functional ecological restoration of mangrove forests: 1. Achieve plant cover similar to that in an adjacent relatively undisturbed control area of mangrove forest. 2. Establish a network of channels that mimic the shape and form of a natural tidal creek system. 3. Establish a heterogeneous landscape similar to that exhibited by local mangrove ecosystems. Few studies exist on trends in biodiversity in restored mangroves, and the range in age, species, and inundation class of restored sites makes generalizations difficult. However, the co-occurrence of many animal species in both restored and comparable natural forests suggest colonization of restoration sites by both mobile and nonmobile fauna is a rapid process, and equivalent populations of mangrove fauna in both natural controls and restored mangrove sites can typically be found within 5–10 years of restoration.
7. SUMMARY Mangrove forest restoration has not been generally successful except where timber production was the goal and monotypic stands were established. Ecological
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restoration of mangrove forests, where the goal is the restoration of a mixed species forest cover and functions equivalent to that of an adjacent reference forest has not typically been a design criteria, and most restoration projects not targeting timber production, but with some general ecological goals have not been successful (Erftemeijer and Lewis, 2000; Lewis, 2005). The chosen restoration sites for many of these projects have been mudflats or seagrass beds lying seaward of the outer edge of existing mangrove forests. These sites are typically planted with nursery grown mangrove seedlings which do not survive due to frequent inundation. Although there are relatively few studies on trends in biodiversity in restored mangroves, the co-occurrence of many animal species in both restored and comparable natural forests suggest colonization of restoration sites by both mobile and nonmobile fauna is a rapid process, with equivalent populations of mangrove fauna in both natural control areas and restored mangrove sites typically found within 5–10 years of restoration. The scientific basis for optimum design of restoration projects to meet certain established criteria, such as increased fish production or more use by wading seabirds, is however very minimal. In the future mangrove restoration projects should be more carefully designed to ensure successful establishment of plant cover at minimal cost over large areas. This can be achieved for example by restoring hydrologic connections to impounded mangrove areas as has been done in Florida (Brockmeyer et al., 1997), Costa Rica, and the Philippines (Stevenson et al., 1999). Funding agencies typically fund mangrove restoration projects with minimal funds dedicated towards quantitative monitoring and reporting over a reasonable and ecologically based time period (5 years minimum). Both failures and successes thus go undocumented, and mistakes are repeated and lessons learned are lost. Funding agencies and governments need to realize that large amounts of limited restoration funds are now being wasted because of these shortsighted efforts, and at a minimum they should regularly review, publish, and teach the lessons learned from both past successes and failures. These same funding agencies and governments are very loath to fund careful examination of the ecological functions of restored mangrove areas. This is somewhat understandable given the large costs of quantitative monitoring, but at a minimum attempts should be made to coordinate restoration projects with graduate training programs to provide, at likely minimal costs, opportunities to graduate students and researchers alike access to restoration sites where good research can be done for minimal costs. These efforts to date have been hampered however by the lack of application of uniform methodologies of sampling and reporting. We have made some recommendations above, but would encourage every researcher to revisit and read carefully the excellent work of Snedaker and Snedaker (1984), where uniform scientific methodologies developed by worldwide consensus for sampling flora, fauna, biochemistry, litter production, and decomposition, among other parameters, were recommended. In addition, the additional detailed work on mangrove forest hydrology of Kjerfve (1990) and Mazda et al. (2007) are essential to round out the beginning point for any mangrove restoration research program. We find most current researchers do not start with these
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references and then add appropriate additional scientific publications to the mix to develop a very well thought out and reviewed plan of study. The result can be data generation of little value in promoting the advancement of the science of ecological restoration of mangrove forests. At the present time, we believe that greater success at EMR could be achieved with a four-step approach that includes: 1. General site selection for restoration sites that includes examination of multiple coastal basins that contain mangroves. 2. Specific site selection that looks at the history of changes in areal cover of mangroves and changes in hydrology at specific potential restoration sites, and targets hydrologic restoration of these sites. 3. Establishing quantitative and measurable success criteria and use uniform criteria between study sites. 4. Monitoring and reporting of progress toward achieving these success criteria, including reporting on lessons learned from both successes and failures.
REFERENCES Al-Khayat, J.A., Jones, D.A., 1999. A comparison of the macrofauna of natural and replanted mangroves in Qatar. Estuar. Coast. Shelf Sci. 49, 55–63. Bosire, J.O., Dahdough-Guebas, F., Walton, M., Crona, B.I., Lewis III, R.R., Field, C., Kairo, J.G., Koedam, N., 2008. Functionality of restored mangroves: A review. Aquatic Botany 89, 251–259. Bosire, J.O., Dahdouh-Guebas, F., Kairo, J.G., Cannicci, S., Koedam, N., 2004. Spatial macrobenthic variations in a tropical mangrove Bay. Biodiversity and Conserv. 13, 1059–1074. Brockmeyer Jr., R.E., Rey, J.R., Virnstein, R.W., Gilmore, R.G., Ernest, L., 1997. Rehabilitation of impounded estuarine wetlands by hydrologic reconnection to the Indian River Lagoon, Florida (USA). Wetl. Ecol. Manage. 4, 93–109. Cintron, G., Lugo, A.E., Pool, D.J., Morris, G., 1978. Mangroves and arid environments in Puerto Rico and adjacent islands. Biotropica 10, 110–121. Cintron, G., Novelli, Y., 1984. Methods for studying mangrove structure. In: Snedaker, S.C., Snedaker, J.G. (Eds.), The Mangrove Ecosystem: Research Methods. Monographs in Oceanographic Methodology 8. UNESCO, Paris, pp. 91–113. Crona, B.I., Ro¨nnba¨ck, P., 2005. Use of replanted mangroves as nursery grounds by shrimp communities in Gazi Bay, Kenya. Estuar. Coast. Shelf Sci. 65, 535–544. Crona, B.I., Ro¨nnba¨ck, P., 2007. Community structure and temporal variability of juvenile fish assemblages in natural and replanted mangroves, Sonneratia alba Sm., of Gazi Bay, Kenya. Estuar. Coast. Shelf Sci. 74 (1), 44–52. Erftemeijer, P.L.A., Lewis, R.R., 2000. Planting mangroves on intertidal mudflats: habitat restoration or habitat conversion? In: Proceedings of the ECOTONE VIII Seminar Enhancing Coastal Ecosystems Restoration for the 21st Century, Ranong, 23–28 May 1999. Royal Forest Department of Thailand, Bangkok, pp. 156–165. Field, C.D. (Ed.), 1996. Restoration of Mangrove Ecosystems. International Society for Mangrove Ecosystems, Okinawa, 250pp. Field, C.D., 1999. Rehabilitation of mangrove ecosystems: an overview. Mar. Pollut. Bull. 37, 383–392. Furukawa, K.E., Wolanski, E., Mueller, H., 1997. Currents and sediment transport in mangrove forests. Estuar. Coast. Shelf Sci. 44, 301–310. Kairo, J.G., 2002. Regeneration status of mangrove forests in Mida Creek, Kenya: a compromised or secured future? Ambio 31, 562–568.
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Kjerfve, B., 1990. Manual for investigation of hydrological processes in mangrove ecosystems. UNESCO/UNDP Regional Project, Research and its Application to the Management of the Mangroves of Asia and the Pacific (RAS/86/120), 79pp. Lewis, R.R., 1992. Coastal habitat restoration as a fishery management tool. In: Stroud, R.H. (Ed.), Stemming the Tide of Coastal Fish Habitat Loss. Proceedings of a Symposium on Conservation of Coastal Fish Habitat, Baltimore, MD, 7–9 March 1991. National Coalition for Marine Conservation, Inc., Savannah, GA, pp. 169–173. Lewis, R.R., 2000. Ecologically based goal setting in mangrove forest and tidal marsh restoration in Florida. Ecol. Eng. 15, 191–198. Lewis, R.R., 2005. Ecological engineering for successful management and restoration of mangrove forests. Ecol. Eng. 24, 403–418. Lewis, R.R., Marshall, M.J., 1997. Principles of successful restoration of shrimp Aquaculture ponds back to mangrove forests. Programa/resumes de Marcuba ’97, Habana, Cuba, 124pp. Lewis, R.R., Hodgson, A.B., Mauseth, G.S., 2005. Project facilitates the natural reseeding of mangrove forests (Florida). Ecol. Restor. 23, 276–277. Lewis, R.R., Gilmore, R.G., 2007. Important conside´rations to achieve successful mangrove forest restoration with optimum fish habitat. Bull. Mar. Sci. 80, 823–837. Macintosh, D.J., Ashton, E.C., Havanon, S., 2002. Mangrove rehabilitation and intertidal biodiversity: a study in the Ranong mangrove ecosystem, Thailand. Estuar. Coast. Shelf Sci. 55, 331–345. Mazda, Y., Wolanski, E., Ridd, P.V., 2007. The Role of Physical Processes in Mangrove Environments. Terrapub, Tokyo. McKee, K.L., 1993. Soil physiochemical patterns and mangrove species distribution – reciprocal effects? J. Ecol. 81, 477–487. McKee, K.L., Faulkner, P.L., 2000. Restoration of biogeochemical function in mangrove forests. Restor. Ecol. 8, 47–259. Medina, E., Fonseca, H., Barboza, F., Francisco, M., 2001. Natural and man-induced changes in a tidal channel mangroves system under tropical semiarid climate at the entrance to the Maracaibo Lake (Western Venezuela). Wetl. Ecol. Manage. 9, 233–243. Perdomo, L, Ensminger, I. Espinosa, L.F., Elster, C., Wallner-Kersanach, M., Schnetter, M-L., 1998. The mangrove ecosystem of Cienaga Grande de Santa Marta (Colombia): observations on regeneration and trace metals in sediment. Mar. Pollut. Bull. 37, 393–403. Proffitt, C.E., Devlin, D.J., 2005. Long-term growth and succession in restored and natural mangrove forests in southwestern Florida. Wetl. Ecol. Manage. 13, 531–551. Ro¨nnba¨ck, P., Troell, M., Kautsky, N., Primavera, J.H., 1999. Distribution patterns of shrimps and fish among Avicennia and Rhizophora microhabitats in the Pagbilao mangroves, Philippines. Estuar. Coast. Shelf Sci. 48, 223–234. Saenger, P., 2002. Mangrove Ecology, Silviculture and Conservation. Kluwer Academic Publishers, Dordrecht, 359pp. Saenger, P., Siddiqi, N.A., 1993. Land from the seas: the mangrove afforestation program of Bangladesh. Ocean Coast. Manage. 20, 23–39. Sasekumar, A., Chong, V.C., 1998. Faunal diversity in Malaysian mangroves. Global Ecol. Biogeogr. Lett. 7, 57–60. Snedaker, S.C., Snedaker, J.G., 1984. The Mangrove Ecosystem: Research Methods. Monographs on Oceanographic Methodology 8. UNESCO, Paris, 251pp. Snedaker, S.C., 1993. Impact on mangroves. In: Maul, G.A. (Ed.), Climatic Changes in the IntraAmerican Seas: Implications of Future Climate Change on the Ecosystems and Socio-Economic Structure of the Marine and Coastal Regimes of the Caribbean Sea, Gulf of Mexico, Bahamas and N.E. Coast of S. America. Edward Arnold, London, pp. 282–305. Spalding, M.D., 1997. The global distribution and status of mangrove ecosystems. Intercoast Network Newsletter Special Edition #1, 20–21. Stevenson, N.J., Lewis, R.R., Burbridge, P.R., 1999. Disused shrimp ponds and mangrove rehabilitation. In: Streever, W.J. (Ed.), An International Perspective on Wetland Rehabilitation. Kluwer Academic Publishers, Dordrecht, pp. 277–297. Tomlinson, P.B., 1986. The Botany of Mangroves. Cambridge Tropical Biology Series. Cambridge University Press, New York, 413pp.
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Turner, R.E., Lewis, R.R., 1997. Hydrologic restoration of coastal wetlands. Wetlands. Ecol. Manage. 4, 65–72. Walton, M.E.M., LeVay, L., Lebata, J.H., Binas, J., Primavera, J.H., 2007. Assessment of the effectiveness of mangrove rehabilitation using exploited and non-exploited indicator species. Biol. Conserv. 138, 180–188. Wilkie, M.L., Fortuna, S., 2003. Status and trends in mangrove area extent worldwide. Forest Resources Assessment Working Paper 63. Forestry Department, Food and Agriculture Organization of the United Nations, Rome, 36pp. Wolanski, E., Mazda, Y., Ridd, P.V., 1992. Mangrove hydrodynamics. In: Robertson, A.I., Alongi, D.M. (Eds.), Tropical Mangrove Ecosystems. American Geophysical Union, Washington, DC, pp. 436–462.
C H A P T E R
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E VALUATION OF R ESTORED T IDAL F RESHWATER W ETLANDS Andrew H. Baldwin, Richard S. Hammerschlag, and Donald R. Cahoon
Contents 1. 2. 3. 4. 5.
Introduction Characteristics of Restored TFW Success Evaluation of Restored Wetlands Ecosystem Attributes Measured at Restored TFW Criteria for Successful Restoration of TFW 5.1. Hydrologic criteria 5.2. Geomorphological criteria 5.3. Soil criteria 5.4. Salinity criteria 5.5. Vegetation criteria 5.6. Seed bank criteria 5.7. Benthic invertebrate criteria 5.8. Fish and wildlife criteria 6. Case Study: Evaluation of Restored TFW of the Anacostia River, Washington, DC, USA 6.1. Characteristics of restored and reference sites 6.2. Evaluation of success of restored TFW 7. Conclusions and Recommendations 7.1. Restoration of TFW in urban landscapes and selection of urban reference sites 7.2. Establishment of vegetation 7.3. Control of nonnative species 7.4. Implications for restoration of TFW Acknowledgements References
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1. INTRODUCTION Tidal freshwater wetlands (TFW) are recognized as highly productive coastal wetlands that support diverse assemblages of plants and animals and complex biogeochemical cycles (Whigham et al., 2009; Megonigal and Neubauer, 2009). Many TFW and their associated ecosystem services have been damaged or destroyed by urbanization, agriculture, and other human activities (Baldwin, 2004; Barendregt et al., 2006). Coastal Wetlands: An Integrated Ecosystem Approach
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Increasing recognition of the value of remaining wetlands and environmental regulations requiring wetland mitigation (i.e., enhancement, creation, or restoration of wetlands to compensate for wetland losses; Kentula, 2000) has driven the restoration of all types of wetlands, including TFW. While considerable information is available in the scientific literature on restoration of saline and brackish tidal wetlands and non-tidal wetlands, little information on restoration of TFW has been available in the refereed literature until recently. Furthermore, criteria for successful restoration of TFW have not been described in the literature. In this chapter, we first describe the distribution of restored TFW in the United States and Canada, their general construction methods, and motivating factors for restoration (Section 2). Then we review literature on general approaches to restoration success evaluation across wetland types (Section 3). In Section 4 we summarize the specific ecosystem attributes and methods that have been monitored to evaluate restored TFW and, based on Sections 3 and 4, in Section 5 we propose criteria for evaluating the success of TFW restoration projects. These criteria are subsequently used to evaluate restored TFW in the Anacostia River watershed in Washington, DC (Section 6). Finally, we present conclusions and recommendations to increase the successful restoration of TFW (Section 7). Throughout this chapter, the terms “restored” and “restoration” are used to identify efforts that have been made to reestablish or construct wetlands in places where they occurred formerly but were lost or degraded due to human activities or natural processes. These terms are not meant to imply that the wetlands were completely restored to a predisturbance ecosystem condition.
2. CHARACTERISTICS OF R ESTORED TFW The distribution and landscape position of restored TFW roughly parallels that of naturally occurring TFW (see Megonigal and Neubauer, 2009; Leck et al., 2009). In the United States, TFW have been restored in the estuaries of the Pacific Northwest (Washington and Oregon), the Sacramento–San Joaquin Delta (California), the midAtlantic coast (Chesapeake Bay and Delaware Bay), and the southeastern United States (North Carolina) (Baldwin, 2009). Just along the Anacostia River in and near Washington, DC, seven separate TFW restoration projects are in place. Furthermore, large-scale restoration is ongoing in the delta plain of the Mississippi River in coastal Louisiana. Many of the restored wetlands in Louisiana can be classified as TFW even though they are not often referred to as such, presumably due to their microtidal hydrologic regime and deltaic geomorphology. On the Fraser River in British Columbia, numerous small TFW projects have been implemented (Levings and Nishimura, 1997; Adams and Williams, 2004). Restoration projects also likely exist in eastern Canada (see the Atlas of Bank Restoration Sites of the St. Lawrence River, www.qc.ec.gc.ca/faune/ AtlasDeRestaurationDesRivesDuSaint-Laurent), although we have found no descriptions of specific TFW restoration projects there. TFW have been restored to mitigate for losses of TFW due to development (roads, buildings), stream channelization, and hydrologic alteration, or to replace ecosystem functions of lost wetlands (Table 1). In the US Pacific Northwest and the
Motivating factors for restoration and construction methods for TFW restoration projects in the United States and Canada
Region
Motivating factors for restoration
Construction methods
References
US Mid-Atlantic
Mitigate road, airport, bridge construction losses; channel maintenance and dredge material disposal; ecosystem and habitat restoration
Excavation of uplands; raising elevation with dredged sediment; cutting tidal channels
US Southeast Atlantic
Mitigate road construction impacts
US Pacific: Sacramento–San Joaquin Delta, California US Pacific Northwest
Ecosystem restoration
Grading, recontouring, cutting tidal channels, breaching berm along river Breach levees surrounding delta islands; excavation to reduce elevation
Kaminsky and Scelsi (1986), Bartoldus and Heliotus (1989), Bartoldus (1990), Bowers (1995), Syphax and Hammerschlag (1995), Marble and Company (1998), Baldwin and DeRico (1999), Gannett Fleming (2001), Quigley (2001), Neff (2002), Hammerschlag et al. (2006) Land Management Group (2004) Simenstad et al. (2000), Orr et al. (2003), Stillwater Environmental Services (2003), Phillip (2005)
Levee breaching; dike removal; excavation of fill material
Simenstad and Thom (1996), Gray et al. (2002), Tanner et al. (2002)
Hydrologic restoration; shoreline protection; freshwater diversion; dredge material placement
Lane et al. (1999), Louisiana Coastal Wetlands Conservation and Restoration Task Force, www.lacoast.gov Kistritz (1996), Levings and Nishimura (1996, 1997), Grout et al. (1997)
US Gulf of Mexico: Louisiana Canada: Fraser River, British Columbia
Mitigation for development; ecosystem restoration; fisheries habitat restoration Reduce marsh erosion and inundation; reduce salinity; increase land-water ratio; promote natural delta development Ecosystem restoration for fisheries
Dredge spoil placement; transplanting of vegetation plugs to barren sites
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Table 1
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Fraser River in British Columbia, Canada, restoring habitat for salmoniddominated fisheries is a primary motivation for restoration of TFW. Extensive restoration of TFW (and more saline wetland types) in the Mississippi River delta plain in Louisiana aims to: (1) stabilize or reverse loss of wetlands due to erosion and increasing relative sea level; (2) increase the ratio of land to open water; (3) reduce saltwater intrusion; and (4) promote the development of natural delta features. Restoration of many TFW involves excavation of upland soils or placement of dredged sediment in open water areas to create a substrate suitable for plant growth at an intertidal elevation similar to those of naturally or previously occurring TFW (Table 1). In locations where former wetlands were surrounded by dikes or levees to dewater them for agriculture, such as the deltas of the Sacramento and San Joaquin rivers in California and some rivers of the Pacific Northwest, breaching of levees and dikes is a common method. Because of the large area of wetlands requiring restoration in the Mississippi River Delta, restoration techniques include diverting sediment-laden river water into deteriorating areas, protecting shorelines against erosion, restoring historical hydrologic connections, and adding dredge material to increase elevation. Most TFW projects have focused on creating wetlands dominated by herbaceous plants (“marshes”) (Figure 1). However, at least one project, the McIntyre Tract associated with the Cape Fear River in North Carolina, has included planted trees and shrubs as well as herbaceous plants with the goal of restoring TFW cypress–gum swamp and tidal marsh/scrub–shrub habitat (Land Management Group, 2004).
Figure 1 Dense and species-rich vegetation at the Duck Island restoredTFW, Delaware River, Hamilton Township, New Jersey. Purple loosestrife is visible in clumps in the foreground. Detailed descriptions are included in Leck (2003) and Leck and Leck (2005). Photo by A.H. Baldwin.
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3. S UCCESS EVALUATION OF R ESTORED W ETLANDS Evaluating the success of wetland restoration projects is among the most important goals of restoration ecology and regulatory enforcement, yet success has proven to be an elusive quantity in North America. In this section we review general approaches that have been used to evaluate the success of wetland restoration projects across different types of wetlands, and summarize success evaluation efforts specifically at restored TFW. Success evaluation can be divided into three broad categories: compliance, functional, and landscape success (Wilson and Mitsch, 1996; Kentula, 2000). Compliance, or legal, success is determined by comparing the restored site with what was specified in the permit (for example as in a Section 404 permit), or by comparing the restored wetland with the wetland that was lost. The most common measures of compliance success are size, vegetation cover, and use by wildlife (Wilson and Mitsch, 1996), which are relatively straightforward to assess. In contrast, determination of functional, or ecosystem, success depends on a comparison of functional and structural attributes of the restored wetland with reference wetlands, aspects that can be considerably more difficult to evaluate. Measures of ecosystem function include hydrologic, biogeochemical, and plant and animal community characteristics (Brinson and Rheinhardt, 1996) that can be measured directly or indirectly through structural indicators of function. Examples of methods used in assessing wetland ecosystem function and structure include measures of water level, soil texture and chemical composition, nutrient cycling, microbial processes, geomorphology, fish and wildlife, invertebrates, vegetation composition and production, and water quality (Mitsch and Wilson, 1996; Wilson and Mitsch, 1996; Zedler and Callaway, 1999; Craft et al., 2003). Other approaches include assessment of ecosystem “services” that provide socioeconomic benefits such as recreation, aesthetics, flood control, shoreline protection, and education (Bartoldus, 1999). These functional or structural measures are compared with reference sites using approaches such as “functional trajectories,” indices of function (e.g., hydrogeomorphic method), or direct comparisons of restored and reference sites (Simenstad and Thom, 1996; Zedler and Callaway, 1999; Craft et al., 2003). Compliance and functional success are necessarily evaluated for specific restoration sites. Increasingly, however, the importance of evaluating the success of the restored wetland in contributing to the hydrological and ecological integrity of the watershed, region, or landscape has been stressed (Kentula, 2000; NRC, 2001). This landscape success category requires evaluating restored sites within the context of their surrounding watershed and how they contribute to broad restoration of ecosystem function such as maintenance of regional biological diversity and watershed hydrology. Furthermore, landscape position strongly influences geomorphology and hydrology and therefore is critical to establishing wetlands that will persist indefinitely, that is, that are sustainable (NRC, 2001). Landscape also becomes an overriding concern in selection of appropriate reference sites or standards for comparison with restored sites. Because human disturbance or development in the watershed surrounding restored sites has a strong
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effect on wetland structure and function, reference standards that are higher than those that can be sustained in the landscape can lead to unrealistic goals of restoration (Brinson and Rheinhardt, 1996; Ehrenfeld, 2000b). This means that restoration standards should represent the highest level of functioning possible of a particular type of wetland in a given watershed or region, for example, in areas of similar urban or agricultural development. Even within highly developed watersheds, highly functioning wetlands may exist. In a Pennsylvania watershed, for example, riverine floodplain sites were found to be high-functioning despite nearby development, possibly because they are more influenced by land use in distant parts of the watershed than nearby land use (Wardrop et al., 2007). Given the myriad potential measures of “success” of wetland restoration efforts, it is difficult to quantify what is meant by success and how successful wetland restoration efforts have been. Determination of legal success requires that regulators or other parties conduct monitoring of restored sites to determine if permit conditions have been met; however, postconstruction monitoring is often conducted for too short of a time period (Mitsch and Wilson, 1996; Zedler and Callaway, 1999), or may not be conducted at all due to personnel or budgetary constraints. Assessments of ecosystem success are often conducted by restoration ecology researchers or environmental managers who typically compare the characteristics and functions of restored and reference wetlands but generally do not make a declaration of success or failure due to the difficulty of defining what constitutes success, rather pointing out similarities or differences between restored and reference sites. Landscape success may be even more difficult to evaluate given the complexity and large spatial scale of watersheds and wetland–stream–upland networks. Despite the difficulties of defining success, several reviews have indicated that wetland restoration projects often do not attain conditions that can be deemed legally or ecologically successful (Mitsch and Wilson, 1996; Zedler and Callaway, 1999; Craft et al., 2003). This includes some restored TFW. For example, Simenstad and Thom (1996) found that only a few of 16 ecosystem attributes monitored were on a functional trajectory approaching that of reference wetlands. Similarly, ecosystem monitoring at restored TFW in the US Pacific Northwest, Fraser River in British Columbia, Louisiana coastal zone, and mid-Atlantic region indicates that persistent differences exist between restored sites and reference wetlands or design goals (Bartoldus, 1990; Levings and Nishimura, 1997; Quigley, 2001; Gray et al., 2002; Neff, 2002, Tanner et al., 2002; Adams and Williams, 2004). Similarly, TFW restored for mitigation purposes met some criteria for legal success (vegetation coverage or survival), but not others (hydrology) (Quigley, 2001; Land Management Group, 2004). Taken together, these findings suggest that there is considerable room for improvement of techniques and approaches for restoration of TFW.
4. ECOSYSTEM ATTRIBUTES M EASURED AT R ESTORED TFW Evaluating the success of wetland restoration efforts depends almost entirely on the relevance of measured ecosystem attributes to restoration objectives and the rigor of ecosystem monitoring. In this section we summarize information on
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ecosystem attributes that have been monitored specifically at restored TFW for evaluating restoration success. This ecosystem monitoring information is used in conjunction with success evaluation approaches (Section 3) to develop restoration success criteria for restored TFW in Section 5. One contributing factor to the mixed success of TFW restoration projects is a lack of consistent monitoring of processes controlling geomorphological development or vegetation establishment. More consistent monitoring would help identify benefits and limitations of specific restoration techniques. Most of the restored TFW studies listed in Table 1 include some form of vegetation monitoring, but other parameters are not routinely measured at restoration sites, and some are included only in specific regions where particular functions of wetlands are of interest (e.g., salmonid fish in the Pacific Northwest and the Fraser River in Canada). Surprisingly, hydrology and geomorphology are not measured at some sites, despite their overriding influence on biogeochemistry, vegetation, and fauna. Similarly, salinity measurements are generally not made, perhaps under the assumption that TFW are by definition “fresh,” defined by Cowardin et al. (1979) as below 0.5 salinity. However, salinity may rise during periods of low water flow during droughts, increasing salinity to 5 or higher (Baldwin, 2007). These periodic events may reduce growth of salt-sensitive plants, altering fish and invertebrate communities, and introducing sulfate and other materials, and may become more frequent as sea level continues to rise (see Neubauer and Craft, 2009). In the case of tidal freshwater swamps, periodic salinity intrusion may promote coexistence and hence diversity of plant species (Baldwin, 2007). Aquatic invertebrate communities are natural integrators of habitat and water quality conditions, and have been described at several restored wetland sites. The importance of invertebrates as a food source for other organisms means that they are also an indicator of the value of the wetland for supporting fish and wildlife (Brittingham and Hammerschlag, 2006). Aquatic invertebrate studies are a wellestablished biological monitoring technique for streams (Resh et al., 1996). While methodologies for wetland invertebrate biomonitoring are not as well-developed as for streams (King and Richardson, 2002), invertebrate monitoring has nonetheless been increasingly used to assess wetland condition (Burton et al., 1999; Guntenspergen et al., 2002; Craft et al., 2003; Hartzell et al., 2007). Assuming that hydrologic and soil conditions permit the development of wetland vegetation and invertebrate communities, and that suitable landscape habitat connectivity and nearby source populations exist, it is almost inevitable that fish, birds, mammals, and herpetofauna (reptiles and amphibians) will visit or colonize the restored site. Because of the visibility, charisma, and value to the public of these organisms, as well as stipulation in regulatory permits, it may be important to document the presence of fish and wildlife at restored sites. While their presence documents that the ecosystem is functioning to provide habitat, monitoring of fish and wildlife may not be a robust measure for comparison between restored and reference sites because of normal spatial and temporal variability in population density, or because the area of restored wetlands is small relative to characteristics of the larger habitat controlling fish and wildlife populations. Focusing on wetlanddependent species, such as waterfowl, may improve resolution and the utility of wildlife surveys for evaluating wetland function.
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5. C RITERIA FOR SUCCESSFUL RESTORATION OF TFW Criteria for evaluating restoration or mitigation success have been developed or applied to many types of wetlands including salt marshes (Weinstein et al., 1997), nontidal wetlands (Wilson and Mitsch, 1996; Cole and Shaffer, 2002), and mitigation banks (large wetlands created, restored, enhanced, or preserved in an off-site location to compensate for authorized natural wetland impacts; Spieles, 2005). Generally evaluation efforts have focused on geomorphology (including soils), hydrology, vegetation development, and wildlife usage. Some studies in salt marshes (Craft et al., 2003) and TFW (Simenstad and Thom, 1996; Baldwin and DeRico, 1999) have gone farther, adding quantitative studies of algae, biogeochemical processes, seed banks, invertebrates, fish, or birds as indicators of ecosystem function relative to reference sites. Based on these studies and literature reviewed in Section 3, it is clear that assessments of geomorphology, hydrology, vegetation, and fauna are accepted measures of restoration success used in many types of wetlands (although they are not consistently measured across sites, as noted previously). Here we propose success criteria specific to restored TFW that address these and related ecosystem attributes (Table 2). As for evaluations of other types of restored wetlands, the success criteria we propose for restored TFW are primarily based on comparison with reference sites. Because success criteria are thus dependent on the condition of the reference site, it is critical to choose reference sites that experience environmental conditions similar to the restored site (Ehrenfeld, 2000b). Watershed urbanization or agricultural development may impose landscape constraints on ecosystem components that cannot be overcome through restoration. For this reason, it is unrealistic to expect that restored TFW in urbanized landscapes with high sediment loads, flashy hydrology, fragmented wetlands, and abundant nonnative species, for example, will closely resemble those of nonurban landscapes (Baldwin, 2004). The criteria proposed in Table 2 reflect the need to apply success criteria that can realistically be achieved given environmental constraints imposed by the landscape or watershed surrounding the restored site. In addition to watershed condition, hydrologic attributes such as tidal range and connectivity to rivers should be similar at the reference and restored sites. If suitable reference sites are not available or evaluated, restored wetlands can be compared with accepted standards of wetland function (Wilson and Mitsch, 1996) or with literature values for naturally occurring wetlands in similar watersheds. Some criteria for restoration success may not require comparison with reference sites. Often these are compliance success criteria specified in permit requirements for mitigation projects implemented to replace wetlands lost to development. These may include goals specifying a certain percentage of vegetation cover, a particular hydrologic regime, a preponderance of hydrophyte plant species, use of restored areas by fish and wildlife, or a maximum threshold of nonnative species. As noted previously, these types of success criteria are clear and relatively simple to evaluate.
Proposed ecological criteria for successful restoration of TFW
Ecosystem attribute
Measurements
Success criteria
Comments
Hydrology
Duration or percentage of time flooded
Differences in elevation of only a few centimeters (e.g., 3–10) can strongly influence the ability of plants to colonize via seedling recruitment, as well as growth of planted vegetation
Geomorphology
Accretion
Elevation of vegetated high marsh should lie at approximately mean sea level, or up to mean high water (MHW); vegetated low marsh should lie approximately between mean sea level and mean low water (MLW) (Odum et al., 1984). High marsh should be flooded up to 30 cm depth for 0–4 h per tidal cycle; low marsh should be flooded up to 100 cm depth for 9–12 h per cycle (Simpson et al., 1983; Mitsch and Gosselink, 2000); similar to reference sitesa Spatially variable; vertical accretion of 5–10 mm/year (Neubauer et al., 2002), average 6–7 mm/year (Craft, 2007); similar to reference sites Little or no net elevation change relative to sea level; similar to reference sites
Elevation change
Channel development
Evidence of small channel formation without excessive erosion or scour of large channels
Topography
Variable within the small range that supports desired vegetation (MLW–MHW)
809
Restored sites constructed from coarse material (sand and gravel) may not accrete organic matter if elevation is sufficiently high to allow oxidation Little information exists on rates of elevation change in naturally occurring TFW. Excessive erosion, subsidence due to dewatering, or compaction of sediments in restored TFW may lead to vegetation species change or dieback; increases in relative sea level may promote vertical increases in elevation and accumulation of organic matter Large channels cut into restored sites may naturally fill with sediment as small channels form naturally in the wetland (Simenstad and Thom, 1996) Naturally occurring TFW may exhibit small hummocks (10–15 cm; Baldwin, pers. obs.) due to vegetation, and lower areas near channels or in small ponds, but are otherwise very flat
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Table 2
(Continued ) Measurements
Success criteria
Comments
Soil
Organic matter
Evidence of organic matter accumulation in the surface horizon or streaking into subsurface horizons; average organic carbon concentrations for US tidal freshwater marshes (Craft, 2007) is 13–22%
Sites created by placement of mineral soil or excavation of upland soil to a hydrologically correct elevation may accumulate little organic matter due to oxidation of any material that accumulates vertically; organic content level of nonurban, naturally occurring TFW (e.g., 20–70%, Odum, 1988) may develop extremely slowly (Zedler and Callaway, 1999; Craft et al., 2003) Urban reference sites may have low organic matter content compared to nonurban sites Concentrations of heavy metals and organic contaminants in dredge material sources should be determined prior to restoration. However, it is unrealistic to try to reduce nutrients or toxins to levels below those of reference sites that experience the same watershed conditions Sites restored by placing river sediment or excavation of uplands will in general have coarser soil texture (e.g., sand and gravel) than reference sites (silts and clays) (Zedler, 2001) Saltwater intrusion events that occur only during exceptionally dry years may be important in maintaining plant diversity and may alter fish and invertebrate communities (Odum et al., 1984; Odum, 1988; Baldwin, 2007).
Similar to reference sites
Salinity
Nutrients, metals, organic contaminants
Similar to reference sites; average nutrient concentrations for US tidal freshwater marshes (Craft, 2007) are 0.9–1.6% N and 0.9–1.6 mg/g P
Texture
Evidence of surface accumulation of materials of similar texture to reference sites; average bulk density for US tidal freshwater marshes is 0.1–0.3 g/cm (Craft, 2007)
Salinity
Average salinity 0.5; pulses up to 5 or higher may occur during droughts
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Ecosystem attribute
Species composition
The list of perennial species should be similar to those of reference sites, but differences in relative abundance should be accepted
Annual species should comprise 20–50% of species (Leck et al., 2009) and a peak annual:perennial biomass ratio of 1:5 (Whigham and Simpson, 1992) Species richness
Similar to reference sites
Biomass or total plant cover
Similar to reference sites
Nonnative species abundance
Similar to or less than reference sites
Species such as cattail (Typha spp.) are adapted to rapid colonization of exposed, moist substrate such as that created in restored TFW; a high abundance of these native invasive species should be accepted as a natural result of their biology and the environmental conditions created by restoring TFW. Species composition of annuals is likely to differ from reference sites due to the inability of some to colonize; seeding may be required to introduce these species Annual species are a key characteristic of naturally occurring TFW (Simpson et al., 1983; Odum et al., 1984). Desired annual species may be slow to colonize restored wetlands sites (Neff and Baldwin, 2005) Additional plantings or seeding may be necessary if few propagules are present in waterways; invasive or nonnative species may contribute substantially to richness in urban areas (Neff, 2002; Rusello, 2006) Standing biomass and total plant cover are indices of primary production. Belowground biomass may be indicative of adverse physical structure or biochemistry of soils for plant growth Urban wetlands that might serve as reference sites often contain nonnative plants. Expectations that restored TFW in urban areas remain free of non-native species may be unrealistic (Baldwin, 2004)
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Table 2
(Continued ) Measurements
Success criteria
Comments
Seed banks
Species composition
Dominant species similar to reference sites
Seed density
Similar to reference sites
Species richness
Similar to reference sites
Benthic invertebrates
Density, species composition, species richness
Similar to reference sites
Fish, birds, mammals, herpetofauna
Density and species composition
Present at restored sites
Because seed bank composition is related to the composition of vegetation and seeds dispersing to restored sites, watershed characteristics are likely to have a strong influence on seed banks of restored TFW and reference sites within the same watershed (Neff et al., 2009) Higher seed density may occur in restored than in reference TFW sites (Baldwin and DeRico, 1999; Leck, 2003; Neff et al., 2009) Richness may be higher in restored than reference sites (Baldwin and DeRico, 1999; Leck, 2003; Neff et al., 2009) Invertebrate communities integrate the water, soil, and vegetation habitat quality functions of the wetland, as are a measure of the capacity of the wetland to support fish, herpetofauna, mammals, and birds It may be useful to document value of wetlands as habitat for particular groups. Comparing restored and reference sites may not be practical due to seasonal and spatial variability in populations. Sampling wetland-dependent guilds may improve resolution
If the restored site is in an urbanized (or agricultural) landscape, reference sites should also be located in an urban (or agricultural) landscape with similar watershed characteristics. a If reference sites are not available, restored sites can be compared to accepted standards of wetland function (Wilson and Mitsch, 1996) or literature values for naturally occurring wetlands in similar watersheds or landscapes.
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Ecosystem attribute
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5.1. Hydrologic criteria Hydrology is the overriding factor controlling ecosystem development in wetlands (Mitsch and Gosselink, 2000), yet is often not measured directly in restored TFW. The high and low marsh zones of TFW are flooded at depths and for periods of time that should be considered in evaluating restored wetland success (Table 2). In practice, it is often relatively simple to install water level recorders at restored and reference sites, and survey the elevation of plots for monitoring vegetation, soils, or other parameters relative to the ground surface at the recorder. Then the duration or proportion of time flooded can be readily calculated from the water level and relative elevation data by assuming a flat water surface. If it is not possible to measure proportion of time flooded, elevation relative to a tidal datum can be a useful surrogate, although without a high-precision global positioning system (GPS) it is often difficult to determine absolute elevation of restored and reference sites where there are no nearby benchmarks. In evaluating the restored wetland for its capacity to support typical TFW vegetation, the duration of flooding may be a more relevant metric than the depth; the emergence of most TFW seed bank species, many of them annuals, is strongly inhibited by shallow (3–5 cm) depths of flooding (Baldwin and DeRico, 1999; Peterson and Baldwin, 2004; Hammerschlag et al., 2006). Because annual species are a major component of TFW vegetation biomass (Whigham and Simpson, 1992) and diversity (Parker and Leck, 1985), small increases in water level have been found to reduce abundance of annual species and to lower plant diversity (Baldwin et al., 2001). Furthermore, plant diversity in TFW was found to be significantly and negatively related to proportion of time flooded, regardless of depth (Neff, 2002).
5.2. Geomorphological criteria Indicators of geomorphologic success include accretion rates, changes in elevation, development of small channels, and microtopographic variation (Table 2). Accretion is likely to be spatially variable, but should be similar to reference sites. Average accretion of 5–10 mm/year was reported in Virginia TFW (Neubauer et al., 2002), and a recent study and review of about 40 studies found average accretion rates of 6–7 mm/year for TFW across different regions of the US coast (Craft, 2007). Elevation may change rapidly across the restored site after construction. Subsidence of dredged sediments may occur due to dewatering, degassing, consolidation, and compaction, or surface sediments may erode, increasing the degree of flooding of the wetland surface. Conversely, eroded sediments may settle in channels or depressions that experience low flow velocity, altering hydrologic exchange and vegetation composition. Sediments may also move unevenly across a restoration site as a result of storm surges with extensive deposition occurring in one place but undergoing scour in another. Once initial rapid geomorphological changes have stopped (e.g., after 1–2 years), the surface of restored TFW should show little change with respect to relative sea level. Where relative sea level is increasing, accumulation of organic matter and mineral sediment should allow the wetland to keep pace with rising sea levels.
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A robust measure of elevation change that incorporates both the surface processes of accretion and erosion and subsurface processes related to soil subsidence or expansion is provided by the Surface Elevation Table – Marker Horizon (SETMH) method (Cahoon et al., 1995). The SET (Boumans and Day, 1993; Cahoon et al., 2002a,b), attached to a benchmark driven to refusal into the substrate, is used in conjunction with cores collected from marker horizons (Cahoon et al., 1996) to measure sediment processes including vertical accretion (deposition) and surface elevation change, from which soil subsidence or expansion can be calculated. A similar technique used at a restored TFW in the Pacific Northwest involves measurements of the height of wooden stakes inserted 0.8 m into the soil, combined with marker horizons (Simenstad and Thom, 1996). The stake-marker horizon method allows for a large number of points to be sampled at relatively low cost, but samples a shallower depth of the substrate than the SET-MH approach. Thus the stake-marker horizon method incorporates a smaller proportion of the substrate processes that influence wetland elevation. Use of both methods together would provide more complete spatial coverage, both horizontal and with depth, for the evaluation of accretion and subsidence dynamics in restored TFW. A successful restoration project should show evidence of small channel development without excessive scour of large channels. Channel development can be observed qualitatively, or mapped, to describe increasing hydrologic connectivity of the wetland to other surface waters. While large tidal channels (5–10 m wide) are sometimes cut into restored TFW during construction, smaller lateral channels (e.g., <1 m wide and <0.3 m deep; Simenstad and Thom, 1996) should develop naturally due to tidal flood and ebb. Excessive erosion and infilling of large channels can also be identified during monitoring. It is likely that large channels constructed to facilitate tidal exchange will decrease considerably in size and depth due to sediment deposition within a few years (Simenstad and Thom, 1996; Baldwin, pers. obs.). Burrowing by muskrats (Ondatra zibethicus) may also create narrow but deep channels (<0.5 m wide, >0.25 m deep; J. Perry, pers. comm., Virginia Institute of Marine Science). Topographic surveys are useful in describing large changes in elevation due to subsidence, or determining adherence to design plans. Elevation should occur within a small range that supports desired vegetation (mean low water to mean high water; Table 2), but microtopographic variation should develop over time, as occurs due to hummock formation, sediment deposition, and channel formation. Microtopographic variation can also be created mechanically by disking after grading the substrate to the correct elevation, which may enhance plant diversity (Moser et al., 2007).
5.3. Soil criteria Because restored coastal wetlands are often created by placement of dredged material or excavation of uplands, their soil often consists of coarse, sandy, or gravelly material, in contrast with the silt or clay texture of soils in naturally occurring wetlands (Zedler, 2001). These coarse soils may require many years before they resemble those of reference sites; Craft et al. (2003) found significantly lower pools of organic carbon and nitrogen in restored salt marsh soils than in reference wetlands, even after almost 30 years. The low rates of organic matter
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accumulation in soils of restored wetlands may be due to stimulation of decomposition by the higher hydraulic conductivity of coarse soils than in reference wetlands (Zedler, 2001; Craft et al., 2003). Alternatively, if the mineral substrate is at the correct elevation for tidal wetland vegetation growth, additional vertical accretion of organic matter may not occur due to oxidation of any material that accumulates above a certain elevation. An exception would be under conditions of sea-level rise or subsidence, when rising relative sea level would promote waterlogging, reducing decomposition rate and promoting accretion of organic material. Scouring in areas of low vegetation cover may cause winter transport loss of organic matter formed during the prior year. A realistic criterion for successful soil restoration is to document evidence of surface accumulation of silts, clays, or organic material, or streaking of organic material into subsurface layers (Table 2). It is unlikely that organic matter content and texture of restored TFW soils will resemble that of reference sites for many years. An exception would be in urbanized or heavily agricultural landscapes where watershed alteration prevents development of soil texture and organic matter content typical of relatively undisturbed naturally occurring wetlands. In these cases, the standard for successful soil development should be reference sites located within the same or similar watersheds. Environmental contaminants such as heavy metals or organic contaminants may occur in dredged material (levels should be determined before construction), and could result in ecological impacts to fish and wildlife if present at elevated concentrations. Again, however, it is unrealistic to expect that pollutant concentrations will be lower than those of appropriately chosen reference sites (Table 2). The same applies to nutrients, which may occur at elevated levels in watersheds due to intensive urban or agricultural land uses.
5.4. Salinity criteria Average salinity of restored TFW should be 0.5 (Cowardin et al., 1979), although pulses up to 5 or higher may occur periodically during droughts (Odum et al., 1984; Odum, 1988; Baldwin, 2007). If oligohaline marshes are the goal of restoration, average salinity should be 0.5–5 (Cowardin et al., 1979). As noted previously, the salinity of surface or interstitial water is often not included in studies of restored TFW. Because it is easy and inexpensive to measure, salinity (or electrical conductivity and temperature, from which salinity can be calculated) should be measured regularly at restored and reference TFW, particularly during droughts that may allow saline water to intrude into normally fresh areas. Salinity data may also be available from environmental agencies as an alternative to measurement. The salinity regimes of restored and reference TFW should therefore be relatively straightforward to assess.
5.5. Vegetation criteria Vegetation is probably the most widely monitored ecosystem component at restored wetlands (Wilson and Mitsch, 1996). The data collected on plant communities most commonly focus on the abundance of individual species, allowing
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comparisons of species composition and richness between restored and reference sites. However, in systems such as TFW where water movement is a dominant factor controlling ecosystem development, it may be more appropriate to focus on ecosystem or landscape processes as goals for successful restoration rather than on species-focused goals (Ehrenfeld, 2000a). Relevant to the vegetation criterion is the appropriate selection of reference sites as discussed previously. For example, TFW in nonurban landscapes generally contain few nonnative species, while those in urban landscapes often contain nonnatives or native invasive species among their community dominants. Specifically for restored TFW, a realistic criterion is that the list of perennial species be similar between the restored and appropriately chosen reference sites (Table 2). Invasive native species such as cattail (Typha spp.) that naturally occur in TFW should be expected as a large component of the restored vegetation given their capacity for rapid colonization of disturbed, moist substrates. Thus it is unrealistic to expect rapid development of patterns of relative abundance of plant species like those of naturally occurring TFW. Additionally, because of the importance of annual species in TFW, annuals should be an important component of the restored TFW vegetation. While seeds of annuals are likely to disperse to restored sites via wind, water, and animals, some desired annuals may be slow to colonize (Neff and Baldwin, 2005). Even if species composition differs between restored and reference TFW, species richness may be similar or even higher at reference than restored sites (Leck and Leck, 2005). Thus, if biodiversity is a criterion for success, it may be appropriate to recognize non-native or other “undesirable” species as a component of diversity (Table 2). This does not mean that degraded or low-functioning wetlands are acceptable targets for restoration, but rather that reference sites or standards representing the characteristic levels of functioning within the landscape or watershed of the restored sites should be chosen, as discussed in Section 3. For example, it is likely that nonnative species occur in reference wetlands in urbanized watersheds at greater abundance than in nonurban wetlands (Baldwin, 2004). Plant biomass and total plant cover are indicative of primary production, and are appropriate indicators of restoration success (Table 2). High biomass of nonnative or native invasive species may provide this ecosystem function to a similar or greater extent than the vegetation of reference wetlands.
5.6. Seed bank criteria Seed banks are an important component of vegetation dynamics of TFW, serving as a reservoir of propagules for annual recruitment into vegetation or for regeneration following disturbance (Parker and Leck, 1985; Baldwin and Mendelssohn, 1998). Seed bank studies, which typically involve enumeration of seedlings emerging from soil samples in a greenhouse or growth chamber, can thus provide an integrative measure of vegetation dynamics. The seed banks of restored TFW may develop rapidly due to seed dispersal into restored sites (primarily via water) or production of seed from colonizing or planted species, in some cases to seed density and diversity levels greater than reference sites (Leck, 2003; Neff and Baldwin, 2005; Neff et al., 2009).
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Seed bank composition is related to vegetation and seed dispersal, and therefore also to watershed condition. A realistic criterion for success is that the dominant species of the reference sites also occur in seed banks of the restored site (Table 2). However, as noted previously, certain species at reference sites may not be dispersed to restored sites, and therefore are unlikely to occur in the seed bank without planting or seeding. Because seed dispersal into restored TFW with connectivity to surface waters and nearby source populations can be prolific (Neff and Baldwin, 2005), rapid development of seed banks with high seed density and diversity is expected (Baldwin and DeRico, 1999; Neff et al., 2009). Therefore, density and diversity of seeds in the seed bank are realistic criteria for success in comparing restored and reference TFW (Table 2).
5.7. Benthic invertebrate criteria As noted previously, aquatic invertebrate communities are useful indicators of ecosystem habitat structure and function and food chain support. Density, species composition, and species richness are simple metrics that can be measured using, for example, aquatic dip net sweeps through vegetation, emergence traps, sediment cores or dredge samples, or artificial substrate samplers, and compared with invertebrate communities of reference sites (Table 2). The benthic community may also reflect water and sediment quality, especially as related to pollutants. As for plants, soils, and other parameters, reference sites should be located within the same watershed or in watersheds with similar levels of urban or agricultural land uses.
5.8. Fish and wildlife criteria The presence of fish, birds, mammals, reptiles, and amphibians at a restored site is indicative of the habitat function provided by that site for food and cover (Table 2). A broad range of animals is associated with TFW, as described in Whigham et al. (2009). Because of the temporal and spatial variability of faunal populations, intensive monitoring at reference and restored TFW may be necessary to detect differences in fish and wildlife usage if the sites differ greatly in vegetation coverage or species composition, extent of mudflats, channel depth or density, or adjacent upland vegetation type. However, because there may also exist differences in these structural features not related to the restoration, it may be difficult to relate detected differences in fish and wildlife communities to the success (or failure) of the restoration.
6. C ASE STUDY : EVALUATION OF R ESTORED TFW OF THE A NACOSTIA R IVER , W ASHINGTON , DC, USA In this section we apply the success criteria proposed in Section 5 (Table 2) to TFW that have been restored along the Anacostia River, located in Washington, DC. We first describe the general characteristics of restored and reference sites, and then evaluate the restored TFW using the proposed success criteria.
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6.1. Characteristics of restored and reference sites As is the case for many urban rivers, the ecosystems of the Anacostia River in Washington, DC, have been substantially altered by human activities. Development and related projects such as flood control, mandated dredging, and landfills destroyed more than 3,900 ha of forested and herb-dominated wetlands, including about 1,000 ha of TFW along the Anacostia River (Schmid, 1994; US EPA, 1997), leaving only a few hectares of fragmented TFW. Since that time several TFW restoration projects have been implemented, including Kenilworth Marsh and Kingman Marsh (Figure 2). Historically a TFW existed at the location of Kenilworth Marsh, but the site was dredged to create a recreational lake in the 1940s (Syphax and Hammerschlag, 1995). The US Army Corps of Engineers (USACE) restored wetlands at the site in 1992–1993 by pumping about 115,000 m3 of sediment dredged from the adjacent Anacostia River into containment cells and planting more than 340,000 plants of 16 species (Bowers, 1995; Syphax and Hammerschlag, 1995). At Kingman Marsh, the USACE restored TFW in early 2000 in a similar fashion by placing Anacostia Anacostia river Anacostia river Dueling creek (natural urban reference)
New York
Pennsylvania
Maryland
Kenilworth marsh (restored 1992–93)
New Jersey Delaware
West Virginia Virginia
Kingman marsh (restored 2000)
Patuxent river
N Anacostia river
1 km
Figure 2 Location of restored TFW associated with the Anacostia River in Washington, DC (Kingman and Kenilworth Marshes), and urban (Dueling Creek) and nonurban (Patuxent River) reference wetlands.The panel on the left shows the location of the Anacostia and Patuxent Rivers within the Chesapeake Bay watershed. The panel on the right shows the approximate boundaries (in white lines) of areas of restored marshes at the Kingman and Kenilworth sites, and the Dueling Creek reference site. Aerial photo from Globe eXplorer was taken in April, 2002.
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River sediment and planting 750,000 plants of seven species. This created about 13 ha of vegetated wetland (AWRC, 2002). Sediment elevations were designed to be lower than those at Kenilworth Marsh to reduce colonization by invasive species (including Lythrum salicaria, Phragmites australis, and Typha spp.). As part of postconstruction monitoring for the Kingman Marsh project, two natural TFW with similar tidal ranges were selected as reference sites to provide a basis for evaluating vegetation development in the restored wetlands (Figure 2). One of these sites, Dueling Creek Marsh, is a remnant 0.41 ha urban wetland located on a small tributary to the Anacostia River 0.8 km upstream of Kenilworth Marsh. The other site, Patuxent River Park, is a nonurban TFW located along the Patuxent River in an adjacent watershed. Throughout the rest of this section, the Kingman Marsh restored site will be referred to as KING (REST) and the Kenilworth Marsh restored site referred to as KEN (REST). The Dueling Creek urban reference site will be abbreviated as DUEL (UREF); the Patuxent nonurban reference site will be abbreviated as PAX (NREF).
6.2. Evaluation of success of restored TFW The goal of ecosystem monitoring at KING (REST) was to “document both the status and the degree to which the reconstructed marsh achieved a wetland condition similar to reference emergent freshwater tidal wetland habitat” (Hammerschlag et al., 2006). Information from monitoring studies is used here to evaluate the success of the KING (REST) and KEN (REST). 6.2.1. Hydrologic evaluation The results of hydrologic monitoring (Hammerschlag et al., 2006) indicate that flooding duration at KING (REST) was more prolonged than at DUEL (UREF) but similar to PAX (NREF) (excluding one transect located in a low marsh area) (Figure 3). KEN (REST), in contrast, experienced less flooding than PAX (NREF) but averaged about the same duration of flooding as DUEL (UREF). Because DUEL (UREF) is in the same watershed as the restored sites, it may be a more appropriate reference site than PAX (NREF). Comparison with DUEL (UREF) suggests successful hydrologic restoration at KEN (REST) but probably not at KING (REST). 6.2.2. Geomorphologic evaluation Surface Elevation Table (SET) and marker horizon measurements were made at five locations at both KING (REST) and KEN (REST) over a 3-year period beginning in October, 2002, about 2½ years after sediment placement. Accretion and elevation change trends were determined from regression analyses. At both wetlands, accretion rates were >20 mm/year and were at least double the rate of elevation gain (Table 3), indicating that soils were subsiding at 13 mm/year at KING (REST) three to 5 years after restoration, and 17 mm/year at KEN (REST) ten to twelve years after restoration. Thus both restored TFW are accumulating sediments and growing vertically within the tidal range despite ongoing consolidation of the dredged material substrate. The slower rates of accretion and
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0.5
a
Inundation (proportion of time flooded)
ab 0.4
a ab
0.3
ab
All transects Without low marsh transect
0.2
b
0.1
0.0 KING1 KING2 KEN1 KEN2 DUEL
PAX
Area
Figure 3 Proportion of time inundated for each wetland area (data from 2001). Values are means þ SE. Means for “all transects” sharing the same letter are not significantly different (Tukey test). Areas are abbreviated as follows: KING1 (Kingman Area 1, restored), KING2 (Kingman Area 2, restored), KEN1 (Kenilworth Mass Fill1, restored), KEN2 (Kenilworth Mass Fill 2, restored), DUEL (Dueling Creek, urban reference), and PAX (Patuxent River, nonurban reference). For the Patuxent site, means are presented for all transects combined (in solid black), and for the mean calculated without one transect that was located in a low marsh area, and thus was considerablylower in elevation and flooded longer than other transects (patterned fill). Low marsh habitat was not sampled at anyof the other sites. Adapted from Neff (2002). Table 3 Mean annual accretion and elevation change trends for 3 years for two restored TFW within the Anacostia River, Washington, DC Location
Process
Kingman
Accretion Elevation change Accretion Elevation change
Kenilworth
Mean Rate (mm/year)
Standard error
d.f. error
F
P
26.5 12.9
3.0 3.0
3
18.35
0.0234
22.3 5.5
3.9 3.9
4
19.7
0.0113
elevation gain at KEN (REST) are consistent with the fact that this wetland is at a higher level within the tidal range than KING (REST). No SET measurements were made at the reference sites, preventing comparisons between restored and reference sites. However, the increases in elevation and demonstrated surface material accretion at KING (REST) and KEN (REST) suggest that success criteria were met (Table 2) and that depositional processes and production of organic matter may lead to development of surface soils at some locations comparable to reference wetlands. However, it should be noted that the
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location of the SETs at KING (REST) did not encompass processes throughout the reconstructed wetlands. Some locations not measured by the SETs were undergoing erosion (Hammerschlag, pers. obs.). Erosion also occurred in the form of channel development at KING (REST) and KEN (REST), and some of the designed large channels dug into the KING (REST) site to restore hydrology were observed to widen and also fill in quickly with sediment eroded from the built marsh (as was noted in a restored wetland studied by Simenstad and Thom, 1996). Channels appeared to follow smooth dendritic patterns with no obvious signs of rapid erosion of underlying material. Taken together, it appears that restoration efforts at KING (REST) and KEN (REST) were successful in establishing accreting wetlands that exhibit microtopographic variation and development of small dendritic channels (Table 2). However, elevation is probably too low at a majority of locations in KING (REST), as inferred from hydrologic measurements (Section 6.2.1). 6.2.3. Soil evaluation Soils were analyzed by collecting soil cores from within the rooting zone (0–30 cm depth) at vegetation monitoring plots at both restored and reference sites and analyzed for texture, organic matter content, total nitrogen, phosphorus, carbon, sulfur, cadmium, copper, chromium, lead, nickel, and zinc. Additionally, redox potential was measured (Neff, 2002; Hammerschlag et al., 2006). Organic matter content was more than three times higher at PAX (NREF) than at any of the urban sites. While low organic matter content might be expected at KING (REST) and KEN (REST) (sites that were restored using river sediment to raise elevations), a similarly low level at the urban natural reference site DUEL (UREF) is indicative of elevated mineral sedimentation or decomposition rates. The two restored sites, KING (REST) and KEN (REST), had a higher sand content than DUEL (UREF) consistent with their construction using Anacostia channel-derived dredged material. These data suggest that organic matter at the restored sites may never approach that at PAX (NREF). Levels of total carbon, nitrogen, and phosphorus paralleled that of organic matter content, which is expected given the abundance of these elements in plants. Levels of heavy metals were highest at DUEL (UREF), with the exception of cadmium, which occurred at the highest concentration at PAX (NREF), possibly because of the presence of cadmium in phosphate-based fertilizers. Redox potential was highest at DUEL (UREF) and lowest at PAX (NREF), another factor possibly contributing to higher oxidation rates and therefore lower organic matter content at DUEL (UREF) than at PAX (NREF). These results indicate that restoration of soil characteristics at KING (REST) and KEN (REST) was successful, based on comparison with DUEL (UREF) (Table 2). 6.2.4. Salinity evaluation Salinity was not measured directly at either restoration site. However, salinity was <0.2 from 1985 to 2005 at a monthly monitoring station on the Potomac River downstream of the Anacostia–Potomac confluence (mddnr.chesapeakebay.net).
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Therefore, droughts do not appear to cause any significant intrusion of saline water into the Anacostia River, and the salinity regime was correct at the restored sites (Table 2). 6.2.5. Vegetation evaluation Our analysis of plant community development from measures of vegetation cover in transects (Neff, 2002; Hammerschlag et al., 2006; Rusello, 2006) reveals significant loss of vegetation cover, species richness, and diversity at KING (REST) but not at any of the other studied wetlands. Thus richness and cover criteria were not met at KING (REST) (Table 2). The vegetation impacts at KING (REST) are attributable to herbivory by resident Canada geese (Figure 4) coupled with effectively lowered sediment elevations following reconstruction (due to removal of sediment containment structures, soil compaction, and sediment loss). Populations
Figure 4 Aerial photos of Kingman Marsh (Cell 1, approximately 2 ha, lowermost area depicted in Figure 2) prerestoration (1998), the first growing season following sediment placement and planting (2000), and in subsequent years (2001^2004). A dramatic reduction in vegetation coverage, due to herbivore grazing and excessive inundation, is visible between 2000 and 2001. Photos from Hammerschlag et al. (2006).
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of resident Canada geese were three to five times larger in the area of KING (REST) than KEN (REST) during, 2001–2004 (Paul et al., 2006). The detrimental impacts of grazing by geese and excessive inundation on vegetation at KING (REST) are evident in the much higher area of “No cover” (73%) compared to the other sites ( 22%) (Table 4). The site with lowest area of “No cover” (i.e., highest vegetation cover) was KEN (REST), which had cover of 10–20% individually for Leersia oryzoides, Peltandra virginica, Phragmites australis, and Typha spp. This finding suggests that rapidly colonizing, high-productivity species such as Phragmites and Typha are important in ecosystem productivity and nutrient cycling, and may provide additional services such as water quality improvement, particularly in highly urbanized landscapes where other species cannot establish or grow due to environmental constraints. Furthermore, these species may function as “nurse” species by providing safe sites for seedling establishment that are protected Table 4 Areal extent of soil containing no plant cover (listed as NO COVER) and the average cover of dominant plant species across 3 years (2002, 2003, and 2004) at restored and reference TFW Species
Kingman Marsh
Kenilworth Marsh
Dueling Creek
Patuxent Marsh
NO COVER Acorus calamus Hydrilla verticillataa Impatiens capensisb Juncus effususc Leersia oryzoides Lemna minor Ludwigia palustris Ludwigia peploides Lythrum salicariaa Nuphar luteac Orontium aquaticum Peltandra virginicac Phalaris arundinacea Phragmites australisa Polygonum arifoliumb Polygonum punctatumb Polygonum sagitattumb Salix nigra Schoenoplectus fluviatilis Schoenoplectus tabernaemontanic Sparganium eurycarpum Typha spp. Zizania aquaticab
73 + 1.7 – – X 2 + 0.4 X – 2 + 0.4 1 + 0.2 4 + 0.4 X – 8 + 0.9 – 1 + 1.1 – 1 + 0.5 X 3 + 0.5 – 1 + 0.3
6 + 2.6 – – 6 + 0.7 X 13 + 1.0 – X – 2 + 0.7 – – 17 + 1.4 4 + 0.5 17 + 1.7 X X X 1 + 0.7 9 + 0.9 X
22 + 4.2 – – 4 + 1.2 – 30 + 1.6 – X – 3 + 1.1 – – 11 + 2.3 4 + 0.7 – 6 + 1.2 10 + 1.2 3 + 0.7 – – X
15 + 3.0 6 + 0.7 10 + 1.4 6 + 0.8 X X 6 + 0.7 X X – 17 + 1.4 3 + 0.3 14 + 1.6 – – 9 + 0.9 X X X X X
– 3 + 0.9 X
– 18 + 1.3 2 + 0.5
3 + 0.5 4 + 2.1 –
3 + 0.3 3 + 1.5 X
Dominant species are the 10 species occurring at the highest mean cover within each site. Values are least-squares means of cover + SE. X = species occurred at the site but was not among the 10 most abundant species at that site. Dashed lines indicate that a particular species was not found at that site. Adapted from Rusello (2006). a Non-native species. b Annual species. c Planted at Kingman Marsh.
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from geese (Baldwin, pers. obs.). In nontidal wetlands Typha latifolia was also found to facilitate establishment of other species by aeration of soil (Callway and King, 1996), but facilitation was not detected in a study of the role of T. latifolia in facilitation of Impatiens capensis in a tidal freshwater marsh (Hopfensburger and Engelhardt, 2007). Rapidly growing invasive species may also ameliorate coarse mineral soils through introduction of organic matter. Most of the dominant species at KING (REST) and KEN (REST) were also dominant at DUEL (UREF) (Table 4). These include the nonnative purple loosestrife, Lythrum salicaria, as well as the invasive Phalaris arundinacea, neither of which occurred at PAX (NREF). Phragmites australis and Typha spp. were both dominant features of the vegetation community at KEN (REST). The annual species I. capensis occurred at both restored and reference sites. However, the tearthumbs Polygonum arifolium and Polygonum sagittatum, which were dominant at one or both reference sites, were rare at the restored sites. 6.2.6. Seed bank evaluation To describe seed bank development at KING (REST), surface soil samples were collected from the restored and reference sites in, 2000, 2001, and, 2003 for seed bank analysis using the emergence method (Baldwin et al., 1996, 2001; Baldwin and DeRico, 1999; Leck, 2003). The seed bank at KING (REST) developed rapidly during the first growing season, showing large increases in emerging seedling density and taxa density between 2000 and 2001 (Neff et al., 2009). In 2003, all restored and reference sites were found to be similar with regard to these parameters. Significantly higher seedling density and species density were also found at a created tidal wetland in Delaware after 1 year of development (Leck, 2003). The nonnative Lythrum salicaria was important at all urban sites in 2003, being the most abundant species at KEN (REST) and DUEL (UREF), and the second most abundant at KING (REST). The results of seed bank studies at the restored and reference sites demonstrate that (1) seed banks of restored TFW develop rapidly, and (2) seeds of nonnative or invasive plants are likely to occur in seed banks of both restored and reference wetlands in urban areas. These studies at KING (REST) and KEN (REST) indicate restoration of seed banks was successful (Table 2). 6.2.7. Benthic invertebrate evaluation Benthic macroinvertebrate organisms were collected using Ekman bottom grab sampler, sediment corer, dip-net, and Hester-Dendy sampler (Brittingham and Hammerschlag, 2006). Samples were collected at least seasonally from tidal channels, tidal mudflats, three vegetation/sediment zones (low, middle, and high marsh), and pools over a 3-year period (2001–2004). Macroinvertebrate density was significantly greater at KING (REST) than KEN (REST) due to more numerous chironomids and oligochaetes (Brittingham and Hammerschlag, 2006). Over 95% of the organisms counted at KING (REST) were either chironomids or oligochaetes and the count was over 85% at KEN (REST). Macroinvertebrate taxa composition at KEN (REST), which had been in existence
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for roughly 10 years since reconstruction, was similar to that of DUEL (UREF), although richness was higher. While about 23 families were represented at KING (REST), KEN (REST), and DUEL (UREF), 30 families were observed at PAX (NREF), demonstrating the importance of choosing a reference site within the same watershed as the restored sites. Unvegetated sites like mudflats and channels generally supported greater numbers of invertebrates (primarily chironomids and oligochaetes) while vegetated sites (increased structural diversity) had higher invertebrate species richness (Brittingham and Hammerschlag, 2006). Benthic invertebrate restoration can be considered successful at both restored wetlands based on comparison with DUEL (UREF) (Table 2). 6.2.8. Fish and wildlife evaluation Bird counts were performed at KING (REST) and KEN (REST) from 2001 to 2004 (Paul et al., 2006). Bird sampling was not performed at the reference sites. Furthermore, fish, mammals, and herpetofauna were not quantitatively sampled at the restored or reference sites. A total of 137 avian species were observed at KING (REST) and 164 at KEN (REST) (177 species total); 124 of the species occurred at both wetlands (Paul et al., 2006). Focusing on several wetland-based guilds helped show the contribution to avian use by the restored wetlands. However, the presence of the overabundant Canada geese year-round made them the dominant avian species. Goose herbivory coupled with excessive inundation reduced vegetation cover at KING (REST) to less than one third its intended area. Despite higher plant cover at KEN (REST), higher numbers of birds were observed at KING (REST), possibly because of the greater extent of mudflat at KING (REST) (Paul et al., 2006). While bird studies did not include samples at reference sites, they do indicate that both restored wetlands provide habitat for numerous species and that success criteria were met (Table 2). However, because birds used mudflats at both sites before restoration, it is not possible to clearly separate the effects of vegetation restoration on bird use. 6.2.9. Summary of restoration success evaluation The restoration at KEN (REST) can be considered a success based on all evaluation criteria when viewed within the framework of the environmental constraints imposed by its highly urbanized watershed. Vegetated marsh established at most areas where sediment was placed at elevations suitable to support plant growth, as envisioned in original plans for the wetland. If the objective was to create a wetland similar to that at PAX (NREF), then the KEN (REST) restoration would likely have been judged unsuccessful for at least the soil, vegetation, and benthic invertebrate criteria. However, the DUEL (UREF) site is a more appropriate reference given its location in the same highly urbanized landscape. The KING (REST) project, in contrast, succeeded in creating considerably less vegetated wetland than the area of sediment originally placed (about 13 ha), due in large part to goose grazing and excessive inundation (due to incorrect elevation). The success evaluation indicates that KING (REST) did not meet hydrology and vegetation success criteria, although the restoration was successful for geomorphology, soil, salinity, seed bank, benthic invertebrate, and fish and wildlife criteria.
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7. CONCLUSIONS AND R ECOMMENDATIONS 7.1. Restoration of TFW in urban landscapes and selection of urban reference sites TFW are among the most productive and species-rich ecosystems of the coastal zone, creating unique challenges for restoration efforts. Monitoring at restored TFW sites in the United States and Canada has demonstrated that restoration of elevation, hydrology, vegetation, geomorphological characteristics and processes, and faunal communities is possible and can be considered successful to varying degrees. The case study of the Kingman and Kenilworth TFW in Washington, DC, highlights the difficulties of reestablishing wetland structure and function in an urbanized landscape. Altered hydrology, environmental pollutants, fragmented landscapes, and nonnative species can override efforts to restore TFW to a structure similar to naturally occurring TFW in non-urban areas (although the abundant suspended sediment in the Anacostia increases the propensity for accretion and the capacity to reestablish sediment elevations). For urban restoration projects, therefore, it makes sense to apply success criteria relative to an urban TFW reference site that preferably is located nearby and in the same watershed as the restored sites.
7.2. Establishment of vegetation Because restoration efforts typically involve extensive earthmoving (e.g., excavation, dredged material placement, grading) and subsequent rapid changes in geomorphology related to tidal hydrology or compaction, a phased approach to wetland restoration is likely to improve success. Increases in inundation due to erosion or subsidence may reduce survival of plantings, but plantings may also help to reduce erosion. If sediment placement, excavation, and grading are completed before or during the dormant season, several months can be allowed for sediment compaction and dewatering. A topographic survey completed at this time will allow determination of suitability of elevation for plant growth, and additional grading can be performed or sediment placed before or during the early spring. While many species are likely to disperse to restored TFW (Neff and Baldwin, 2005), planting or seeding of native species not expected in dispersal pathways may be necessary during the spring to establish desired species, stabilize sediments against erosion, and possibly reduce establishment of nonnative species. If herbivores such as Canada geese are present at or nearby restoration sites, it may be necessary to reduce populations through management or protect sites with fencing for several years until vegetation has established. Dense vegetation dominated by native species has established at two recently restored TFW projects in the Anacostia (Heritage Island and River Fringe) where geese have been excluded, although changes in vegetation that will occur following future fence removal are unknown (Hammerschlag, pers. obs.).
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7.3. Control of nonnative species The benefits of controlling nonnative species in restored wetlands should be weighed against the negative environmental impacts of chemical use, as well as labor and materials costs, particularly if the nonnative species also occur in reference wetlands. Furthermore, the beneficial ecological functions of nonnative species should be considered in decisions regarding their control. However, governmental agencies and conservation groups often emphasize establishment of local native plants. In such cases it may be necessary to support an additional program to suppress unwanted invasive species and promote a habitat that reflects project goals. Based on the Anacostia experience, elevations at or just below mean high tide will support most native high marsh species but will reduce the vigor of invasive species.
7.4. Implications for restoration of TFW In a larger context, this review brings to light a number of considerations that are likely to improve the success of TFW restoration: • Clear objectives or goals for restoration should be established during the early planning stages. This need has been stated repeatedly for wetland restoration in general (Mitsch and Gosselink, 2000; Zedler, 2001), and it applies equally to TFW restorations. • Realistic criteria for success in meeting goals or objectives should be clearly established, preferably with regard to appropriately chosen reference sites. Planners may envision a pristine, diverse, exotic-free wetland as the goal, but this may not be possible in a highly urbanized or agricultural landscape (Ehrenfeld, 2000b; Baldwin, 2004). • Success criteria should include not only criteria for compliance success but also for ecosystem and landscape success. These criteria should be clearly conveyed to groups monitoring the restored wetlands. • The role of nonnative and native species in ecosystem development deserves more investigation. • Consistent assessment of hydrologic and geomorphologic aspects of restored TFW is likely to improve restoration success. • Increased use of adaptive management for several years following restoration should be encouraged, for example, to fine-tune elevations, introduce additional plantings or seeds, or spot-control nonnative plants. • Restoration of TFW should be viewed ecologically as catastrophic landscape disturbances that create high-light, high-nutrient, and moist-soil conditions optimal for rapid colonization by native and nonnative wetland species adapted to colonizing disturbed substrates. These species should be expected as a natural initial phase of vegetation and community development, with the expectation that vegetation development will continue for many years, as influenced by hydrology, geomorphology, seed and propagule supply, and watershed condition. Restoration of TFW is increasingly practiced in North America and Canada. Because of the biological and hydrogeological complexity of TFW, outcomes of
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restoration are generally uncertain. We hope that the success criteria proposed here stimulate discussion and promote dissemination of information that will improve the restoration potential of these wetlands.
ACKNOWLEDGEMENTS The authors thank Chris Craft, Mary Kentula, Mike Haramis, Jim Perry, and an anonymous reviewer for their helpful comments of the manuscript.
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Phillip, M., 2005. Decker Island wildlife area: enhancing delta wetlands one phase at a time. Outdoor California March–April 2005, 4–8. Quigley, P.A., 2001. Wetlands monitoring report, Henderson freshwater tidal wetland restoration, John Heinz National Wildlife Refuge, Monitoring Year 2000. Patricia Ann Quigley, Inc., Norristown, PA. Resh, V.H., Myers, M.J., Hannaford, M.J., 1996. Macroinvertebrates as biotic indicators of environmental quality. In: Hauer, F.R., Lamberti, G.A. (Eds.), Methods in Stream Ecology. Academic Press, San Diego, CA, pp. 647–667. Rusello, K., 2006. Wetland restoration in urban settings: studies of vegetation and seed banks in restored and reference tidal freshwater marshes. M.S. Thesis, University of Maryland, College Park, MD. Schmid, J.A., 1994. Wetlands in the urban landscape of the United States. In: Platt, R.H., Rowntree, R.A., Muick, P.C. (Eds.), The Ecological City: Preserving and Restoring Urban Biodiversity. The University of Massachusetts Press, Amherst, MA, pp. 106–133. Simenstad, C.A., Toft, J., Higgins, H., Cordell, J., Orr, M., Williams, P., Grimaldo, L., Hymanson, Z., Reed, D., 2000. Sacramento/San Joaquin Delta Breached Levee Wetland Study (BREACH). University of Washington School of Fisheries, Seattle, WA. Simenstad, C.A., Thom, R.M., 1996. Functional equivalency trajectories of the restored Gog-Le-Hi-Te estuarine wetland. Ecol. Appl. 6, 38–56. Simpson, R.L., Good, R.E., Leck, M.A., Whigham, D.F., 1983. The ecology of freshwater tidal wetlands. BioScience 33, 255–259. Spieles, D.J., 2005. Vegetation development in created, restored, and enhanced mitigation wetland banks of the United States. Wetlands 25, 51–63. Stillwater Environmental Services, 2003. Decker Island Phase II habitat development and levee rehabilitation project, initial study/mitigated negative declaration public review draft. Stillwater Environmental Services, Davis, CA. Syphax, S.W., Hammerschlag, R.S., 1995. The reconstruction of Kenilworth Marsh, the last tidal marsh in Washington, DC. Park Sci. 15, 15–19. Tanner, C.D., Cordell, J.R., Rubey, J., Tear, L.M., 2002. Restoration of freshwater intertidal habitat functions at Spencer Island, Everett, Washington. Restor. Ecol. 10, 564–576. US EPA (US Environmental Protection Agency), 1997. An environmental characterization of the District of Columbia – a scientific foundation for setting an environmental agenda. US Environmental Protection Agency – Region 3, Philadelphia, PA. Wardrop, D.H., Kentula, M.E., Jensen, S.F., Stevens, D.L., Hychka, K.C., Brooks, R.P., 2007. Assessment of wetlands in the Upper Juniata watershed in Pennsylvania, USA using the hydrogeomorphic approach. Wetlands 27, 432–445. Weinstein, M.P., Balletto, J.H., Teal, J.M., Ludwig, D.F., 1997. Success criteria and adaptive management for a large-scale wetland restoration project. Wetlands Ecol. Manage. 4, 111–127. Whigham, D.F., Baldwin, A.H., Barendregt, A., 2009. Tidal freshwater wetlands. In: Perillo, G.M.E., Wolanski, E., Cahoon, D.R., Brinson, M.M. (Eds.), Coastal Wetlands: An Integrated Ecosystem Approach. Elsevier Science, Amsterdam, pp. 515–534. Whigham, D.F., Simpson, R.L., 1992. Annual variation in biomass and production of a tidal freshwater wetland and comparison with other wetland systems. Va. J. Sci. 43, 5–14. Wilson, R.F., Mitsch, W.J., 1996. Functional assessment of five wetlands constructed to mitigate wetland loss in Ohio, USA. Wetlands 16, 436–451. Zedler, J.B., 2001 Handbook for restoring tidal wetlands. CRC Press, Boca Raton, FL, 439pp. Zedler, J.B., Callaway, J.C., 1999. Tracking wetland restoration: do mitigation sites follow desired trajectories? Restor. Ecol. 7, 69–73.
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P A R T
V I I
COASTAL WETLAND SUSTAINABILITY AND LANDSCAPE DYNAMICS
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C H A P T E R
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S URFACE E LEVATION M ODELS John M. Rybczyk and John C. Callaway
Contents 1. Introduction 2. Measuring Processes that Affect Wetland Elevation 3. Types of Models 3.1. Zero-dimensional mineral sediment models 3.2. Zero-dimensional organic sediment process models 3.3. Geomorphic models 4. Future Directions for Model Improvement 4.1. Data gaps 4.2. Integrating models 4.3. Improved linkage between sediment models and vegetation 4.4. Spatialization 5. Conclusions References
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1. INTRODUCTION Observed and predicted increases in the rate of eustatic sea level rise (ESLR), caused by climate change, have led to concerns regarding the long-term sustainability of coastal wetlands worldwide (Reed, 1995), the essential question being, can wetland elevation keep pace with rising sea levels? Recent observations have revealed an ESLR rate of 3.1 mm/year for the period 1993–2003 compared to a background rate of 1–2 mm/year for the 19th and the early 20th centuries (Meehl et al., 2007). Predicting the future rate of sea level rise; however, is an uncertain business at best because it is a function of numerous complex processes, both physical (e.g., the thermal expansion of water and the melting of glaciers and ice caps) and political (e.g., future carbon emissions). The fourth and most recent Intergovernmental Panel on Climate Change (IPCC) assessment predicts a sea level rise of 0.18–0.59 m by end of the 21st century (Meehl et al., 2007); however, many feel that this estimate is conservative. Note, for example, that the observed rate of ESLR from 1993 to 2003 has already exceeded the lower limit predicted by the fourth IPCC assessment. Rahmstorf (2007) discusses the conservative nature of Coastal Wetlands: An Integrated Ecosystem Approach
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the IPCC reports, largely due to uncertainty surrounding the predictably of melting ice sheets; more rapid melting of global ice sheets, primarily in Greenland and the Antarctic, could lead to even greater rates of sea level rise. Rahmstorf (2007) predicted a sea level rise of 0.5–1.4 m by the end of the 21st century. Sea level, as it appears to an ecologist standing in an estuarine Spartina meadow or to an individual Spartina plant, is a function of both the absolute volume of water in the ocean (the eustatic sea level) and the vertical displacement of the land surface due to processes such as subsidence (Nuttle et al., 1997). The apparent sea level rise in an estuary that results from both ESLR and land subsidence is referred to as relative sea level rise (RSLR; Table 1). The dramatic effects of high rates of RSLR have already been observed in regions such as coastal Louisiana (Boesch et al., 1994) where high rates of land subsidence and insufficient sedimentation exacerbated rising eustatic sea levels and led to wetland loss rates of 60 km2/year in the 1980s and 1990s (Boesch et al., 1994). Regions with high rates of RSLR can serve as models for other estuarine systems as the rate of ESLR begins to accelerate. As an aside, in any discussion of wetland elevation relative to sea level, there is often some ambiguity regarding the terminology itself. The term “sea level rise” alone is ambiguous as it could refer to the absolute rise in sea level (i.e., ESLR) or the apparent rise caused by both ESLR and land subsidence. Similarly, land subsidence could include deep geologic subsidence, shallow subsidence caused by primary compaction and organic matter decomposition, or both. We define the terms, as used in this chapter (Table 1) and we encourage the use of them elsewhere.
Table 1
Definitions of some of the processes affecting wetland elevation relative to sea level
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Surface vertical accumulation of mineral and organic sediment, usually over some marker horizon. May also integrate processes occurring on and within the upper part of the marsh substrate (e.g., root growth and decomposition)a Primary compaction, decomposition, and dewatering that occurs in upper sediments (up to10 m)a Deep primary compaction, secondary compaction, and other processes such as geosynclinic downwarping and tectonic activityb Global sea level rise caused by changes in the volumes of glaciers and ice caps and by water density/temperature dependent relationshipsb Long-term, absolute vertical relationship between the land and the water surface. On the marsh surface, RSLR should be calculated as ESLR þ deep subsidence þ shallow subsidence. However, RSLR, measured using tidal gauge records, represents only ESLR þ deep subsidence = accretion – shallow subsidence – deep subsidence – ESLR or = accretion – RLSR
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Cahoon et al., 1995. Chen and Rybczyk, 2005.
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Estuarine wetland elevation, relative to sea level, is a function of numerous processes including mineral and organic matter accretion, sediment compaction, deep subsidence, and ESLR, all operating at different time scales (Figure 1). A number of studies have shown that estuarine wetlands can persist for long periods of time (thousands of years) in the face of rising sea levels when sediment accretion equals or exceeds the rate of land subsidence plus ESLR, as is the case for most wetlands worldwide under current rates of ESLR (Gornitz, 1995; Orson and Howes, 1992; Morris et al., 2002; Rybczyk and Cahoon, 2002). In practice, estuarine wetlands exist in a dynamic equilibrium between the forces that lead to their establishment and maintenance, such as sediment accretion, and the forces that lead to their deterioration such as increasing rates of ESLR and subsidence (Day et al., 1999). Changes in either side of the maintenance/deterioration equation could lead to changes in the inundation regime, anaerobic stress levels throughout the wetland, and shifts in habitat types. For example, an increase in inundation frequency and duration could lead to a shift in the relative distribution of vegetated Autochthonous organic matter production
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Figure 1 Processes that affect wetland elevation relative to sea level rise. Processes shown below the time line decrease elevation while those shown above the time line increase elevation. Processes shown in italics are commonly entered as forcing functions into existing wetland elevation models.
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habitats across a wetland (Warren and Niering, 1993; Kirwan and Murray, 2008) (e.g., shifts from areas dominated by Spartina patens to Spartina alterniflora). Over longer time periods, these types of changes would lead to conversion of more and more area to unvegetated mudflats. At the wetland–upland transition, there could be conversion of upland buffers to wetlands as wetlands “migrate” inland. However, in many cases, coastal wetlands are surrounded by development or steep habitats, such that migration would not be possible (Titus, 1991).
2. MEASURING P ROCESSES THAT AFFECT W ETLAND E LEVATION While many coastal wetlands accumulate sediment at a rate that keeps pace with current rates of ESLR, the focus of most managers presently, and especially in the future, is and will be on the loss of relative elevation and the submergence of tidal wetlands due to increased rates of ESLR. The potential for coastal wetland submergence has traditionally been determined by calculating a net accretion balance (Table 1). This is accomplished by comparing rates of vertical accretion to rates of RSLR (ESLR plus deep subsidence). Rates of accretion are typically estimated by measuring the accumulation of sediments, both organic and mineral, over some known and dated marker horizon such as feldspar clay, 137Cs, or 210Pb. Estimates of deep subsidence are usually based on long-term records from tide gauges that are mounted on stable piers, bridges, or pilings that extend through the shallow subsidence zone (and thus do not include shallow subsidence). A tidal gauge record spanning at least 18.6 years is required to factor out variations due to the moon’s nodal cycle (Turner, 1991). Typically, mean annual or monthly water levels are regressed against time to yield a rate of RSLR (Emery and Aubrey, 1991). To estimate the deep subsidence component of RSLR, current ESLR is subtracted from the water level rise recorded from the pier-mounted tidal gauge. ESLR is derived from the analysis of tide gauge data from coasts worldwide that are assumed to be geologically stable (Penland and Ramsey, 1990). Recent research has shown that the compaction in the upper 5–10 m of sediment contributes a shallow component that is often overlooked when calculating RSLR (Cahoon et al., 1995). This results in an underestimation of the accretion balance. Shallow subsidence is the result of primary sediment consolidation and the decomposition of organic matter (Table 1). It is an especially important process in coastal systems under stress (e.g., from flooding or salt) where belowground plant structures, such as roots and rhizomes, die and decompose, leading to rapid subsurface collapse (Cahoon et al., 2003). It is also an active process in coastal systems that are rapidly accreting as surface sediments consolidate (Day et al., 1999). For several reasons, these calculations to determine accretion balances must be viewed with caution, even if shallow subsidence is considered. First, short-term field measurements do not necessarily integrate long-term processes that affect wetland elevation, such as compaction, decomposition, and pulsing events such as stormrelated sediment deposition. Even programs that span a decade or more may not capture infrequent sediment deposition events from hurricanes or major river floods
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(Rybczyk and Cahoon, 2002). Second, these types of measurements may not take into account possible feedback mechanisms on the processes themselves. For example, a change in elevation typically alters flooding patterns that can, in turn, affect rates of sediment deposition, decomposition, and autogenic primary production. Over time, these measured rates change in a nonlinear fashion. Finally, the direct comparison of rates of elevation change to rates of RSLR can be problematic because of the large amount of uncertainty involving measurements of the processes affecting marsh elevation and sea level rise. For these reasons, a simulation model that considers all of the relevant processes over appropriate time scales, incorporates feedback mechanisms, and has methodologies for dealing with uncertainty (i.e., sensitivity analysis) can be a complimentary tool for examining the response of wetland elevation to increasing rates of sea level rise (Rybczyk and Cahoon, 2002).
3. TYPES OF MODELS Over the past 25 years, a variety of surface elevation models have been developed to address the relationship between sea level rise and coastal wetland elevation. They differ in the spatial scales that are considered and the processes that are simulated within the model, as opposed to processes that are input as forcing functions or not considered at all (Figure 1). A forcing function is an independent input variable that drives a model but is not modified by the model itself; it is a variable that is not simulated. Although there is overlap, we can divide surface elevation models into three groups: (1) landscape models that simulate processes over large regions (i.e., entire estuaries or coastlines), (2) geomorphic and ecogeomorphic models that simulate physical and ecological processes across a marsh platform or transect, and (3) zero-dimensional models that simulate the change in elevation at one point rather than across an entire marsh. In general, landscape models excel at simulating general trends at large spatial scales but often do not mechanistically simulate the processes that contribute to wetland elevation (in other words, many of the processes are input as forcing functions and are not simulated within the model). In one of the first efforts of this type, Park et al. (1989) developed a model (SLAMM2) that was used for 93 sites and predicted changes in land classes based on elevation, erosion, surrounding land type, accretion rates, and inundation. Land classes were based on classified Landsat data using 500 500 m cells. Elevation data on the same scale was obtained from USGS topographic maps. Modified versions of the SLAMM model are still used for assessing landscape scale impacts of sea level rise on coastal systems. Other types of landscape scale models are linked to complex and computer-intensive oceanographic hydrodynamic models that simulate changes in sea level, salinity, and sediments (Martin et al., 2000; Reyes et al., 2000). Changes in habitat type are simulated by simple switching functions that are dependent on these three parameters. At the other end of the spatial scale are the zero-dimensional wetland surface elevation models or “point” models that simulate change in wetland elevation for
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one specific point in a coastal wetland. Early zero-dimensional models tended to focus on either mineral deposition or organic sediment processes although there is now overlap between the two. The mineral sediment models, as might be expected, were first developed for minerogenic wetlands (e.g., San Francisco Bay and England). These models provide detailed mechanistic modeling of sediment transport and settling, but they have little information about organic matter processes. The organic matter process models focus on aboveground and belowground organic matter production, decomposition, and sediment compaction with most models focusing particularly on belowground dynamics. Both of these groups of models lack horizontal spatial articulation, but they excel at simulating site-specific processes in a mechanistic fashion. Spanning the gap between zero-dimensional models and regional-scale landscape models are the geomorphic models that simulate morphodynamics (i.e., sedimentation, channel development, and erosion) across a marsh platform (a two-dimensional model) or a marsh transect (a one-dimensional model). These models are said to be “ecogeomorphic” or “biogeomorphic” if they additionally consider the feedbacks between marsh vegetation and physical processes such as sedimentation and erosion. Since many of the processes that affect wetland elevation, such as sedimentation rates, vary across the marsh surface, these types of models have recently been used to examine the response of salt marshes and tidal flats to sea level rise (Kirwan and Murray, 2007). There is no fundamental reason why zero-dimensional models, and even some landscape scale models, could not also be considered “geomorphic”. They all simulate changes in marsh morphology at some scale (although at varying levels of detail). We divide them here as a matter of organizational convenience. In the following sections, we discuss the zerodimensional and geomorphic surface elevation models. Landscape models are covered in detail by Reyes (2009).
3.1. Zero-dimensional mineral sediment models These models focus on mineral sedimentation processes in coastal wetlands. The algorithm common to all of these models describes a negative feedback between marsh surface elevation and mineral sediment accretion; the higher the marsh elevation, with respect to sea level, the less it is flooded, and the less sediment is transported and deposited. The first of these models was developed by Krone (1987) for San Francisco Bay. The model was designed to evaluate the historic location of the mean high water datum by simulating changes in wetland elevation over time. Krone mechanistically simulated mineral accretion rates based on predicted periods of tidal inundation combined with assumptions concerning concentrations and settling rates for suspended sediments. The model was calibrated with 14 C data, and Krone found that the accretion rate increased with higher rates of sea level rise. Under constant rates of sea level rise, there was little change in the relative elevation of the wetland because of the feedback between relative elevation and the period of tidal inundation. The method that Krone developed is, to this day, the general approach that is used for mechanistically simulating mineral sediment dynamics in most wetland elevation models.
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Using a similar approach to Krone (1987), Allen (1990, 1995) developed two exploratory models of mineral sedimentation that extended the understanding of sediment dynamics in coastal wetlands. Allen modeled the development of a wetland rather than focusing on a mature wetland as in Krone’s model. He also modeled changes in suspended sediment caused by variations in the daily tidal range and seasonal fluctuations. Allen’s first model (Allen, 1990) simulated changes in wetland surface elevation over a 2,500-year time scale, with model predictions compared to data from Pethick (1981) for the age and elevation of a variety of developing wetlands. Allen used this exploratory model to show the link between forcing factors (including sea level rise and mineral and organic matter inputs) and sediment characteristics (mineral and organic content of the sediment and rates of accretion). With current rates of sea level rise, the model simulated a steep increase in wetland elevation in newly developing wetlands, with a gradual slowdown to a stable elevation below the level of extreme high tides. A later version of the model (Allen, 1995) added a variable sea level rise term and variable mineral inputs to simulate intercalated silt and peat sediments that are commonly found in northwest Europe. Neither of these models explicitly addressed impacts due to predicted increases in sea level rise; however, the models simulated the processes that are necessary for these predictions. French (1993) built upon the approach of Krone (1987) and Allen (1990) to develop a mass balance model of sediment accumulation and wetland surface elevation that was designed specifically to predict impacts of increased sea level rise for wetlands in Norfolk, UK (Figure 2). He calibrated the model based on the same relationships between wetland elevation and age, using revised data from
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Figure 2 In one of the first mineral sediment models designed specifically to examine the effects of increasing rates of sea level rise on vertical marsh growth and stability, French (1993) simulated net accretion balances (Table 1) for north Norfolk, UK, coastal marshes under different ESLR scenarios (adapted from French, 1993).
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Pethick (1981) over a 2,000-year time scale. Following model calibration, French used a series of constant and variable rates of sea level rise (ranging from 0.11 to 1.5 cm/year) to evaluate changes in wetland surface elevation over time. He concluded that by the year 2100, the wetland would be submerged only with the highest predicted rates of sea level rise; 1.5 cm/year. A constant increase in sea level rise of 1.0 cm/year caused a shift in relative elevation of approximately 50 cm by 2100 (compared to conditions under current sea level rise). Using the 1990 IPCC best estimate of sea level rise projections, the modeled wetland began losing elevation by about 2040 (accretion deficit); however, the decrease was slight with only moderate changes in elevation by 2100. Using a slightly different approach, Allen (1994) and Woolnough et al. (1995) developed simulation models that focused on hydrology and the delivery of mineral sediment to the wetland. This model tested the mode of sediment delivery in a hypothetical linear wetland: either over the front of the wetland (adjacent to a bay or sea) or via a network of perpendicular tidal creeks through the wetland (Allen, 1994). Allen (1994) compared model-generated sediment characteristics based on these two scenarios to actual conditions and confirmed that sediment delivery via tidal creeks was more realistic. Allen (1996) used a series of case studies with this model to evaluate sediment stratigraphy and wetland development for wetlands in the Severn Estuary. Woolnough et al. (1995) used a similar approach as Allen (1994) and developed a mathematical model of sediment mass transport and deposition to evaluate sedimentation processes over a 12-h tidal period. They developed the model using simplified sediments of a single grain size and expanded it to include a “population” of mixed sediments. The mixed sediment model confirmed sediment patterns that are found in real wetlands, with more deposition adjacent to water and the accumulation of finer sediments toward the upland edge of the wetland. As a validation exercise, Temmerman et al. (2003) used Krone’s model to simulate, and replicate, historic marsh accretion rates in the Scheldt Estuary, Belgium. Krone’s model (1987) assumed a constant suspended sediment concentration for all phases of a tidal cycle and for all tidal heights. However, Temmerman et al. (2003) showed that suspended sediment concentrations tend to increase with increasing tidal heights. By modifying Krone’s original model to reflect this positive relationship, they showed that simulated sedimentation rates to a marsh, summed over an entire tidal cycle, increased exponentially as a function of tidal height (as opposed to linearly when suspended sediment concentrations are held constant). Furthermore, simulated accretion in the estuary only matched historic records when this modification was employed.
3.2. Zero-dimensional organic sediment process models This type of zero-dimensional model mechanistically simulates organic matter accumulation dynamics in wetland sediments and belowground processes including organic matter production, decomposition, and sediment compaction (Figure 3). Mineral matter deposition also is usually simulated. Many of these models use a “cohort” approach, following changes in a given age class or cohort of sediment as
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Figure 3 A conceptual diagram of a generic organic sediment process model, using standard Odum energy systems diagramming (Odum and Odum, 2000). Rates of primary production, decomposition, and mineral matter accumulation are all functions of relative elevation. The arrow-shaped polygons represent the interaction of two or more processes. Arrows represent flow of material to or from one state variable to another.
it is buried to greater depths. This approach was first used by Morris and Bowden (1986), although they modeled nutrient dynamics not sedimentation processes or sea level rise. Morris and Bowden’s model used a mechanistic approach to simulate belowground decomposition, including separate labile and refractory pools. The use of a cohort approach for sediment modeling is useful because it is possible to generate modeled profiles of sediment characteristics that can be compared to data from actual cores (Figure 4). These data are commonly available from field studies of sediment accretion. In addition, it is easy to evaluate modelgenerated accretion rates by following the depth of burial for a cohort of a particular age or date. Even though Morris and Bowden (1986) did not explicitly model accretion rates, the cohort approach that they developed has become the basis for models of sediment organic matter processes. Chmura et al. (1992) developed a simple sediment cohort model that was specifically designed to simulate relative wetland elevation and stability under various sea level rise scenarios for Barataria Basin, Louisiana. However, this model made no distinction between organic and mineral matter inputs and, as a consequence, assumed a homogeneous sediment composition with depth.
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Additionally, inputs of sediment were modeled as a constant and were not a function of elevation. Only four cohorts were used in the model, so belowground changes in sediment characteristics were modeled on a coarse scale. The model was calibrated using sediment accretion rates based on 14C dating. Sensitivity analysis indicated that accretion rate and sea level rise had the greatest effect on relative wetland elevation. The model predicted critical rates of accretion for wetland stability, and maximum measured rates in Barataria Basin, Louisiana, were lower than the critical values, confirming the instability of these wetlands under the high rates of local subsidence. Callaway et al. (1996) used the cohort approach to model sediment processes in coastal wetlands in Mississippi and Great Britain. Their model used annual cohorts over a 300-year time period and was calibrated based on 137Cs and 210Pb dating, as well as vertical profiles of sediment characteristics (organic matter content, bulk density, and pore space). The use of multiple methods of calibration strengthens the overall reliability of the model. The model used a mechanistic approach for production and decomposition; however, the inputs for these processes were model calibrated rather than based on field data. A “model-calibrated” input (the decomposition rate of organic matter, for example) is an input parameter that is varied over many program runs until the model output (percent organic matter in the soil, for example) matches the observed field conditions. Model-calibrated
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parameters are often used when there are no actual, site-specific field measurements for that given parameter and should only be varied over a biologically reasonable range. The Callaway et al. (1996) model also simulated compaction as a function of the density of sediment above a given cohort. Because no universal compaction function was identified, the compaction algorithm was calibrated separately for each wetland under consideration (Callaway et al., 1996). Mineral inputs were modeled as a linear function of relative wetland elevation, with maximum rates at elevations below mean sea level and no mineral inputs above mean high water. Callaway et al. (1996) evaluated changes in elevation and predicted accretion rates over a series of constant increases in sea level ranging from 0.05 to 0.5 cm/year. The relative elevation of the wetland was stable at rates of 0.1 cm/year, and the model predicted a loss in elevation of approximately 1 m with increases in sea level of 0.5 cm/year (given a 6-m tidal range). Sensitivity analyses indicated that the model-generated accretion rates were most affected by changes in pore space (i.e., compaction), mineral matter deposition, initial elevation, ESLR, and belowground organic matter production. Rybczyk et al. (1998) developed a simulation model to address sediment dynamics in a subsiding forested wetland in the coastal zone of Louisiana that received wastewater effluent. This model extended the work of the previous organic-based models by modifying and adding new subroutines to address the characteristics of forested wetlands that are not associated with the annual systems that previous models simulated (e.g., perennial aboveground biomass). It also included a primary production submodel that simulated organic matter production as a function of elevation. Organic matter production was entered as a forcing function in both Morris and Bowden (1986) and Callaway et al. (1996) model. Decomposition processes were simulated more mechanistically in this model, because decomposition was likely to be affected by the addition of wastewater. Modified versions of this model also have been used to predict the vulnerability of coastal salt marshes to increasing rates of ESLR in the Venice Lagoon (Day et al., 1999) and salt marshes in Louisiana (Rybczyk and Cahoon, 2002) and to predict sediment collapse in a mangrove swamp decimated by Hurricane Mitch (Cahoon et al., 2003). Morris et al. (2002) developed, and later generalized (Morris, 2006), a noncohort theoretical feedback model that predicts change in salt marsh elevation in response to sea level rise. Like earlier sediment models (i.e., Krone, 1987; French, 1993), sediment deposition increases with increasing rates of sea level rise and concomitant flooding and approaches zero as marsh elevation nears the level of the highest tides. However, the model also simulates the positive effects of aboveground plant biomass and productivity on sediment trapping and elevation change. The model can be used to predict the rate of sea level rise for which salt marsh elevation at a given location will be optimal for plant growth under various rates of sediment loading. Their simulations revealed that the optimal rate of sea level rise maximizes sediment accretion and productivity but also represents the upper limit of flooding tolerance for a given plant species. The model predicted that their study site, a S. alterniflora marsh in South Carolina, could tolerate a rate of RSLR of up to 1.2 cm/year.
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The organic-based models have added to the understanding of coastal wetland sediment dynamics by highlighting the importance of belowground processes to accretion balances. The models indicate that compaction (both shallow and deep) and belowground production are important in determining overall accretion rates. Because of feedback between compaction and production in these models, changes in the relative elevation of the wetland are not linearly related to mineral inputs. The models indicate that wetland elevation is especially sensitive to predicted rates of ESLR and to estimated rates of subsidence, both of which are modeled as forcing functions and are outside of the control and influence of wetland managers. For areas with high rates of local subsidence, such as the Louisiana, and the Po and Ebro River deltas, the models predict immediate problems due to the imbalance of sediment accretion and RSLR (Day et al., 1999). The organic sediment models lack spatial approaches, so large-scale predictions of impacted acreage are not possible.
3.3. Geomorphic models These models emphasize the physical processes that affect marsh surface topography and typically include mechanistic algorithms that describe wave energy, tidal currents, shear stress, and sediment erosion. Ecogeomorphic models also consider how marsh biota influences these physical processes, especially rates of deposition and erosion. For example, geomorphic and ecogeomorphic models have been used to simulate the cross sectional evolution of a salt marsh channel (D’Alpaos et al., 2006), the evolution of tidal flats and salt marshes in the Venice Lagoon (Fagherazzi et al., 2006; Defina et al., 2007), three-dimensional channel network development and tidal marsh accretion (Kirwan and Murray, 2007), and the response of salt marshes to sea level rise (D’Alpaos et al., 2007; Kirwan and Murray, 2008). While most elevation models describe a smooth elevational transition from open water to high marsh, or vice-versa, as sedimentation and vegetative growth dynamics change, Fagherazzi et al. (2006) observed that, in some shallow tidal basins (micro- to mesotidal systems where resuspension of bottom sediments is dominated by wind waves and not tidal currents), elevation is bi-modal; tidal flats exist at lower elevations, salt marshes at the higher elevations but no landforms are found at intermediate elevations. They developed a numerical conceptual model where change in marsh elevation over time was a function of sediment porosity, rates of erosion and sedimentation, and bottom shear stress. The model revealed that there were only two stable equilibrium conditions for Venice Lagoon: a stable tidal flat if the annual rate of deposition was smaller than the erosion rate or an emergent marsh if the annual deposition was higher than the erosion rate. Intermediate elevations would only be stable if the deposition rate exactly equaled the rate of erosion. In one of the most recent efforts to simulate the effects of sea level rise on marsh wetland elevation, Kirwan and Murray (2008) developed an exploratory, ecogeomorphic model for Westham Island, a salt marsh/tidal flats system in the Fraser River Delta, British Columbia. The model utilized a modification of the deposition algorithm from Morris et al. (2002) that accounted for distance from tidal channels, simulated wave height and wave-induced erosion, and linked marsh elevation to a
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Figure 5 Simulated loss of vegetated area and annual productivity (indicated here as total biomass) for Westham Island, BC under a RSLR (ESLR plus local deep subsidence) scenario of 1.036 m in 100 years.While loss of productivity continues to increase after 2060, loss of vegetated area begins to stabilize. This is because the most productive high marshes are converted to less productive transitional and low marshes (adapted from Kirwan and Murray, 2008).
simple plant productivity switching function (i.e., plant productivity decreased with decreasing elevation). It also incorporated a sediment deposition/elevation/plant productivity feedback function (plant productivity affects sediment deposition rates which, in turn, affect elevation, which influences plant productivity). Output was generated as two-dimensional elevation and vegetation zone maps of the island under various sea level rise scenarios. They found that a RSLR of 1.036 m in the next 100 years resulted in a loss of approximately 38% of total vegetated area (Figure 5).
4. FUTURE D IRECTIONS FOR M ODEL I MPROVEMENT Certainly different surface elevation models are built with different objectives in mind. The intent of a landscape model may be to elucidate coast-wide wetland loss trends over the next 100 years while the objectives of a two-dimensional geomorphic model may be to accurately simulate channel development for a single marsh under vegetated and unvegetated conditions. Although each type of model has its strengths and weaknesses, it is valuable to identify general improvements for the next generations of models. We propose the following improvements to current models: (1) address data gaps, (2) integrate existing models (3) further link vegetation and sedimentation processes, and (4) improve horizontal spatialization. Below we discuss what is necessary to make these improvements and what specific advancements each of these improvements will allow.
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4.1. Data gaps For all models, certain processes have only been model-calibrated because data, at least at the site-specific level, do not exist for these processes. To mechanistically simulate ecosystem processes, we must collect field data to initialize, calibrate and validate these components of the models. In particular, there is little information about many of the belowground processes that contribute to shallow subsidence in coastal wetlands, including the following. 4.1.1. Compaction Shallow compaction rates are poorly studied in wetland sediments. A study (Pizzuto and Schwendt, 1997) of coastal wetland sediments in Delaware has modeled compaction in freshwater peat and salt marsh sediments. This model looked at sediments over a 6,000-year time scale; however, their approach could be incorporated into short-term models that are more applicable to concerns about predicted increases in sea level. Besides this modeling effort, most work investigating compaction has been done with freshwater peat and has focused on engineering issues (Landva and Pheeney, 1980; Hobbs, 1986) or it has been done qualitatively and data are not applicable to calibration of simulation models. The data that are needed to improve the model are both short- and long-term rates of sediment compaction, including an understanding of how these rates vary depending on the amount and type of sediment (bulk density, organic content) that accumulate on top of a particular sediment layer. 4.1.2. Belowground production To model changes in organic matter accumulation throughout the sediment column, we need detailed belowground production estimates; however, data on belowground production, and root distribution with depth, in coastal wetlands are limited, compared to information on aboveground production. Data are needed for belowground production by depth, for different plant species across a wetland and under different growing conditions. The few studies that have looked at production with depth have been done at a coarse scale, so we do not have data to adequately describe changes in belowground production by depth. In addition, almost no work has been done to evaluate changes in production at different elevations, under different environmental conditions (e.g., effects of hydrology or salinity), or with different plant communities across the wetland. These types of data are needed in order to improve current modeling efforts of belowground organic matter accumulation. 4.1.3. Long-term decomposition As with belowground production, problems associated with sampling have limited estimates of belowground decomposition rates. While we do have some basic data on belowground decomposition (van der Valk and Attiwill, 1983; Schubauer and Hopkinson, 1984; Buth, 1987), we lack detailed information about changes in rates
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with depth of burial (van der Valk and Attiwill, 1983; Hemminga et al., 1988) and under different rates of flooding (Hemming et al., 1991). A further limitation for the decomposition data is that all currently available information for belowground decomposition rates concerns short-term (mostly up to 5 years) not long-term rates. However, slight changes in the long-term rates of decomposition can lead to large changes because these rates are integrated throughout the sediment column. Given the importance of decomposition to overall sediment dynamics, we need to develop a method to estimate long-term rates.
4.2. Integrating models A second advancement for the development of wetland elevation models will be to incorporate the best aspects of currently available models into a single integrated approach. For example, many of the mineral sediment models (Allen, 1990, 1994, 1995; French, 1993) and cohort-based models (Callaway et al., 1996) are “deposition-only” models and do not mechanistically consider physical erosion. Some of the newer geomorphic models, on the other hand, do simulate erosion (Fagherazzi et al., 2006; Marani et. al., 2007; Kirwan and Murray, 2007), but they ignore the belowground processes that are mechanistically simulated in organic cohort models. One of the difficulties in incorporating these models into an integrated model is that they simulate processes over a series of different time scales, from daily tidal flooding regimes (for mineral sediment dynamics) to long-term decomposition rates (for organic matter).
4.3. Improved linkage between sediment models and vegetation As the relative elevation of a wetland changes, vegetation communities are likely to change, and this shift could cause large changes in sediment dynamics (D’Alpaos et al., 2007). Vegetation shifts due to changes in elevation have been described in sediment core analyses by Warren and Niering (1993) and are implied in many developmental studies of salt marshes (Redfield, 1972; Orson and Howes, 1992). Although it is clear that elevation is not the only determinant of vegetation community, it is a dominant factor. Along these lines, Rybczyk et al. (1998) made connections between above- and belowground production and sediment surface elevation in their models of forested wetlands. In models of salt marsh sediment dynamics (Day et al., 1999), both the dominant vegetation community and the modeled annual productivity change with 10-cm increments in elevation. Recent ecogeomorphic models (D’Alpaos et al., 2006, 2007; Kirwan and Murray, 2007, 2008; Marani et al., 2007) have also been used to simulate the effects of vegetation on sediment erosion and wetland resilience to sea level rise. Kirwan and Murray (2007) developed a three-dimensional model of tidal marsh accretion and channel network development that included a biomass productivity state variable that was a function of water depth. Their model showed that, in the absence of vegetation and under a moderate sea level rise scenario, intertidal platforms would degrade to subtidal elevations due to tidal channel expansion and erosion. These models consider plant composition and
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productivity to be a function of elevation (or water depth) only. As other factors affecting plant growth are quantified (e.g., salinity and sediment composition), they should be incorporated into models.
4.4. Spatialization Many of the processes that affect wetland elevation, such as sedimentation rates, vary across the marsh platform (e.g., proximity to tidal channels). To address this, several of the previously discussed geomorphic models have already spatialized the zero-dimensional models of Krone (1987), Allen (1990), and French (1993) to the scale of a marsh platform although they do not consider belowground processes. Thus far, these geomorphic models have not been extended to the landscape scale (i.e., entire estuaries or coastlines). Spatializing over larger scales will allow for the coupling of mechanistic morphodynamic models to regional hydrodynamic models that are capable of simulating sediment delivery and salinity levels, both of which are factors likely to be affected by climate change. The biggest difficulty with generating mechanistic, geomorphic spatial models will be in the collection of the field data required to calibrate and validate them. For example, the data gaps that we have discussed above concern calibration and validation of a unit model for a single point within a wetland. When the model is spatialized, data shortcomings will likewise be compounded.
5. C ONCLUSIONS Simulation models have been used effectively to integrate our understanding of wetland sediment dynamics. Three different approaches have been used to address sediment dynamics in coastal wetlands, including landscape models, ecogeomorphic modeling, and zero-dimensional simulations of mineral and organic sediment processes. These models have been successful; however, the combination of process-based models with spatial models for estimating large-scale impacts due to predicted increases in eustatic sea level has been limited. In order to improve future models, we must address data gaps which exist for the present simulation models, integrate the strengths of existing models, incorporate interactions between vegetation and sediment dynamics, and spatialize the integrated process model so that more realistic predictions of future impacts from sea level rise can be made.
REFERENCES Allen, J.R.L., 1990. Salt-marsh growth and stratification: a numerical model with special reference to the Severn Estuary, southwest Britain. Mar. Geol. 95, 77–96. Allen, J.R.L., 1994. A continuity-based sedimentological model for temperate-zone tidal salt marshes. J. Geol. Soc. Lond. 151, 41–49. Allen, J.R.L., 1995. Salt-marsh growth and fluctuating sea level: implications of a simulation model for Flandrian coastal stratigraphy and peat-based sea-level curves. Sediment. Geol. 100, 21–45.
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Allen, J.R.L., 1996. Shoreline movement and vertical textural patterns in salt marsh deposits: implications of a simple model for flow and sedimentation over tidal marshes. Proc. Geol. Assoc. 107, 15–23. Boesch, DF., Josselyn, M.N., Mehta, A.J., Morris, J.T., Nuttle, W.K., Simenstad, C.A., Swift, D.J., 1994. Scientific assessment of coastal wetland loss, restoration and management in Louisiana. J. Coast. Res. Spec. Issue 20, 1–103. Buth, G.J.C., 1987. Decomposition of roots of three plant communities in a Dutch salt marsh. Aquat. Bot. 29, 123–138. Cahoon, D.R., Hensel, P., Rybczyk, J.M., McKee, K., Proffitt, C.E., Perez, B.C., 2003. Mangrove peat collapse following mass tree mortality: implications for forest recovery from Hurricane Mitch. J. Ecol. 91, 1093–1105. Cahoon, D.R., Reed, D.J., Day Jr., J.W., 1995. Estimating shallow subsidence in microtidal salt marshes of the southeastern United States: Kaye and Barghoorn revisited. Mar. Geol. 128, 1–9. Callaway, J.C., Nyman, J.A., DeLaune, R.D., 1996. Sediment accretion in coastal wetlands: A review and a simulation model of processes. Curr. Top. Wetl. Biogeochem. 2, 2–23. Chen, Z., Rybczyk, J.M., 2005. Coastal Subsidence. In: Schwartz, M. (Ed.), Encyclopedia of Coastal Science. Kluwer Academic Publishers, Dordrecht, pp. 359–362. Chmura, G.L., Costanza, R., Kosters, E.C., 1992. Modelling coastal marsh stability in response to sealevel rise: a case study in coastal Louisiana, USA. Ecol. Model. 64, 47–64. D’Alpaos, A., Lanzoni, S., Marani, M., Rinaldo, A., 2007. Landscape evolution in tidal embayments: modeling the interplay of erosion, sedimentation, and vegetation dynamics. J. Geophys. Res. 112. doi:10.1029/2006JF000537. D’Alpaos, A., Lanzoni, S., Mudd, S.M., Fagherazzi, S., 2006. Modeling the influence of hydroperiod and vegetation on the cross-sectional formation of tidal channels. Estuar. Coast. Shelf Sci. 69, 311–324. Day, J.W., Rybczyk, J.M., Scarton, F., Rismondo, A., Are, D., 1999. Soil accretional dynamics, sealevel rise and the survival of wetlands in Venice Lagoon: a field and modeling approach. Estuar. Coast. Shelf Sci. 49, 607–628. Defina, A., Carniello, L., Fagherazzi, S., D’Alpaos, L., 2007. Self-organization of shallow basins in tidal flats and salt marshes. J. Geophys. Res. 112.doi:10.1029/2006JF000550. Emery, K.O., Aubrey, D.G., 1991. Sea Levels, Land Levels, and Tide Gauges. New York: SpringerVerlag. Fagherazzi, S., Carniello, L., D’Alpaos, L., Defina, A., 2006. Critical bifurcation of shallow microtidal landforms in tidal flats and salt marshes. Proc. Natl. Acad. Sci. U. S. A. 102 (22), 8337–8341. French, J.R., 1993. Numerical simulation of vertical marsh growth and adjustment to accelerated sealevel rise, North Norfolk, U.K. Earth Surf. Process. Landf. 18, 63–81. Gornitz, V., 1995. Sea-level rise: A review of recent past and near-future trends. Earth Surf. Process. Landf. 20, 7–20. Hemming, M.A., de Leeuw, J., de Munck, W., Koutstaal, B.P., 1991. Decomposition in estuarine salt marshes: the effect of soil salinity and soil water content. Vegetatio 94, 25–33. Hemminga, M.A., Kok, C.J., de Munck, W., 1988. Decomposition of Spartina anglica roots and rhizomes in a salt marsh of the Westerschelde Estuary. Mar. Ecol. Prog. Ser. 48, 175–184. Hobbs, N.B., 1986. Mire morphology and the properties and behaviour of some British and foreign peats. Q. J. Eng. Geol. 19, 7–80. Kirwan, M.L., Murray, A.B., 2007. A coupled geomorphic and ecological model of tidal marsh evolution. Proc. Natl. Acad. Sci. 104, 6116–6122. Kirwan, M.L., Murray, A.B., 2008. Ecological and morphological response of brackish tidal marshland to the next century of sea-level rise: Westham Island, British Columbia. Glob. Planet. Change 60, 471–486. Krone, R.B., 1987. A method for simulating historic marsh elevations. In: Kraus, N.C. (Ed.), Coastal Sediments ’87: vol. I. American Society of Civil Engineers, New York, pp. 316–323. Landva, A.O., Pheeney, P.E., 1980. Peat fabric and structure. Can. Geotech. J. 17, 416–435. Marani, M., D’Alpaos, A., Lanzoni, S., Carniello, L., Rinaldo, A., 2007. Biologically-controlled multiple equilibria of tidal landforms and the fate of the Venice Lagoon. Geophys. Res. Lett. 34. doi:10.1029/2007GL030178.
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Martin, J.F., White, M.L., Reyes, E., Kemp, G.P., Day, J.W., Mashriqui, H., 2000. Evaluation of coastal management plans with a spatial model: Mississippi delta, Louisiana, USA. Environ. Manage. 26, 117–129. Meehl, G.A., Stocker, T.F., Collins, W.D., Friedlingstein, P., Gaye, A.T., Gregory, J.M., Kitoh, A., Knutti, R., Murphy, J.M., Noda, A., Raper, S.C.B., Watterson, I.G., Weaver, A.J., Zhao, Z.C., 2007. Global climate projections. In: Solomon, S., Qin, D., Manning, M., Chen, Z., Marquis, M., Avery, K.B., Tignor, M., Miller, H.L. (Eds.), Climate Change 2007: The Physical Science Basis. Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge. Morris, J.T., 2006. Competition among marsh macrophytes by means of geomorphological displacement in the intertidal zone. Estuar. Coast. Shelf Sci. 69, 395–402. Morris, J.T., Bowden, W.B., 1986. A mechanistic, numerical model of sedimentation, mineralization and decomposition for marsh sediments. Soil Sci. Soc. Am. J. 50, 96–105. Morris, J.T., Sundareshwar, P.V., Nietch, C.T., Kjerfve, B., Cahoon, D.R., 2002. Response of coastal wetlands to rising sea-levels. Ecology 83 (10), 2869–2877. Nuttle, W.K., Brinson, M.M., Cahoon, D., Callaway, J.C., Christian, R.R., Chmura, G.L., Conner, W.H., Day, R.H., Ford, M., Grace, J., Lynch, J., Orson, R.A., Parkinson, R.W., Reed, D., Rybczyk, J.M., Smith III., T.J., Stumpf, R.P., Williams, K., 1997. Conserving coastal wetlands despite sea-level rise. EOS 78 (25), 257–261. Odum, H.T., Odum, E.C., 2000. Modeling for all Scales. San Diego, CA: Academic Press. Orson, R.A., Howes, B.L., 1992. Salt marsh development studies at Waquoit Bay, Massachusetts: Influence of geomorphology on long-term plant community structure. Estuar. Coast. Shelf Sci. 35, 453–471. Park, R.A., Trehan, M.S., Mausel, P.W., Howe, R.C., 1989. The effects of sea-level rise on U.S. coastal wetlands. In: Tirpak, D.A., Smith, J.B. (Eds.), The Potential Effects of Global Climate Change on the United States: Appendix B, Sea-level Rise. U.S. Environmental Protection Agency, EPA-230-05-89-052, Washington, DC, pp. 1–55. Penland, S., Ramsey, K.E., 1990. Relative sea-level rise in Louisiana and the Gulf of Mexico: 1908–1988. J. Coast. Res. 6 (2), 323–342. Pethick, J.S., 1981. Long-term accretion rates on tidal salt marshes. J. Sediment. Petrol. 51 (2), 521–577. Pizzuto, J.E., Schwendt, A.E., 1997. Mathematical modeling of autocompaction of a Holocene transgressive valley-fill deposit, Wolfe Glade, Delaware. Geology 25 (1), 57–60. Rahmstorf, S., 2007. A semi-empirical approach to projecting future sea-level rise. Science 315, 368–370. Redfield, A.C., 1972. Development of a New England salt marsh. Ecol. Monogr. 42, 201–237. Reed, D.J., 1995. The response of coastal marshes to sea-level rise: survival or submergence. Earth Surf. Process. Landf. 20, 39–48. Reyes, E., 2009. Wetland landscape spatial models. In: Perillo, G.M.E., Wolanski, E., Cahoon, D.R., Brinson, M.M. (Eds.), Coastal Wetlands: An Integrated Ecosystem Approach. Elsevier Science, Amsterdam, pp. 885–908. Reyes, E., White, M.L., Martin, J.F., Kemp, G.P., Day, J.W., 2000. Landscape modeling of coastal habitat change in the Mississippi delta. Ecology 81, 8–22. Rybczyk, J.M., Cahoon, D.R., 2002. Estimating the potential for submergence for two wetlands in the Mississippi River Delta. Estuaries 25 (5), 985–998. Rybczyk, J.M., Callaway, J., Day, J.W., 1998. A relative elevation model (REM) for a subsiding coastal forested wetland receiving wastewater effluent. Ecol. Model. 112 (1), 23–44. Schubauer, J.P., Hopkinson, C.S., 1984. Above- and belowground emergent macrophyte production and turnover in a coastal marsh ecosystem, Georgia. Limnol. Oceanogr. 29 (5), 1052–1065. Temmerman, S., Govers, G., Meire, P., Wartel, S., 2003. Modelling long-term tidal marsh growth under changing tidal conditions and suspended sediment concentrations, Scheldt estuary, Belgium. Mar. Geol. 193, 151–169. Titus, J.G., 1991. Greenhouse effect and coastal wetland policy: How Americans could abandon an area the size of Massachusetts at minimum cost. Environ. Manage. 15 (1), 39–58.
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Turner, R.E., 1991. Tide gauge records, water level rise, and subsidence in the Northern Gulf of Mexico. Estuaries 14 (2), 139–147. van der Valk, A.G., Attiwill, P.M., 1983. Above- and below-ground litter decomposition in an Australian salt marsh. Aust. J. Ecol. 8, 441–447. Warren, R.W., Niering, W.A., 1993. Vegetation change on a northeast tidal marsh: Interaction of sea-level rise and marsh accretion. Ecology 74, 96–13. Woolnough, S.J., Allen, J.R.L., Wood, W.L., 1995. An exploratory numerical model of sediment deposition over tidal salt marshes. Estuar. Coast. Shelf Sci. 41, 515–543.
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C H A P T E R
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S ALT M ARSH –M ANGROVE I NTERACTIONS IN A USTRALASIA AND THE A MERICAS Neil Saintilan, Kerrylee Rogers, and Karen McKee
Contents 1. Introduction 2. Distribution/Geomorphic Settings – Where do Mangrove and Salt Marsh Coexist? 2.1. Mangrove distribution 2.2. Salt marsh distribution 2.3. Coexisting mangrove and salt marsh 3. Long-Term Dynamics 3.1. Tropical northern Australia 3.2. Southeastern Australia 3.3. Western Atlantic–Caribbean Region 4. Recent Interactions 4.1. Air photographic evidence of mangrove–salt marsh dynamics in SE Australia 4.2. Saltwater intrusion in Northern Australia 4.3. Western Atlantic–Gulf of Mexico 5. Stressors Controlling Delimitation of Mangrove 5.1. Geomorphic and hydrological controls 5.2. Climatic controls 5.3. Physicochemical factors 5.4. Biotic interactions 6. Conclusions References
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1. INTRODUCTION In many places in the world mangroves and salt marsh coexist and competitively interact. These interactions can be studied over a range of timescales. Stratigraphic and palynological evidence has been used to reconstruct distribution of mangrove and salt marsh communities over geological timescales with the greatest clarity emerging from the Holocene. At this scale, interactions between mangroves Coastal Wetlands: An Integrated Ecosystem Approach
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and salt marshes are governed by geomorphic processes, most notably patterns of sedimentation following the postglacial marine highstand. More recently, over the historical time period, archival air photographs, and other remotely sensed images have been used to study changes in distribution of these communities over wide geographic areas. In SE Australia, the technique has revealed a consistent replacement of salt marsh by mangrove over the previous five decades. Mangroves compete with salt marshes at the limits of their physiological tolerance. At their latitudinal extreme, the vigor of mangrove growth is constrained by cold, and survival may be dictated by frost. The landward limit of mangroves in the intertidal zone is constrained by the often severe environment of the upper intertidal zone, where low soil moisture and high salinity may prohibit seedling establishment and growth. This chapter reviews a range of environmental variables: geomorphic, hydrologic, climatic, physicochemical, and biotic, which governs mangrove and salt marsh interactions. An emerging theme is the potential importance of climate change as a key control on dynamics: mangrove competitive advantage may well be enhanced by higher temperatures, mangroves being tropical plants and salt marshes predominantly temperate and sub-Arctic, increasing in diversity with increasing latitude (Saenger et al., 1977). Furthermore, higher sea levels could promote the landward encroachment of mangrove into salt marsh. The changing distribution of mangrove and salt marsh may serve as an important indicator of climate change impacts, a sentinel of change for the broad range of ecosystem services dependent on these habitats.
2. D ISTRIBUTION /GEOMORPHIC SETTINGS – W HERE DO M ANGROVE AND S ALT M ARSH C OEXIST ? 2.1. Mangrove distribution Mangroves are the typical intertidal vegetation of sheltered tropical and subtropical coastlines. However, some mangrove species will readily occupy sheltered coastlines at lower temperate latitudes (Saenger, 2002). The global distribution of mangroves has been examined on numerous occasions (Chapman, 1977; Duke et al., 1998a; Duke, 2006; Figure 1) and is primarily limited by the physiological tolerance of mangrove species to low temperature. Other factors, such as the availability of suitable habitat and climate, dispersal and establishment of propagules, continental drift, and tectonic events are also important (Duke et al., 1998a). Mangroves tend to be restricted to coastlines where mean air temperatures of the coldest month are higher than 20C and the seasonal range is not greater than 10C (Walsh, 1974; Chapman, 1977, from Duke et al., 1998a), which appears to correlate with the 20C isotherm for seawater (Duke et al., 1998a). In the Americas, distribution of mangroves roughly corresponds to the sea-surface winter isotherm of 20C, except for the North Atlantic coast where the northern limit corresponds to the 27C isotherm (i.e., the isotherm of the annual average for monthly maximal sea temperature where areas poleward are always cooler) (see Davy
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and Costa, 1992). The cold Humboldt and Falkland (Malvinas) currents influence sea temperatures along the western and eastern coasts of South America so that the winter 20C isotherm occurs at different latitudes (Figure 1). Similarly, the distribution of mangroves on the west coast of Africa is more restricted than on the east coast due to the different latitudinal positions of the 20C isotherm (Figure 1). The reduced occurrence of mangroves on the western coasts of Africa and South America also coincides with the limits of arid regions (Saenger, 2002) and implies that the development of mangroves in these regions is limited by aridity in addition to temperature (Saenger and Moverly, 1985; Smith and Duke, 1987). However, latitude and its relationship with low temperatures and/or the occurrence of frosts are the primary explanation for the latitudinal distribution of mangroves in the Americas (Sherrod and McMillan, 1985; McMillan and Sherrod, 1986). Mangroves thus extend from about 31N (Baja California, Peinado et al., 1994) to 3450 S (Peru, Clusener and Breckle, 1987) on the Pacific coast and from about 29N (Indian River Lagoon, Florida, USA) to 28300 S (Laguna, Brazil, (Schaeffer-Novelli et al., 1990) on the Atlantic coast. The northernmost mangroves in this region occur in Bermuda (32180 N), off the east coast of the United States, due to the warm waters of the Gulf Stream and absence of freezing temperatures; species found there include Rhizophora mangle, Avicennia germinans, and Conocarpus erectus (Thomas, 1993). In contrast, mangroves historically extended only to Ponce de Leon Inlet (29040 N) on the Atlantic coast of Florida and the Mississippi River delta in the Gulf of Mexico (29120 N) where they are periodically killed by freezing temperatures. Recent reports indicate northernmost individuals of R. mangle at 29400 N (Fort Matanzas, FL, USA) (Zomlefer et al., 2006) and of A. germinans at 3001.0120 N (Indian River Lagoon, FL, USA) (I.C. Feller, pers. comm.).
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Exceptions to the distribution pattern of mangrove in waters cooler than 20C also occur around Australia and the North Island of New Zealand. In these locations, the species of mangrove with the greatest latitudinal extent is Avicennia marina, with its distribution extending to Corner Inlet (38450 S) in Southeastern Australia. Explanations for the southerly distribution of A. marina in Australia include transmission by ocean currents or that they are relict populations that had greater poleward distributions in the past (MacNae, 1966; Duke et al., 1998a). The latter explanation is more widely accepted, particularly since A. marina var. australisica is genetically distinct from more tropical varieties, such as var. marina and var. eucalyptifolia (Duke et al., 1998a,b). A. marina var. australasica is the only mangrove species found in New Zealand, and is restricted to the north island with a latitudinal limit at Kutarere (38030 S) (Lange and Lange, 1994). Like its Australian counterpart, its southern distribution does not appear to relate to climatic factors (Lange and Lange, 1994) and may be a relict of greater poleward distributions (Mildenhall and Brown, 1987; Saenger, 2002). Due to the successful establishment of A. marina in more southerly locations, it has been the latitudinal distribution of mangroves in New Zealand is hypothesized to be in disequilibrium with climatic factors. Furthermore, low current velocities and large distances between suitable habitats inhibit the distribution of mangrove to the same extent observed in Australia (Lange and Lange, 1994).
2.2. Salt marsh distribution In contrast, salt marsh may occur on many of the world’s shorelines (Adam, 1990, Mendelssohn and McKee, 2000). Information about the global distribution of salt marsh species is poor and extensive detailed mapping of salt marsh has not been undertaken for some time (Chapman, 1977; Long and Mason, 1983). At a global scale, salt marsh establishes on shorelines where mangrove establishment is precluded or development is limited (Kangas and Lugo, 1990). For this reason salt marshes are most common in temperate, sub-Arctic, and Arctic zones (Long and Mason, 1983; Mitsch and Gosselink, 2000; Mendelssohn and McKee, 2000). In North America, salt marshes occur from the southern boundary of Central America to Alaska and Canada, with greatest development in the temperate United States (Mendelssohn and McKee, 2000). The northernmost limits of salt marsh on that continent occur along the Arctic Ocean at 70N latitude and extend southward along the Pacific shoreline of Alaska, Canada, the continental United States, Mexico, and Central America. Salt marsh also occurs along the shoreline of Hudson Bay (55N) and the Atlantic coasts of Canada and the United States, extending along the Gulf of Mexico into Central America where the distribution overlaps with that of mangrove vegetation. Salt marsh species can be found in Belize (16–17N), Guyana (6–8N), Peru/Ecuador (0–5S), Brazil (0–32S), Argentina (34–51S), and Chile (40–52S) (Table 2 in Costa and Davy, 1992). Salt marsh vegetation can be classed into six biogeographical types: Arctic, Boreal, Temperate, West Atlantic, Dry Coast, and Tropical (Adam, 1990). A description of salt marsh vegetation in North America using this classification can be found in Mendelssohn and McKee (2000). Salt marshes in the Arctic and sub-Arctic regions are
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dominated by species such as Puccinellia phryganoides and Carex spp. Boreal salt marsh along the Atlantic coast is characterized by Spartina patens, S. alterniflora in the low marsh, whereas Boreal salt marsh in the Pacific–Oceanic region, including British Columbia, contain Triglochin maritima, Salicornia virginica, and Distichlis spicata. Temperate salt marsh on the Atlantic coast is dominated by S. alterniflora, which extends southward into northern Florida. Other temperate species along the Atlantic coast include Juncus balticus (Boreal), J. gerardii (Boreal to Temperate), and J. roemerianus (Temperate), which also occurs along the Gulf of Mexico. Temperate salt marsh on the Pacific coast contains Spartina foliosa, Salicornia virginica, and D. spicata. In tropical zones, species such as S. patens, Spartina spartinae, D. spicata can occur, but may be replaced in hypersaline settings by succulents such as Batis maritima, Borrichia frutescens, Suaeda maritima, and Sesuvium portulacastrum.
2.3. Coexisting mangrove and salt marsh Salt marsh typically occurs in settings where mangrove development is limited (West, 1977; Kangas and Lugo, 1990) enabling salt marsh to establish. Extensive coexisting mangrove and salt marsh communities can be found in the temperate regions of Australia, New Zealand, and southern continental United States. Salt marsh may also establish behind mangrove communities within tropical and subtropical climates where rainfall is low and soil salinities in these areas become hypersaline (Chapman, 1977; Long and Mason, 1983). In Mexico, Central America, and Florida, for example, salt marsh may occur on the margins of mangrove forests (either colonizing seaward mudflats or the saline soils on landward edges), within mangrove woodlands with more open canopies, or in disturbed areas (West, 1977; Lopez-Portillo and Ezcurra, 1989). A distinction must be made between salt marsh vegetation typical of the low intertidal zone and that restricted to the “high marsh”, that is, areas in the upper intertidal range that are characterized by dryer, more saline conditions. More specifically, it is necessary to identify salt marsh and mangrove physiographic equivalents when considering interactions between coexisting mangrove and salt marsh vegetation and future changes in climate and sea level. For example, in North America, extensive stands of the temperate low marsh dominant S. alterniflora can be found as far south as subtropical Florida, Louisiana, and Texas. Although S. alterniflora can be found throughout Latin America, its occurrence is infrequent and mainly limited to mangal fringes (Costa and Davy, 1992). Other salt marsh species such as S. spartinae, D. spicata, B. maritima, Sesuvium portulacastrum, and Sporobolus virginicus may be more abundant at tropical latitudes (Costa and Davy, 1992), possibly due to their greater tolerance of arid conditions and dessication. Species differences in tolerance of inundation, salinity, dessication, and frost as well as ability to acquire resources under differing stresses will determine mangrove versus salt marsh shifts in response to global drivers. The physiographic equivalent of S. alterniflora is R. mangle, but the latter cannot tolerate freezing temperatures and does not extend far into subtropical latitudes. Consequently, in the southern United States S. alterniflora often intergrades with A. germinans, which is more cold tolerant (Figure 2).
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Figure 2 Ground (top) and aerial (bottom) views of a salt marsh ^ mangrove community in coastal Louisiana, USA. Black mangrove (Avicennia germinans) occurs along creekbanks and intergrades with smooth cordgrass (Spartina alterniflora) in the marsh interior.
Coexisting mangrove and salt marsh communities occur in a range of geomorphic settings. Along the coastline of New South Wales in Australia, mangrove and salt marsh are primarily located within drowned river valleys and barrier estuaries (Roy et al., 2001). While some salt marsh may be located within saline coastal lagoons, mangroves are generally excluded because mangrove propagule dispersal to these estuaries is restricted by the near permanent closure of the estuary mouth. Periodic extended flooding within coastal lagoons may also lead to the dieback of mangroves, yet allow the survival of salt marsh. Due to high wave energies along the southern Victorian coastline, mangrove and salt marsh are almost exclusively distributed within three coastal embayments; Port Phillip Bay, Western
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Figure 3 Mangroves (Avicennia marina) invading a mudflat in New Zealand (Coromandel Penninsula). Marsh vegetation in foreground is Leptocarpus similis (jointed rush).
Port Bay and Corner Inlet. In New Zealand, mangroves are located within major embayments of the northern part of the North Island (Burns and Ogden, 1985) and may occur with species such as Apodasmia similis (jointed rush) (Figure 3). Along the east coast of Florida, mangrove and salt marsh coexist in low-energy environments such as the Indian River Lagoon (Reimold, 1977; Montague and Wiegert, 1990). Along the northern coastline of the Gulf of Mexico mangrove and salt marsh may coexist within the inactive deltaic environments of the Mississippi River and chenier coastal plains in Louisiana (Patterson and Mendelssohn, 1991), in embayments along the west coast of Florida from Tampa Bay to the Cedar Keys (Kangas and Lugo, 1990; Stevens et al., 2006), and in lagoons along the southeast coast of Texas (Sherrod and McMillan, 1985; Everitt et al., 1996). In tropical Mexico, salt marshes are often associated with mangroves in coastal lagoons or near river deltas with low sediment loads (Olmsted et al., 1993). Elsewhere in Latin America, salt marsh may develop in settings that promote the development of hypersaline conditions and in disturbed mangrove areas (Costa and Davy, 1992). The species composition of salt marsh–mangrove communities in any geographic location may vary substantially due to differences in tide range, local topography, wave energy, and temperature regime. This variation is exemplified along the southern coast of the United States. On the east coast of Florida in the Indian River Lagoon complex, a typical shoreline might be dominated by R. mangle or S. alterniflora in front of a mid-marsh mixed community of Salicornia spp., D. spicata, Borrichia frutescens, B. maritima, and A. germinans (Montague and Wiegert, 1990). In south Florida, S. alterniflora might form a narrow fringe in front of a welldeveloped mangrove zone (R. mangle, A. germinans, Laguncularia racemosa) and a back mangal zone dominated by J. roemerianus (Davis, 1940). On the west coast of Florida (north of Tampa Bay to the Cedar Keys), S. alterniflora stands contain small
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stature, but abundant A. germinans. Similarly, the salt marsh–mangrove communities in coastal Louisiana and south Texas consist of S. alterniflora and A. germinans. The mangrove R. mangle occurs farther south along the Texas coast due to its greater freeze sensitivity. Geographic variation in species co-occurrence and respective inherent environmental tolerances and competitive abilities will likely determine mangrove–salt marsh interactions in any specific location.
3. LONG-T ERM D YNAMICS 3.1. Tropical northern Australia Tropical northern Australia is drained by several macrotidal rivers, each receiving seasonal floods in relation to the monsoons (Woodroffe et al., 1989). Mangroves are confined to active depositional environments, including channel banks, the edges of mid-channel islands, and prograding coastlines (Woodroffe et al., 1985). The wide upper-intertidal and supratidal flats are bare in lower rainfall areas, or covered in grasses, sedges, and, in wetter areas, Melaleuca forest (Woodroffe et al., 1989). These wide estuarine flats are known as blacksoil plains. Beneath the blacksoil plains lay extensive mangrove peat deposits. These have been located in King Sound (Semeniuk, 1980, 1982), the Fitzroy River (Jennings, 1975), the Ord River (Thom et al., 1975), and the Daly and Alligator Rivers (Woodroffe et al., 1985). Radiocarbon dating of these mangrove facies shows a consistency in age across the estuarine plains and also between river systems. For example, in the South Alligator River, 33 radiocarbon dates returned values within a range of 5370–6860 years BP with no spatial trends (Woodroffe et al., 1985). These dates corresponded with dates returned from other systems in northern Australia, including the Fitzroy (5800–7500 years BP) and the Ord (ca. 6700 BP Thom et al., 1975). From these dates Woodroffe et al. (1985) hypothesized that extensive mangrove deposits occupied much of the estuarine plains adjacent to the macrotidal rivers near the end of the postglacial marine transgression. Jennings (1975) had provided a climate interpretation of the extensive cover of mangrove, or “big swamp” phase in northern Australian estuaries, postulating higher rainfall as a possible mechanism. In tropical north Queensland, mangroves still occupy the entire estuarine plain in some rivers experiencing high, year-round rainfall, but can occur together with salt marsh species (Figure 4). By contrast, Woodroffe et al. (1985) presented a geomorphological explanation for this phase. In this model, continued sedimentation following sea-level stabilization raised intertidal elevations above those tolerated by mangroves, to the point that by 5500 BP most of the mangrove extent was confined to the fringes of channels. In dryer areas, mangroves gave way to hypersaline flats, while in wetter areas; upper intertidal and supratidal environments were colonized by salt-tolerant grasses and sedges. The more detailed investigation by Woodroffe et al. (1985) had shown that the relationship between sedimentation and sea level was the driver of mangrove–salt marsh/saltflat dynamics at the scale of the Holocene, at least in northern Australia.
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Figure 4 Salt marsh (Distichlis sp.) growing adjacent to a stand of mangroves (Ceriops tagal) in Queensland, Australia.
These studies provided a theoretical context for later work in the estuaries of southeastern Australia.
3.2. Southeastern Australia Roy (1984) presented a model of estuarine infill for SE Australia that described phases in the availability of intertidal habitat. Under this model, the deeply incised drowned river valleys characteristic of SE Australia support little in the way of intertidal vegetation immediately following postglacial marine transgression. As these valleys infill with sediment, the fluvial bayhead delta progrades seaward, supporting wide intertidal flats. It is during this “intermediate” stage of infill that estuaries support the greatest extent of mangrove and salt marsh. With the completion of infill floodplains accrete above intertidal elevations, flow is channelized throughout the length of the estuary, and intertidal habitats are restricted to channel fringes and cutoff embayments. The Roy (1984) model assumes stable sea level following the cessation of the postglacial marine transgression at approximately 6500 BP. Recent dating of encrusting organisms at several locations in New South Wales have suggested that mid-Holocene sea levels may have been somewhat higher (50–150 cm) than present levels (Baker and Haworth, 1997, 2000), an observation which might explain the widespread occurrence of potentially acid sulfate soils above current sea levels on estuarine floodplains (Wilson, 2005). Were mangroves more widely distributed in the mid-Holocene than at present? Early work on the interactions between mangrove and salt marsh in eastern Australia was dominated by succession theory. Pidgeon (1940) saw mangroves as the initial colonizers of exposed estuarine mudflat under conditions of extreme
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salinity and saturation. Mangroves were seen as active land builders, which accumulated sediment amongst aerial roots, inviting invasion by subsequent colonizers. Saltmarsh plants, the first of which are Salicornia australis, Suaeda australis and Samolus repens, invade the landward margin of the mangrove area. With subsequent silt accumulation the mangroves are forced further out into the estuary and in their place a saltmarsh is formed. Pidgeon (1940, p. 224) Under this model, continued accretion led to colonization by salt-tolerant grasses (Sporobolus virginicus), followed by the rush Juncus spp., then Casuarina glauca. As freshwater conditions replace saline, Eucalyptus invades, with the “climax” community appearing as a mixed Eucalypt forest (Pidgeon, 1940). Evidence supporting this dynamic succession was seen as the occurrence of “relict species in more advanced zones” and the active invasion of Juncus into the Sarcocornia meadow. The succession model can be tested by coring within the salt marsh in search of relict mangrove peat material. To this end, Mitchell and Adam (1989a) retrieved shallow sediment cores from several salt marshes in Botany Bay and the Georges River, in the Sydney region. Their coring failed to find any evidence of previous occupation of the salt marsh habitat by mangrove, and suggested a model which had salt marsh species as primary colonizers, followed by invasion by mangrove. This model accorded with observations of initial colonization of newly created intertidal flats by salt marsh on the northern foreshore of Botany Bay, and the spreading of mangroves into the salt marsh zone at various intertidal locations in the Sydney region (Mitchell and Adam, 1989b). The Hawkesbury River in central NSW is a large drowned river valley in an intermediate phase of infill. Unlike the macrotidal estuaries of tropical northern Australia, the Hawkesbury supports a diverse array of vegetation on depositional terraces through its tidal length of some 106 km. There is no “big swamp” phase, nor does there appear to have been at any point in the Holocene or late Pleistocene. Present-day intertidal flats support widespread mangrove in the central reaches of the estuary. Seaward of these, mudflats exposed at low tide are unvegetated. Headward, intertidal flats are dominated by salt marsh, principally Juncus kraussii. Saintilan (1997) suggested that gradual infill has led to the replacement of mangrove with salt marsh on successive intertidal flats as the river progrades seaward within its valley. Evidence for this model of geomorphically driven vegetation succession is found in the stratigraphy of intertidal flats currently occupied by salt marsh. Using techniques similar to those of Mitchell and Adam (1989b), Saintilan and Hashimoto (1999) found mangrove peats well preserved 20–30 cm beneath the present-day marsh surface, at the approximate elevation of contemporary mangrove root systems. Beneath these peats, estuarine shells dated to approximately 5000 years BP. The age of the mangrove peats varied with distance from the edge of wide intertidal flats, from 1200 to 1700 years BP at the upslope fringes of the flats to 500 years BP close to the current mangrove/salt marsh boundary, suggesting gradual infill.
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Mangrove root material retrieved from beneath salt marsh in small creeks in southern NSW told a similar story. Mangrove root material dating to 1300 BP in Currambene Creek and 1900 BP in Cararma Inlet were retrieved from beneath the contemporary salt marsh (Saintilan and Wilton, 2001). The lack of any significant fluvial input into Cararma Inlet suggested tidal reworking of aeolian and washover deposits as the most likely mechanism for infill. Within Currambene, preserved mangrove root material was more sporadic, and the suggestion was made that channel migration had erased evidence of prior mangrove occupation in some locations. At neither site was there evidence of widespread mangrove colonization earlier than 3000 BP, though more headward estuarine floodplains were not sampled.
3.3. Western Atlantic–Caribbean Region Early work in Florida and the Caribbean Region was also influenced by succession theory and the concept of mangroves as “land-builders” (Spackman et al., 1966; Cohen and Spackman, 1977). Davis (1940) proposed that mangrove–salt marsh zonation in Florida reflected ecological succession and seaward progradation by mangroves as sea level rose. However, Egler (1952); Thom (1967) and others later challenged both the zonation–succession interpretation and the idea that mangroves were geological agents capable of building land. Egler (1952) presented evidence that mangroves had retreated landward during sea-level rise, invading freshwater habitats in the Everglades. More recent work in the Caribbean has shown that mangroves growing in sediment-deficient settings can build vertically via peat formation (Woodroffe, 1981, 1983; Cameron and Palmer, 1995; Macintyre et al., 1995, 2004; McKee and Faulkner, 2000; Toscano and Macintyre, 2003; McKee et al., 2007). In some cases, mangrove peat deposits reach 10 m in thickness, and radiocarbon dating shows that this biogenic accretion has kept pace with sea-level rise (e.g., Twin Cays and Tobacco Range, Belize) (Macintyre et al., 2004; McKee et al., 2007). Paleostratigraphic evidence of fluctuations in mangrove and salt marsh dominance is limited for the Western Atlantic–Caribbean Region and generally does not record the interplay between these specific vegetation types. Early work in Florida showed basal freshwater peat beneath mangrove peat in the modern Everglades–mangrove complex, suggesting mangrove invasion of freshwater marsh during sea-level rise (Cohen and Spackman, 1977). Some of the deepest and oldest peat deposits occur in Belize where basal peats have been radiocarbon dated to about 8000 Cal BP (Toscano and Macintyre, 2003; Macintyre et al., 2004; McKee et al., 2007). Most of these studies identified mangrove-derived peat throughout the stratigraphic profile and no components were identified as “salt marsh” peat. Work by Woodroffe (1981, 1983) at Grand Cayman also showed predominately mangrove peat in radiocarbon dated cores. These latter observations show the continued dominance of mangroves throughout the Holocene despite changes in sea level. Other locations, including Jamaica (Digerfeldt and Hendry, 1987) and Trinidad (Ramcharan, 2004), show fluctuations between mangroves and brackish water
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sedges (e.g., Cladium sp.) or freshwater species, respectively. In the Negril Swamp, Jamaica, radiocarbon dated cores show sedge (Cladium sp.) peat at the base (ca. 7000–8000 Cal BP) and interspersed at intervals with mangrove peat (Rhizophora or Rhizophora–Conocarpus) or mixtures of Cladium–Rhizophora or Cladium–Conocarpus peat (Digerfeldt and Hendry, 1987). Pollen records from a late Holocene core on the Mexican Caribbean coast (Quintana Roo) showed mangroves dominating during humid periods and Conocarpus erectus and nonmangrove taxa dominating during drier periods, for example, that coincided with the decline in the Mayan culture (Islebe and Sanchez, 2002).
4. R ECENT INTERACTIONS 4.1. Air photographic evidence of mangrove–salt marsh dynamics in SE Australia On the basis of the stratigraphic evidence described above the trend in SE Australia over the later Holocene has been one of salt marsh replacing mangrove with the infilling of estuaries, and mangrove occupying freshly accreted habitat. This pattern has been consistent across a range of settings, from drowned river valleys (Saintilan and Hashimoto, 1999) to smaller barrier estuaries (Saintilan and Wilton, 2001). The pattern presents a contrast to trends identified from more recent times from air photographic records. Numerous studies from SE Australia have demonstrated the encroachment of mangrove into upper-intertidal salt marsh. Saintilan and Williams (2000) cited 28 surveys that demonstrated salt marsh loss to mangrove encroachment over the period covered by archival air photographs (usually 1940s – present). The trend is apparent across all east coast bioregions and a range of geomorphic settings. Within southern Queensland, Pleistocene sand barriers protect wide, shallow backbarrier deposits which support extensive mangrove and salt marsh. Mangrove encroachment into salt marsh in these environments is well documented (McTainsh et al., 1986; Hyland and Butler, 1988; Morton, 1994; Manson et al., 2003). In northern NSW, mangroves and salt marshes occupy the mouths of large rivers. While losses of mangrove and salt marsh to agriculture have been extensive (West, 1993), mangroves are encroaching upon salt marsh and some agricultural pastures (Saintilan, 1998). In central coast NSW, widespread losses of salt marsh to mangrove have been reported from both shallow coastal lakes (Winning, 1990) and drowned river valleys (Mitchell and Adam, 1989a,b; Evans and Williams, 1997; Williams and Watford, 1997; Williams et al., 1999; McLoughlin, 2000; Haworth, 2002; Williams and Meehan, 2004). South coast NSW estuaries, mostly smaller “barrier” estuaries (Roy et al., 2001) have shown similar trends with a median loss of approximately 40% of the salt marsh to mangrove encroachment (Meehan, 1997; Chafer, 1998; Saintilan and Wilton, 2001). Within Victoria, salt marshes and mangroves occupy the shorelines of large coastal embayments. Here, loss of salt marsh to mangrove encroachment has been
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consistent though less dramatic. Declines of 5–12% of salt marsh to mangrove encroachment have been reported for the salt marshes of Western Port Bay (Rogers et al., 2005b) and losses have also been noted for Corner Inlet (Vanderzee, 1988) and the Gulf St. Vincent in South Australia (Burton, 1982). Within New Zealand, the proliferation of mangrove is more commonly described as a seaward colonisation (Cragg et al., 2001; Park, 2001; Morrisey et al., 2003) though landward encroachment has been noted (Burns and Ogden, 1985).
4.2. Saltwater intrusion in Northern Australia Marked geomorphological changes to estuaries and coastal plains in Northern Australia over the past 50 years have been associated with saltwater intrusion in the Alligator River Region (Winn et al., 2006) and Mary River (Knighton et al., 1991; Mulrennan and Woodroffe, 1998). The gradual extension of tidal influence along stream channels, the expansion of tidal creeks and the formation of new tidal creeks (Winn et al., 2006) is linked to the encroachment of mangrove and saline mudflats into freshwater vegetation (Finlayson et al., 1998), localized scour and dieback within Melaleuca forests, accretion of sediment on floodplains (Knighton et al., 1991; Woodroffe and Mulrennan, 1993; Bell et al., 2001), changes in subsurface hydrology (Jolly and Chin, 1992) and land cover changes (Ahmad and Hill, 1995; Bell et al., 2001). Changes since 1950 are significant with bare saline mudflats on the East Alligator River exhibiting a ninefold increase and an associated loss of 64% of Melaleuca forests by 2000 (Winn et al., 2006). More than 17,000 ha of freshwater vegetation have been adversely affected on the Mary River and a further 35–40% of floodplains are immediately vulnerable to intrusion (Mulrennan and Woodroffe, 1998). A single cause for saltwater intrusion has not been identified (Mulrennan and Woodroffe, 1998). Instead, it is apparent that factors such as drier-than-average monsoonal conditions, low-frequency and low-intensity cyclonic events, and above average ocean water levels (Winn et al., 2006) facilitate extension of tidal influence into freshwater environments, while tributary development, large tidal range, small elevation differences over floodplains, and uncontrolled feral buffalo promote the expansion of tidal influence (Knighton et al., 1991). Due to the desiccation of floodplain sediments in the dry seasons, the process of saltwater intrusion now appears to be internally driven and is likely to continue until an equilibrium state is reached between floodplain elevation and tidal influence (Mulrennan and Woodroffe, 1998; Winn et al., 2006).
4.3. Western Atlantic–Gulf of Mexico Information on historical shifts in salt marsh and mangrove vegetation in the Western Atlantic–Gulf of Mexico region is limited to site-specific observations (Lonard and Judd, 1985; Sherrod and McMillan, 1985; McMillan and Sherrod,
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1986; Montague and Wiegert, 1990; Montague and Odum, 1997; McKee et al., 2004; Stevens et al., 2006). In the northern Gulf of Mexico, the persistence and spread of mangroves is controlled mainly by temperature (Sherrod and McMillan, 1985; Stevens et al., 2006) although hydroperiod, sedimentation, salinity, and nutrients as well as propagule dispersal and predation determine local patterns of establishment and spread (Patterson and Mendelssohn, 1991; Patterson et al., 1993, 1997). In coastal Louisiana, for example, mangroves tend to dominate shorelines of tidal creeks, and salt marsh occupies the interior position. The A. germinans zone is characterized by higher elevation, salinity and soil bulk density, whereas the S. alterniflora zone is lower in elevation with greater flooding and reducing soils (Patterson and Mendelssohn, 1991). Exclusion of A. germinans from the interior marsh zone has been attributed to limited retention of and higher predation on propagules (Patterson et al., 1997) and plant competition in combination with greater flooding stress (Patterson et al., 1993). Periodic freezes have killed or damaged mangroves allowing expansion of salt marsh, most recently during the 1980s (McMillan and Sherrod, 1986; Montague and Wiegert, 1990; Montague and Odum, 1997; Stevens et al., 2006). Stevens et al. (2006) review reports of major damage to mangroves in Florida as a result of severe freezes in 1962, 1977, 1981, 1983, 1985, 1989, and 1996. A. germinans has some capacity to recover from freeze damage through coppicing (stump sprouting), but periodic damage results in a stunted, scrublike growth form. In Louisiana, the last severe freeze that caused widespread dieback of mangroves (A. germinans) occurred in December, 1989 (K.L. McKee, pers. obs.). Although A. germinans can recover from mild freezes, the 1989 event killed mature trees. However, propagules that had already dispersed survived this freeze and appeared to be the primary means by which the population reestablished. Thousands of propagules were observed establishing in the months after the freeze, and their viability was further confirmed in greenhouse culture (K.L. McKee, unpublished data.). The resistance of A. germinans propagules in Louisiana to freezing temperatures is consistent with previous studies of cold tolerance of this species (Markley et al., 1982; Norman et al., 1984; Sherrod et al., 1986). Most observations indicate that salt marsh replaces freeze-killed mangroves in the Gulf of Mexico within 4–5 years (Stevens et al., 2006). However, relatively mild winters and the lack of severe freezes in recent years has allowed mangrove expansion in Florida (Stevens et al., 2006) and Louisiana (K.L. McKee, pers. obs.). Mangroves have expanded (average of 21% increase in cover) at the Cedar Keys, Florida (29080 N) between 1995 and 1999 during a period free from severe freezes (Stevens et al., 2006). Recent mangrove expansion in coastal Louisiana, USA, has been attributed to a drought-induced, widespread dieback of S. alterniflora during 2000 (McKee et al., 2004). Over 40,000 ha of salt marsh were severely damaged with areas up to 5 km2 reduced to mudflat. However, A. germinans was unaffected and even thrived during and after this marsh dieback event (Figure 5).
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Figure 5 Salt marsh (Spartina alterniflora) dieback in coastal Louisiana, USA (top). Black mangrove (Avicennia germinans) was unaffected and later became dominant in some dieback areas (bottom).
5. STRESSORS C ONTROLLING D ELIMITATION OF M ANGROVE Mangroves commonly exhibit a gradual reduction in development with latitude (Lot et al., 1975; Saenger and Snedaker, 1993; Saenger, 2002) and stunted forms at their latitudinal limit may be due to an increasingly stressful environment (Schaeffer-Novelli et al., 1990). Kangas and Lugo (1990) hypothesize that without “stress,” in their case frost stress, mangrove vegetation is competitively superior and coastlines would be dominated by mangrove. However, field observations of salt marsh dieback in coastal Louisiana, USA, and A. germinans expansion into areas formerly occupied by S. alterniflora (McKee et al., 2004) suggest that at subtropical latitudes some mangrove species are competitively inferior. In addition, greenhouse
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and field experiments show that S. alterniflora competitively inhibits growth of A. germinans (Patterson et al., 1993; McKee and Rooth, 2008). Where mangrove and salt marsh coexist, environmental stressors control the delimitation of mangrove so that these communities typically form three zones: a seaward zone dominated by mangrove, a landward transition zone containing both mangrove (possibly stunted) and salt marsh, and an interior zone dominated by salt marsh. Numerous stressors, including geomorphic, hydrological and climatic controls, physicochemical factors, and biotic interactions may act to delimit mangrove establishment (Patterson et al., 1993, 1997; McKee and Rooth, 2008) and enable salt marsh development within the interior of marshes. In some extreme environments, however, marsh vegetation may facilitate mangrove recruitment through propagule trapping and/or amelioration of stressful soil conditions (McKee and Rooth, 2008).
5.1. Geomorphic and hydrological controls Inundation by tides is not an ecophysiological requirement of mangroves and salt marshes, but it is a typical feature and plays an important role by limiting excessive buildup of salts within soils (Saenger, 2002) and distributing mangrove and salt marsh propagules (Adam, 1990; Clarke, 1993). Indeed, inundation may play a significant role in mangrove and salt marsh zonation as tidal flushing may act to alter soil salinity (Patterson and Mendelssohn, 1991) and control the transport of propagules along an elevation gradient (Adam, 1990; Patterson et al., 1997; Saenger, 2002). The hydrologic role of inundation is largely controlled by the geomorphology of a marsh so that inundation frequency and water volume is negatively proportional to land elevation. However, the relative positions of salt marsh and mangrove vegetation along a topographic gradient varies with geographic region and the species involved. In the northern Gulf of Mexico, mangroves coincide and compete with low marsh species such as S. alterniflora, which are highly flood tolerant (Patterson and Mendelssohn, 1991; Patterson et al., 1993, 1997). In these communities, mangroves (A. germinans) occupy the higher elevation creek banks and salt marsh species dominate the interior, lower elevation areas. In other locations, low elevation sites are commonly inhabited by mangrove, are inundated more frequently, and receive higher water volumes, while higher elevations are commonly inhabited by salt marsh. The latter pattern can be found in coexisting mangrove and salt marsh communities in Australia as well as in Florida, USA (Montague and Wiegert, 1990). In the latter case, the low elevation sites along shorelines are dominated by R. mangle and higher elevations are occupied by high marsh species such as B. maritima, D. spicata, Borrichia frutescens, and Salicornia spp.; the upland transition may be occupied by J. roemerianus, Spartina bakerii, or S. patens (Montague and Wiegert, 1990). Adaptations to cope with submergence and high salt environments are evident within both mangrove and salt marsh species (summarized in Adam, 1990; Saenger, 2002) and species tolerance to submergence and salinity are hypothesized to cause the zonation of plants observed within marshes. Numerous studies have demonstrated the role of salinity in causing mangrove species to segregate (Ball, 1988a,b, 2002; Lin and Sternberg, 1992; Ball and Pidsley, 1995; Lo´pez-Hoffman et al., 2006)
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and coexist with temporal salinity variation (Ball, 1998). However for salt marsh, Adam (1990) argues that since soil salinity varies spatially and temporally, zonation may be more readily explained on the basis of species tolerance to frequency and duration of submergence. Furthermore, while soil salinity and tidal regime are important factors, these factors do not appear to explain observed distributions of salt marsh species across the intertidal zone (Silvestri et al., 2005). The zonation of coexisting mangrove and salt marsh communities in Sydney (Clarke and Hannon, 1971) and Brazil (Santos et al., 1997, cited in Saenger, 2002) were found to be salinity based, with species exhibiting definable salt tolerances (Clarke and Hannon, 1971) and their distribution being determined by a salinity model based on topographic level and upland runoff (Santos et al., 1997). While causes of mangrove and salt marsh zonation remain incompletely understood (Snedaker, 1982; Adam, 1990), it is evident that zonation, soil salinity and submergence are all maintained along elevation gradients and is likely to contribute to the distribution of mangrove and salt marsh vegetation (Snedaker, 1982; Saenger, 2002). Zonation as a consequence of differential dispersal of mangrove propagules has also been proposed. This is primarily based on the concept of tidal sorting of propagules proposed by Rabinowitz (1975, 1978a,b,c) for mangroves in Panama. However, the relationship between propagule size and zonation has not been supported for mangroves in other locations (Ball, 1980; Saenger, 1982; McKee, 1995a) In addition, differential dispersal and survival of propagules with tides has not been established for salt marsh species and since many salt marsh species are able to vegetatively propagate the contribution of tides to dispersal is likely to be limited (Adam, 1990). Regardless of the role of tidal sorting of propagules in zonation, tides do distribute mangrove propagules and it is likely that zonation of mangrove to salt marsh may be related to differential seedling mortality at different tidal levels (Ellison and Farnsworth, 1993). High soil salinity decreases the rate of pericarp shedding by mangrove propagules, thereby limiting survival (Downton, 1982; Clarke and Myerscough, 1991). Thus, higher soil salinities and reduced inundation frequency, common in the high marsh, causes mangrove propagules to desiccate and suffer high mortality (Clarke and Allaway, 1993). However, work in the northern Gulf of Mexico found that mortality of A. germinans propagules was associated with excessive flooding and damage by predators in low marsh areas, resulting in decay and loss of viability (Patterson et al., 1997). Observed interactions between mangrove and salt marsh globally have highlighted the role of hydrology and geomorphology in increasing the dispersal and survival of mangrove propagules and seedlings. For example, a strong relationship has been found between relative sea-level rise and the upslope migration of mangrove into salt marsh environments in Southeastern Australia (Rogers, 2004). In this case, relative sea-level rise incorporates both geomorphological and hydrological changes and may be altered by eustatic sea-level rise, which relates to changes in the volume of water within oceans through the melting of ice caps and thermal expansion of water; and/or surface elevation changes of the land through surface processes of vertical accretion or subsurface processes of subsidence and autocompaction.
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Hydrological changes due to global warming has caused global sea level to increase by 1.0–2.0 mm/year in the 20th century the rate of sea-level rise is predicted to accelerate toward 2,100 (Solomon et al., 2007). Recent studies of relative sea-level trends for salt marshes in Tasmania and New Zealand show higher rates over the past century (2.5–3.1 mm/year) compared to the previous four centuries (Gehrels et al., 2007). Numerous studies predict that global sea-level rise will cause a shift in the boundary between mangrove and salt marsh (Snedaker, 1995; Blasco et al., 1996) and indeed the study of Rogers et al. (2006) related sealevel changes to the expansion of mangroves in Southeastern Australia. However, changes to the water levels inundating a marsh may also occur due to altered tidal prisms resulting from engineering works, such as dredging of estuary entrances. These hydrological changes may increase the mean water level and/or alter the tidal range within a marsh, thereby altering soil salinity and increasing the distribution potential of mangrove propagules. Both reduced soil salinities and increased inundation promote the survival of mangrove within salt marsh and increases their competitive potential. Geomorphological changes to marsh elevation alter the inundation frequency and duration within a marsh and may occur as a result of surface processes such as vertical accretion. Increased sediment availability and vertical accretion within estuaries has been hypothesized as a mechanism promoting mangrove encroachment of salt marsh in Southeastern Australia (summarized by Saintilan and Williams, 1999; McLoughlin, 2000) by creating more areas suitable for mangrove establishment and providing a soft substrate for mangrove propagule establishment within the salt marsh. Certainly the spread of mangrove along estuarine foreshores seems related to the availability of new habitat, and commonly the landward encroachment of mangrove is accompanied by seaward encroachment also. Perhaps the overall increase in population of A. marina, and increased fecundity made possible by higher nutrient loads has increased propagule availability within estuaries, including the upper-intertidal environment. However, increased vertical accretion is also more likely to increase marsh elevation, reducing inundation frequency, and thereby reducing the likelihood of mangrove survival due to propagule desiccation. Further, the “fertilization” of the salt marsh from anthropogenic sources does not necessarily lead to improved mangrove colonization. Experimental fertilization of A. marina propagules has failed to show an impact on propagule survival (Clarke and Myerscough, 1993; Saintilan, 2003). Rates of vertical accretion have been used as an estimate for marsh elevation and a surrogate for investigating the cause of mangrove–salt marsh interaction. This is appropriate at some sites, such as Homebush Bay and Western Port Bay, Australia, where salt marsh surface elevation is maintained primarily by vertical accretion (Rogers et al., 2005a,b). However, this assumption is not always true. Recent literature has highlighted the role of belowground processes in altering marsh elevations (Cahoon et al., 1999, 2006; McKee et al., 2007). Processes such as subsidence, soil compaction, plant productivity and dieback, groundwater availability, and tidal flooding interact to alter soil volumes over a range of spatial and temporal scales.
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The relationship between marsh elevations and sea level has primarily been explored through the use of surface elevation tables (described by Cahoon et al., 2002). By incorporating analyses of surface processes such as vertical accretion, the belowground component of surface elevation change can be differentiated. In southeastern Australia, Surface Elevation Tables (SET) were used to show that mangrove surface elevation gain was consistently lower than salt marsh surface elevation gain despite vertical accretion being consistently higher in the mangrove zone (Rogers et al., 2005a,b, 2006). Mangrove encroachment of salt marsh in the same region has been attributed to autocompaction or subsidence, which causes the observed disequilibrium between mangrove surface elevation change and vertical accretion (Rogers et al., 2006). Those marshes with the highest rates of relative sealevel rise were those with the highest rates of mangrove encroachment (Rogers et al., 2005b). Autocompaction of sediments corresponded to a severe El Ninorelated drought, and surface elevation change correlated strongly with total monthly rainfall (Rogers et al., 2005a), the Southern Oscillation Index and groundwater depth (Rogers and Saintilan, 2008). This research is consistent with other studies of surface elevation and groundwater (Schmidt and Burgman, 2003; Watson, 2004; Whelan et al., 2005) and suggests that rainfall and subsequent groundwater recharge plays a significant role in determining interannual variability in surface elevation with respect to sea level.
5.2. Climatic controls Climatic factors are commonly proposed to promote or limit the survival of mangrove within coexisting mangrove and salt marsh communities as a result of extreme low temperatures and frost causing the death of mangrove, or rainfall influencing soil salinity. Many tropical and subtropical plants exhibit physiological dysfunction and damage in response to chilling temperatures (van Steenis, 1968) and this appears to be the case for mangroves (McMillan, 1975; Markley et al., 1982; McMillan and Sherrod, 1986; Kangas and Lugo, 1990; Olmsted et al., 1993). Freeze sensitivity is particularly apparent along the Gulf and Atlantic coasts of North America as discussed in the previous section. In the northern Gulf of Mexico, mangroves tend to expand during periods of mild winters (Stevens et al., 2006) and when weather extremes or other factors stress or eliminate salt marsh (McKee et al., 2004). When mangroves are periodically damaged or killed by freezes, salt marsh quickly regains dominance (Stevens et al., 2006). Consequently, future changes in global climate regime and local weather patterns will likely influence the relative dominance of mangrove and salt marsh vegetation in this region. Mangroves show greater luxuriance and species richness in high rainfall areas (Tomlinson, 1986; Hutchings and Saenger, 1987; Smith and Duke, 1987) and species distribution correlates strongly with freshwater influence from rainfall (Bunt et al., 1982), controlling for latitude. In fact, low rainfall and a shortened rainfall season in Senegal in the 1980s contributed to the decline of mangroves and increased surface areas of bare salt flats and saline grasslands (Diop et al., 1997). In addition, studies in Australia have suggested that freshening of salt marsh
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environments from rainfall might promote survivorship of mangroves (McTainsh et al., 1986) and may explain the establishment of mangrove at the salt marsh– terrestrial boundary in tropical and subtropical areas of Australia (Duke, 2006). There is other evidence that mangrove encroachment in tropical and subtropical climates may be related to rainfall trends. In these situations, the proportion of mangrove and salt marsh in estuaries is correlated to annual rainfall (Bucher and Saenger, 1991). Where salt marsh occurs, it often occupies a middle position between seaward and landward mangrove zones. Increased rainfall in recent decades has been used to explain the encroachment of tall mangrove forest into saltpan in Hinchinbrook Channel (Duke, 1995) and stunted mangrove into salt marsh in Moreton Bay (Duke, pers. comm.). Conversely, declines in rainfall have resulted in mangrove dieback in the upper intertidal environment in some locations (Duke, pers. comm.) The correlation between rainfall and the proportion of mangrove and salt marsh breaks down in New South Wales (Saintilan, 2003). Here, the upslope limit of mangrove is more consistently related to frequency of tidal inundation, and there is no landward mangrove fringe.
5.3. Physicochemical factors Physicochemical factors have been hypothesized to limit the establishment of mangrove within salt marsh areas. In particular, mangroves may be excluded from salt marsh environments due to nutrient deficiency and/or high soil salinity within salt marshes. Salinity decreases pericarp shed and increases desiccation of mangrove propagules (Downton, 1982; Clarke and Myerscough, 1991) and may contribute to zonation of mangrove and salt marsh species due to differential species tolerance to soil salinity (summarized in Adam, 1990; Saenger, 2002) as discussed above. Experimental fertilization of mangroves indicate that nitrogen and phosphorus limitation is common (Boto, 1983; Boto and Wellington, 1983; McKee et al., 2002) and upon fertilization mangroves exhibit increases in growth and biomass (Naidoo, 1987; Clarke and Allaway, 1993; Feller et al., 2003a,b; Lovelock et al., 2004, 2006). However, since the effect of nutrient enrichment does not become apparent until mangrove cotyledons have been exhausted (Naidoo, 1987; Clarke and Allaway, 1993), it has less impact on early mangrove establishment within salt marsh. Saintilan (2003) confirmed that mangrove mortality within salt marsh is not reduced through nutrient enrichment of seedlings. In some geographic regions, mangrove mortality within salt marshes is the result of desiccation of propagules (Clarke and Allaway, 1993; Saintilan, 2003), which is dependent on climatic, geomorphic and hydrologic factors. However, changes in tissue chemistry by fertilization or light conditions (McKee, 1995b) could influence mangrove seedling susceptibility to herbivores.
5.4. Biotic interactions Several studies suggest that salt marsh precedes mangrove in successional sequences due to tidal transport of mangrove propagules and shading of salt marsh by mangrove seedlings (summarized in Kangas and Lugo, 1990). Recent work by
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McKee and Rooth (2007) found that establishment, survival, and growth of mangrove seedlings were facilitated by herbaceous species, Sesuvium portulacastrum and D. spicata. Both species promoted mangrove propagule trapping and ameliorated soil conditions (aeration, temperature). Alternatively, dense stands of saltbush (Tecticornia spp.) have been suggested to inhibit mangrove establishment in salt marshes in Western Port Bay, Australia by either shading seedlings or collecting sediments and building up the marsh surface to exclude mangrove (Rogers et al., 2005b). Similar competition has been reported between mangroves and S. alterniflora in marshes in Louisiana (Patterson et al., 1993; McKee and Rooth, 2008). Plant–animal interactions are also evident within coexisting communities and differential predation of mangrove propagules by crabs is suggested as a cause of zonation within mangroves (Smith, 1987; Smith et al., 1989), but this hypothesis is not supported in other locations (McKee, 1995a; Sousa and Mitchell, 1999). A study of A. germinans seedlings found that predator damage to propagules (and frequent flooding) leads to decay of propagules within the Spartina zone in coastal Louisiana (Patterson et al., 1997).
6. C ONCLUSIONS The interaction between mangrove and salt marsh is of interest to climate change research for several reasons. Having evolved in the tropics (Chapman, 1977; Specht, 1981) mangroves are most prolific in lower latitudes and decrease in diversity and vigor toward the poles (Lot et al., 1975; Saenger and Snedaker, 1993). In contrast, salt marshes increase in floristic diversity with increasing latitude within the temperate zone (Adam, 1990), and in the absence of mangroves form the characteristic intertidal vegetation in these latitudes. In the predominantly temperate latitudes where the two community types coexist, salt marsh distribution is confined in part by the presence of mangrove and in part by the extreme environmental conditions of the upper intertidal zone. Environmental variability can therefore profoundly influence the competitive interactions between mangrove and salt marsh. In southern United States, where the proliferation of mangrove is periodically checked by frost, higher temperature and decreased freeze frequency and severity could enhance the survival and growth of species such as A. germinans. Wider and more continuous stands of this species may be expected in the northern Gulf of Mexico and the Atlantic coast of Florida. Less cold-tolerant species, such as R. mangle, may be able to shift northward along the coasts of Baja, California, Texas, and Florida. In SE Australia and New Zealand, the growth of the mangrove A. marina will be aided not only by increased temperatures toward its southern limit of distribution, but also by higher sea levels. Presently, the upslope limit of A. marina is defined by the mean high water mark. While there is continued debate over the causes of mangrove encroachment into the salt marsh of eastern Australia, there is growing evidence for a role of relative sea-level rise, evidenced by the relationship between the rate of encroachment and relative sea-level trends, and an increase rate of
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relative sea-level rise in temperate Australasian salt marshes over the past century. The subtle elevation gradients defining the position of mangrove and salt marsh in these situations provide an early indication of the effects of sea-level rise in the coastal wetlands of the region. The concern in these situations is that the landward transition of salt marsh may be impeded by topographic constraints, both cultural and natural. Where feasible, thought should be given to the designation of landward “accommodation space” in anticipation of projected rates of sea-level rise, so that the full range of wetland vegetation communities can continue to coexist in these estuaries.
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Lovelock, C.E., Feller, I.C., McKee, K.L., Engelbrecht, B.M.J., Ball, M.C., 2004. Experimental evidence for nutrient limitation of growth, photosynthesis, and hydraulic conductance of dwarf mangroves in Panama. Funct. Ecol. 18, 25–33. Macintyre, I.G., Little, M.M., Littler, D.S., 1995. Holocene history of Tobacco Range, Belize, Central America. Atoll Res. Bull. 430, 1–18. Macintyre, I.G., Toscano, M.A., Lighty, R.G., Bond, G.B., 2004. Holocene history of the mangrove islands of Twin Cays, Belize, Central America. Atoll Res. Bull. 510, 1–16. MacNae, W., 1966. Mangroves in eastern and southern Australia. Aust. J. Bot. 14, 67–104. Manson, F.J., Loneragan, N.R., Phinn, S.R., 2003. Spatial and temporal variation in distribution of mangroves in Moreton Bay, subtropical Australia: a comparison of pattern metrics and change detection analyses based on aerial photographs. Estuar. Coast. Shelf Sci. 57, 653–666. Markley, J.L., McMillan, C., Thompson Jr., G.A., 1982. Latitudinal differentiation in response to chilling temperatures among populations of three mangroves, Avicennia germinans, Laguncularia racemosa, and Rhizophora mangle, from the western tropical Atlantic and Pacific Panama. Can. J. Bot. 60, 2704–2715. McKee, K.L., 1995a. Mangrove species distribution patterns and propagule predation in Belize: an exception to the dominance-predation hypothesis. Biotropica 27, 334–345. McKee, K.L., 1995b. Interspecific variation in growth, biomass partitioning, and defensive characteristics of neotropical mangrove seedlings: Response to light and nutrient availability. Am. J. Bot. 82, 299–307. McKee, K.L., Cahoon, D.R., Feller, I.C., 2007. Caribbean mangroves adjust to rising sea level through biotic controls on soil elevation change. Global Ecol. Biogeogr. 16, 545–556. McKee, K.L., Faulkner, P.L., 2000. Mangrove peat analysis and reconstruction of vegetation history at the Pelican Cays, Belize. Atoll Res. Bull. 468, 46–58. McKee, K.L., Feller, I.C., Popp, M., Wanek, W., 2002. Mangrove isotopic (d 15N and d13C) fractionation across a nitrogen vs. phosphorous limitation gradient. Ecology 83, 1065–1075. McKee, K.L., Rooth, J.E., 2007. Mangrove recruitment after forest disturbance is facilitated by herbaceaous species in the Caribbean. Ecol. Appl. 17 (6), 1678–1693. McKee, K.L., Rooth, J.E., 2008. Where temperate meets tropical: multi-factorial effects of elevated CO2, nitrogen enrichment, and competition on a mangrove-salt marsh community. Global Chang. Biol. 14, 971–984. McKee, K.L., Mendelssohn, I.A., Materne, M.D., 2004. Acute salt marsh dieback in the Mississippi River deltaic plain: a drought-induced phenomenon? Global Ecol. Biogeogr. 13, 65–73. McLoughlin, L., 2000. Estuarine wetlands distribution along the Parramatta River, Sydney, 1788–1940: implications for planning and conservation. Cunninghamia 6, 579–610. McMillan, C., 1975. Adaptive differentiation to chilling in mangrove populations. In: Walsh, G.E., Snedaker, S.C., Teas, H.J. (Eds.), Proceedings of the International Symposium on Biology and Management of Mangroves, vol. 1. University of Florida, Gainesville, pp. 62–68. McMillan, C., Sherrod, C.L., 1986. The chilling tolerance of black mangrove, Avicennia germinans, from the Gulf of Mexico Coast of Texas, Louisiana and Florida. Contrib. Mar. Sci. 29, 9–16. McTainsh, G., Iles, B., Saffinga, P., 1986. Spatial and temporal patterns of mangroves at Oyster Point Bay, South East Queensland, 1944–1983. Proc. R. Soc. Qld. 99, 83–91. Meehan, A., 1997. Historical Changes in Seagrass, Mangrove and Saltmarsh Communities in Merimbula Lake and Pambula Lake. Honours Thesis, University of Wollongong. Mendelssohn, I.A., McKee, K.L., 2000. Salt marshes and mangroves. In: Barbour, M.G., Billings, W.D. (Eds.), North American Terrestrial Vegetation. Cambridge University Press, Cambridge, pp. 501–536. Mildenhall, D.C., Brown, L.J., 1987. An early Holocene occurrence of the mangrove Avicennia marina in Poverty Bay, North Island, New Zealand: its climatic and geological implications. N. Z. J. Bot. 25, 281–294. Mitchell, M.L., Adam, P., 1989a. The relationship between mangrove and saltmarsh communities in the Sydney region. Wetlands (Australia) 8, 37–46. Mitchell, M.L., Adam, P., 1989b. The decline of saltmarsh in Botany Bay. Wetlands (Australia) 8, 55–60. Mitsch, W.J., Gosselink, J.G., 2000. Wetlands, third ed. John Wiley and Sons, New York.
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C H A P T E R
3 2
W ETLAND L ANDSCAPE S PATIAL M ODELS Enrique Reyes
Contents 1. 2. 3. 4.
Introduction Physical Models Hydraulic Modeling Hydrodynamic Modeling 4.1. Finite difference solutions 4.2. Finite element solutions 4.3. Finite volume solutions 5. Ecological Models 6. Individual-Based Modeling 7. Eco-Geomorphological Modeling 8. Ecosystem-Level Modeling 9. Desktop Dynamic Modeling 10. Conclusions Acknowledgments References
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1. INTRODUCTION Wetlands are heterogeneous ecosystems with an ample diversity of vegetation and fauna occupying different habitats at different times and space. These arrays of organisms and their interactions take place at specific locations and respond differently to global or regional environmental drivers. Cultural eutrophication, headwater diversions, and land use conversion are examples of local drivers that require sound long-term knowledge, including forecasting capabilities, to allow the application of scientifically based management. Dynamic wetland modeling as a research tool should include environmental and biotic components that interact with each other. For example, as water flows over a wetland, vegetation response may vary with sediment deposition, duration of flooding, and nutrient inputs. Conversely, as the plant biomass increases, water flow and tidal channel morphology could be altered accordingly. Thus, the challenge for wetland landscape research is to examine this inherent complexity at pertinent temporal and spatial scales. Coastal Wetlands: An Integrated Ecosystem Approach
2009 Elsevier B.V. All rights reserved.
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The recognition that wetland ecosystems are highly complex requires that multidisciplinary, holistic studies be conducted for better understanding of their functioning. To incorporate this information into a modeling framework, holistic perspectives begin with the construction of conceptual and mathematical models to synthesize current knowledge and direct further research (Heemskerk et al., 2003; Twilley, 2003). Wetland dynamic modeling uses mathematical representations of a system and its interacting components as a systematic approach to environmental problems, where current knowledge and understanding about that particular system are tested on a particular spatial scale (Costanza and Voinov, 2004). Considering the state-of-the-art at the time, Mitsch et al. (1988) provided a review of the diverse wetland modeling efforts and their holistic approaches designed to represent dynamic behavior. However, most of the studies presented in the review treatment wetlands as homogeneous systems; the one exception was the work by Costanza et al. (1988), who described how spatial dynamic modeling can analyze spatially explicit responses. Presently, this particular approach has expanded considerably. The ability of dynamic models to project cumulative impacts at different temporal and spatial scales makes them extremely useful for the understanding, experimentation, and forecasting of complex systems. The inherent predictive nature of dynamic models allows the ecosystem researcher to evaluate diverse scenarios and their consequences. Particularly, the need for a wetland dynamic model is in response to the need to simplify complexity of landscape or watershed processes that drive the fluxes and that transport water or any material into the wetland area. An example of such processes is the rainfall/ runoff-related transport of nutrients from the earth’s uplands, flow through fluvial systems, and finally transport through wetlands and into the coastal zone (Moustafa and Hamrick, 2000; Li et al., 2003). This chapter describes several spatial approaches to wetland ecosystem modeling. It begins with modeling approaches following a general ecosystem conceptual model for a wetland (Figure 1; Sklar et al., 1990; Ogden et al., 2005). Water, suspended particles (conservative or not), and biotic components are the three most Environmental parameters Marsh vegetation
Channel water
Consumers
Marsh water Suspended sediment Belowground biomass
Groundwater
Deposited/buried sediment
Figure 1 Wetland ecosystem conceptual diagram. Black arrowheads indicate material flux and white arrowheads indicate processes.
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common variables modeled which, along with their interactions, can provide insight into ecosystem response to natural and anthropogenic influences. The approaches and models here presented focus either on one of the three components or all of them depending on the particular research objectives. The first section describes water-based models, as water flow is the principal environmental parameter responsible for wetland heterogeneity. The following sections highlight approaches to represent spatial distributions using mass-balance approaches, individual-based models for mixed ecosystems, or fully integrated ecosystem simulations. Wetland dynamic modeling thus is defined and treated here as process-based computations of physical and biological processes. Empirical dynamic models based on statistical methods are another approach used primarily to discern relationships and suggest mechanisms. However, the emphasis here is placed on process-based deterministic models, as they allow for feedbacks and interactions among different components, making them more suitable for detailed forecasting.
2. PHYSICAL M ODELS Water presence and flow are the determinant environmental drivers for a wetland. Flood duration and flow allow materials to be distributed across an area and to interact with the diverse organisms that constitute a wetland community. Physical models of water can take two approaches: hydraulic and hydrodynamic. Hydraulic models are commonly used as a rapid assessment tool for wetland water balances (Arnold et al., 1998) and focus mostly on water volume and its budget in a wetland. Hydrodynamic models aim to describe water levels and velocities with a high degree of precision and thus require detailed mathematical descriptions of physical processes such as momentum.
3. HYDRAULIC M ODELING Hydraulic wetland models account for water at a particular location using a combination of linear relationships and decision rule trees (Figure 2). In heterogeneous environments, different driving functions will affect this accounting, so it must be combined with specific decision rules for several types of vegetation to represent different communities and their distribution. Linear relationships accounting for water budgets are commonly based on equations such as Wt = W þ
t X ðRi Qi ETi Pi Þ
ð1Þ
t=1
where water (W) at time (t) is a function of previous water volume and processes like precipitation (R), surface discharge (Q), evapotranspiration (ET), and percolation (P). Each of these processes then can be described either as linear functions or Boolean logic (e.g., if air temperature is higher than 0C then runoff can occur).
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Weather data
Wetland hydrology
Surface water No
Air temp >0?
Air temp >0?
Yes
No
Yes
Water volume
Vegetation Groundwater
No
Water depth diff.
Yes
Types
Figure 2 Decision rule trees forWETSIM hydraulic model. Each component calculates water budgets according to environmental clues (e.g., temperature). Modified from Voldseth et al. (2007).
Using this combined approach, Park et al. (1986) developed the sea level affecting marshes model (SLAMM). This GIS-based model computes four environmental processes that affect wetland vegetation under different scenarios of sea level rise. At each location, and with a time step of 5–25 years, factors such as inundation, erosion, overwash, and saturation are combined as relative sea level rises, which in turn determines marsh vegetation type. An important feature of this model is the ability to allow marsh migration if conditions on adjacent land are favorable under the simulated sea level rise to examine its cumulative effects (Clough and Park, 2006). Decision rules (i.e., Boolean logic) and linear relationships allow large areas (hundreds of square kilometers) to be modeled at high resolution (usually 30 m2; Glick et al., 2007). It should be noted that SLAMM is run with large time steps. This makes validation difficult because long-term historical data, such as aerial photographs, are seldom available. The most common model used for freshwater wetlands is the wetland simulation (WETSIM) model developed in the 1990s (Poiani et al., 1996; Johnson et al.,
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2005) for prairie wetlands (Figure 2). This model computes several spatially explicit vegetation algorithms on semipermanent wetlands with a variable time step (daily or monthly). Initially, only wetland vegetation types with a small footprint (9.5 m2) were accounted for and the simulation was restricted to approximately 47,000 m2. Subsequent versions (WETSIM 3.2; Voldseth et al., 2007) include agricultural cover types and improved algorithms to better represent evapotranspiration.
4. H YDRODYNAMIC M ODELING Because of the ample variety of wetland types, hydrodynamics and their modeling can be quite variable. Governing equations must include continuity, momentum, and constituent water transport if large scales are of interest. Conversely, when the processes of interest are on small scales (e.g., length of the basin is smaller than the tidal wavelength), such as in the case of some geomorphological models, then simplifying assumptions can be made (Rinaldo et al., 1999a,b). Dynamic modeling of water on a large scale (thousands square meters) in wetlands can be approached by solving the same set of equations as used for hydrodynamic models of rivers or oceans to simulate water flow and quality. Governing equations for the rate of change on a fluid in motion can be described as u
@q @q @q þv þw @x @y @z
ð2Þ
where q is a quantity of interest; x, y, and z are Cartesian coordinates; and u, v, and w are the components of the velocity on each axis. For a particular direction (as when building one- or two-dimensional models), this total derivative can be computed on a particular direction such that
x-direction:
du du du du du = þu þv þw dt dt dx dy dz
ð3Þ
y-direction:
dv dv dv dv dv = þu þv þw dt dt dx dy dz
ð4Þ
z-direction:
dw dw dw dw dw = þu þv þw dt dt dx dy dz
ð5Þ
The selection of using either or all of these derivatives gives the option to simplify these equations into one-, two-, or three-dimensional models. One-dimensional models are commonly used for streams (McCutcheon, 1982) and recently for
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wetland tidal network (Mudd et al., 2004). Two-dimensional models assume a homogeneous distribution of q over one of the axes (commonly z as in a wellmixed estuary; Defina, 2000; French and Clifford, 2000). For small tidal embayments, the approach developed by Rinaldo et al. (1999a,b) to calculate water inundation and velocity on tidal marshes is especially useful. This formulation simplified the two-dimensional shallow water equations assuming that a balance exits between water surface slope and friction in the momentum equations. Thus, for water surface, , then @z @z þ z =K z @x @y
ð6Þ
where K represents the forcing and friction terms. An important distinction between estuarine and riverine wetland models is the presence or absence of reversing currents (i.e., tidal pumping). Differences among tidally influenced wetland models lie in how the governing equations are solved (i.e., which numerical method is used), scope of parameters, and functional structure (i.e., how many of the three dimensions are accounted for). Depending on the overall objectives of the model, additional equations can be incorporated to handle differences in state that relate density to temperature, salinity, suspended sediment, and in some cases, biologically active elements (i.e., nutrients, oxygen, and chlorophyll). Wetland hydrodynamic modeling can be classified according to the computational method used to calculate a continuous fluid. All methods depend on the creation of a series of geometrical figures that fill the spatial domain with no overlaps or gaps. Thus, these shapes (i.e., cells) form a volume grid that then is used to solve the governing equations of motion. Such grids can be structured (regular-shaped cells) or unstructured (irregular-shaped cells), with the decision on which type of grid to use depending on the numerical method of choice (Figure 3). The three most common numerical methods are finite difference (Axelsson, 2004), finite element (Brenner and Carstensen, 2004), and finite volume (Barth and Ohlberger, 2004). The finite difference method is mostly used with structured grids, the finite elements method uses unstructured grids, and finite volume method employs the most versatile discretization technique (described below), since the variable of interest is located at the centroid of the cells, regardless of its shape. Structured grids can be stretched over specific regions within the grid of different lengths; as such, they provide greater flexibility in matching land–water boundaries and can assume a curvilinear aspect to the grid (orthogonal grids; Umgiesser et al., 2004; Leupi et al., 2006; Liang et al., 2006; Shen et al., 2006; Peng and Zeng, 2007). However, drawbacks include the added terms in the transformed governing equations that could require up to 30 additional terms because they involve nonorthogonal transformations of Cartesian terms. A common characteristic of all wetland models is that they require a “wet-dry” capability in which the grid or lattice to represent water flow and transport varies spatially under different temporal conditions. Dynamic modeling of ebbing and flooding water requires the solution of governing equations using numerical
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(a)
(b) j+1
y
D
D
D
D
D D D D D D
j j–1
x
i –1
i
i+1 (d)
(c)
Neuse River
d
un
ico ml
So
d
un
Pa
ico ml
So
Pa
Figure 3 Structured and unstructured grids for hydrodynamic modeling. Panels A and C show a conceptual structured grid and its application to the Albemarle-Pamlico Estuary as an orthogonal grid. Panels B and D show an unstructured grid (image courtesy of Jesse Feyen, NOAA).
schemes that differ in simulation of advection and diffusion and how changes in time are handled (Roig, 1989; Elder, 1994). A seminal effort on vertically averaged two-dimensional models to simulate wetting and drying was achieved by Leendertse and Gritton (1971) and Leendertse (1989). A thin layer of water was assumed to remain over the “dry” grid point using the depth computed just prior to the calculation of the negative depth. The mass balance equation is solved for the dry grid point using the previous depth and setting advection and diffusion to zero, thus preserving constituent mass. Presently, two methods of the simulating wetting and drying involve flow cells that can be turned on or off as the depth of flow reaches a critical level (White and Day, 1998; Reyes et al., 2000, 2004a; Oey, 2005). Turning flow cells on and off can violate mass conservation (Roig, 1989). The second method reformulates the numerical grid at each time step (Lynch and Gray, 1980; Casulli and Walters, 2000; Chen et al., 2003; Leupi et al., 2006; Shen et al., 2006). Dynamic modeling of estuarine flow requires the solution of partial differential equations. The numerical solution of differential equations is by the discretization of the governing equations. Discretization is the process in which an ordinary or a partial derivative is approximated by difference equations that compute an approximation of the interval as it approaches zero. There are three approaches to discretizing a given equation: (1) finite difference solutions, (2) finite element solutions, and (3) finite volume solutions. These three techniques are
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roughly comparable in complexity when it comes to their implementation and usually yield identical discretized formulas. The criteria used to determine which solution to use depends mostly on the spatial geometry of the problem and computational limitations.
4.1. Finite difference solutions The finite difference technique consists of transforming the partial derivatives in difference equations over a small interval. This is done by replacing the derivatives appearing in the differential equation with difference approximations. The implementation of this particular method requires a discretization on a regular geometry. The most frequently used computational grid in finite-difference schemes is a rectangular grid with fixed grid spacing (Wang, 1978, 1992; Wang and Kravitz, 1980; Wang et al., 1990). The grid could be rectangular, but the spacing between the grid points in each dimension is rigid (Figure 3). The definition of spacing between points must take into account the presence of islands and channels, making the grid spacing so small that computations could become impractical. Nevertheless, efforts using this approach to examine fluid dynamics in great detail continue as shown by the regional ocean modeling system (ROMS) in highly stratified estuaries, such as Chesapeake Bay (Li et al., 2005). Cheng et al. (1993) developed the semi-implicit, finite-difference tidal residual intertidal mudflat (TRIM) model to simulate wetting and drying. Depths are defined at midpoints of a rectangular mesh, providing a 30% higher resolution than if the depths were defined at the corners of the mesh. When the total depth along a side goes to zero, no mass flux is permitted through that side. The authors stated that this approach conserves mass. Recently, these authors (Cheng and Casulli, 2001) introduced an unstructured grid version named UnTrim that allows for computation in an orthogonal grid. Subsequently, Shen et al. (2006) used this model to analyze a “100-year storm” event (Hurricane Isabel) for Chesapeake Bay, demonstrating the utility of this approach for large regions. Roig (1989, 1996) also modified the finite element RMA-2 model to simulate wetting and drying. Domain coefficients based on water surface elevations were added to the solution to identify partially wet elements. Larger spatial grids were possible on the tidal flats, which decreased computation times (Elder, 1994). However, problems were noted with a deforming grid. Storage volumes appeared discontinuous, a vertical wall effect occurred, and velocities were unrealistic near dry areas (Roig and King, 1992). Another example of a fixed-grid approach is SIMSYS2D (Leendertse, 1987, 1991), a two-dimensional circulation and water quality model with an implicit solution scheme. The model solves vertically averaged equations of long-wave motion and equations for the transport of heat, salt, and other water quality constituents over a rectilinear grid. The original SIMSYS2D was modified into a two-dimensional, vertically integrated model, named the surface water integrated flow and transport (SWIFT2D) model (Swain, 2005). SWIFT2D is now part of a larger project entitled “tides and inflows in the mangrove ecotone (TIME).” The TIME project
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for the Everglades (Schaffranek et al., 2001) is an ambitious approach to massively integrate several modules. In addition to SWIFT2D, a three-dimensional, variable density ground water flow model SEAWAT (a modification itself of the Modular Ground Water Flow model; MODFLOW) and a modular solute transport model (MT3D) are being integrated for a comprehensive accountability of the water cycle. Objectives of the integrated TIME model were to simulate flow exchanges and dissolved salt fluxes between the surface and ground water systems of the landmargin interface of the Everglades with Florida Bay (Schaffranek et al., 2001). The integration of SWIFT2D and SEAWAT was tested for the lower Everglades (Langevin et al., 2004) to quantify flow and salinity patterns for a 7-year simulation period. Results indicated the high degree of connectivity of the surface and ground water flows and the importance of topographic features in redirecting and impeding water flow. Models for ground water flow across wetlands are dominated by the original MODFLOW by Restrepo et al. (1998) with sophisticated developments ranging from code optimization (Harbaugh, 2005), applications to estuaries (Langevin, 2001), and combinations with other models to compute solute concentrations (Langevin and Guo, 2006). All of these efforts have contributed to the creation of a robust series of studies using this finite-difference model.
4.2. Finite element solutions When the domain of interest is complex or changes through time or the desired precision varies over the entire domain (coastal areas vs. open ocean), the approach is to subdivide it into a series of smaller regions on which differential equations can be used for their solution (Figure 3). Each region is then referred as an element, each of which is connected to specific points (nodes) that do not require them to be regularly spaced into the so-called grid or mesh. Additionally, for coastal areas and marsh platforms with complicated geometry and large changes in bottom depth, it is difficult to resolve simultaneously the water column equally well and efficiently in both shallow and in deep regions of a basin using traditional depth layering approaches (equal number of layers, equal depth for each layer). A better simulation for mixed layers is possible when the vertical coordinate system takes into account changes in topography. Such is the case for the sigma coordinate system. In a sigma coordinate system, the number of vertical levels in the water column is the same everywhere in the domain irrespective of the depth of the water column. However, each depth layer is normalized across the sigma coordinate system. This simplifies the computation on the w derivative [see Equation (5)], while the layer thicknesses can vary widely from node to node (Casulli and Cheng, 1992; Ezer and Mellor, 1997). An example of an unstructured grid is the estuarine coastal and ocean model (ECOM-3D), a horizontal and vertically explicit model to simulate the effect of tides, winds, and density gradients on water levels and three-dimensional currents (Blumberg and Mellor, 1980, 1987). The model uses a sigma grid stretched on the vertical plane and curvilinear coordinates on the horizontal plane. A more popular version of this model is the Princeton Ocean Model (POM) that has been used to
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investigate real-time coastal forecasting in the Mississippi Sound (Szczechowski and Carron, 1999). It is important to note that wet and dry capabilities for POM were not implemented until recently (Oey, 2005). Coastal flooding has become a critical issue to be examined under the forecasted consequences of climate change. Among models approved by the Federal Emergency Management Agency (FEMA; Bellomo, 2004) for flood insurance purposes, two can calculate two-dimensional flows, are public domain, and incorporate drywet routines. One is a three-dimensional hydrodynamic model (TABS) and the other is the advanced circulation model (ADCIRC). The TABS-MD model simulates steady and dynamic flow, water-surface elevations, sedimentation, and constituent transport (King, 1982, 1988). It is a finite element model (Thomas and McAnally, 1985) that computes steady or dynamic simulations of three-dimensional flow, salinity, and sediment transport. TABS-MD is the preferred tool used by the US Army Waterways Experiment Station and the COE New Orleans District to simulate sediment loads and storm surges on the Mississippi Delta. A new version of this code has been parallelized to decrease computational time (Rao and Medina, 2006). In an effort to integrate results, a computational environment (i.e., Graphic User Interface) was developed to enhance visualization and facilitate preprocessing and postprocessing. The surface water modeling system (SMS) acts as a comprehensive environment for several hydrodynamic modeling efforts. The USACE-ERDC supports TABS-MD and ADCIRC (Scientific Software Group, 2007). The Advanced Circulation Model (Luettich et al., 1992) uses an interesting feature by combining a grid space discretized using finite element methods, where time is solved using a finite difference method. It increases the computational performance and allows for highly precise calculations of water elevations (Westerink et al., 2004). The ADCIRC model currently has been optimized for performance under parallel computer architectures (Ceyhan et al., 2007).
4.3. Finite volume solutions The finite volume method is similar to the finite difference method in which values are calculated at discrete places on a meshed geometry. The finite volume method evaluates the small volume surrounding each centroid on a grid (Letter D on panel B, Figure 3). Using this method, the volume changes computed by a partial differential equation can be evaluated as fluxes at the surfaces of each finite volume. In this method, volume integrals are converted to surface integrals in a partial differential equation that contains a divergence term. These terms are then evaluated as fluxes at the surfaces of each finite volume. This method is conservative because the flux entering a given volume is identical to that leaving the adjacent volume. Another advantage of the finite volume method is that it is easily formulated to allow for unstructured meshes. The MIT General Circulation Model (Marshall et al., 1997a,b) was the state-of-the-art in finite volume modeling used to simulate three-dimensional oceans. However, as it based on a rectangular grid, it is not easily implemented to estuaries and coastal wetlands. In response, Chen et al. (2003) developed an
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unstructured grid, finite-volume, three-dimensional primitive equation coastal ocean circulation model (FV-COM). The computation of fluxes in an integral form for the equations of motion allows grid flexibility (finite-element method) to be combined with numerical efficiency (finite-difference method). The computation of momentum, mass, dissolved particles, and even heat can be achieved with minimal computational load and high performance. It is particularly useful for coastal wetlands because it includes a mass conservative wet/dry point treatment technique. In combination with FVCOM, different submodules include threedimensional passive tracer equations with inclusion of sinking, sedimentation, and resuspension processes capable of simulating suspended sediments and other passive organic particles. Because the flux entering a given volume is identical to that leaving the adjacent volume, these methods are mass conservative. The clear advantage of the finite volume method is that it can easily be adapted to both structured and unstructured grids.
5. ECOLOGICAL MODELS Ecological models incorporate the biotic component of a wetland into the simulation. The conceptual and mathematical approaches vary according to the amount of data available and the goals related to the understanding of plant or animal population structure, distribution and life history, and the relationship of these to habitat conditions. Recently, research on ecological models has changed focus from attempting to understand how the environment affects populations to how the biota can affect their environment. These types of models, which incorporate biotic feedbacks into environmental parameters, can be classified into two types: (1) those focused on abiotic characteristics (Section 7) and (2) those that maintain a focus on biotic responses (Section 8).
6. I NDIVIDUAL -BASED MODELING For certain populations, it is important to account for the particular characteristics of individuals because details of behavior and physiology of these organisms and their local interactions with other species can determine the behavior, responses, and future status of the population. In such cases, the individual-based modeling (IBM) approach can be used. In this approach, both the environment and all organisms, or a representative sample of all organisms in a population or community, are modeled as individuals, usually by means of a combination of probabilistic rule-based and process-based modeling. In this manner, tracking of individual organisms is accomplished, as opposed to other modeling techniques in which average population responses are computed.
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Individual-based models are commonly “spatially explicit,” which means that the location of each individual in a realistic environment is “tracked” through time. Although of obvious use for motile organisms, it can also be used for nonmobile organisms (e.g., trees or oysters) to determine effects of environmental gradients and their response. Spatially explicit individual-based models for wetlands cover a broad spectrum of spatial scales. They have been applied to trees (baldcypress; Xiao et al., 2002) and to a wide variety of fauna, including brown shrimp using marshes as nursery grounds (Haas et al., 2004), to mammals and birds (Abbott et al., 1997; DeAngelis et al., 2000). Chau (2006) provided a more detailed review of this approach. Two models for mangroves in Southwest Florida have been developed using this IBM approach, based on the classical forest-gap model (Shugart, 1980, l984). These are the MANGRO-SELVA model (Doyle et al., 2003a,b) and FORMAN (Chen and Twilley, 1999). Both simulate interspecific competition for Avicennia germinans, Laguncularia racemosa, and Rhizophora mangle. The MANGRO-SELVA (spatially explicit landscape vegetation analysis) accounts for the effects of windthrow and hurricane disturbances via a probability function. This information is then used by the MANGRO model to predict the status of intertidal mangrove trees based on a series of unique sets of environmental factors and forest history. The FORMAN model (forest mangrove) accounts for population changes under salinity and nutrient gradients (Chen and Twilley, 1999). This model simulates individual tree growth over a long period of time (e.g., 500 years) to predict how the mangrove forest will respond to environmental variability such as sea level rise, changes in nutrient availability, and forecasted climate change. Simulation of populations using IBM methods differs from traditional ecosystem approaches (response of the average population levels) in that the latter allows for environmental conditions to be modified as the population responds, thus creating a feedback between the physical and the biological components.
7. ECO -GEOMORPHOLOGICAL MODELING This type of spatially articulated model aims to simulate how marsh platforms and channel network on tidal marshes evolve. The abiotic and biotic feedbacks are generated from the assumption that water flow and velocity are reduced within vegetated zones, allowing for suspended sediments to settle, accrete, and increase platform elevation, thereby reducing tidal inundation. Reduction of inundation and sediment deposition should in turn affect plant biomass and be conducive to faster soil compaction and lower elevation, thus reinitiating this morphologic cycle. To simulate this cycle, two main approaches have been tried. One aims to describe in detail soil dynamics in three-dimensional fashion, examine large spatial scales, and specifically investigate how plant tussocks affect tidal flows (Temmerman et al., 2007). The other approach relies on a simplified two-dimensional hydrodynamics [see Equation (6)] and examines long-term effects (hundreds of years) on platform evolution (D’Alpaos et al., 2007).
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Temmerman et al. (2005, 2007) utilized a finite-difference model (Delft-3D; http://delftsoftware.wldelft.nl/) to explicitly account for the effect of stem density on flow. In combination with advection–diffusion equations for sediment transport and stem growth density [see Equation (7)], the authors examined vegetation growth and tidal network formation. Temporal change in stem density was expressed as a function of plant establishment (est), lateral expansion to neighboring areas via diffusion (diff), a logarithmic growth equation for present biomass (growth), mortality by flow stress (flow), and mortality due to inundation (inund). @nb @nb @nb @nb @nb @nb þ þ = @t @t est @t diff @t growth @t flow @t inund
ð7Þ
After 30 years of simulations, Temmerman et al. (2007) concluded that, contrary to the view that marsh plants protect soil from water erosion, their results indicated an acceleration of channel incision and erosion in areas adjacent to plant clumps. Several efforts to examine the negative feedbacks between surface elevation and its inorganic accretion have been based on the work by Rinaldo et al. (1999a,b) and Defina (2000) on the use of two-dimensional shallow wave equations for partially dry areas. Assessment of the effects of long-term accelerated sea level rise had been done from experimental (Kirwan and Murray, 2007) and empirical perspectives (Fagherazzi et al., 2004; Marani et al., 2004; Silvestri and Marani, 2004; D’Alpaos et al., 2006, 2007; D’Alpaos and Defina, 2007). All of them integrate water, sediment, and biomass dynamics for small areas (100s of square meters). Critical to the computation of deposition flux on these models was the determination of sediment trapping by plant biomass. D’Alpaos et al. (2007) assumed the following relationship: 8 > <
zmax zb B1ps = zmax zmin > : 0
! If zmin zb < zmax
ð8Þ
If zb < zmin or zb zmax
where local peak season biomass (Bps) is a function of salt marsh elevation (Z).
8. ECOSYSTEM-LEVEL M ODELING The overall objective of an ecosystem wetland model usually is to incorporate physical processes in order to understand and forecast the resultant effects on biological responses. Some of these models aim to find management solutions to problems such as changes in water quality and land loss. Thus, the models require the coupling of hydrodynamics, water quality parameters (e.g., sediment loads, nutrients, and other biologically active constituents), and organisms. The coupling
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of hydrodynamics and other model components can be made via direct or indirect approaches. Direct approaches use simultaneous flow and water quality simulations, that is, both reside in the same code. This approach allows for explicit feedback mechanisms and interactions. The indirect approach uses first a flow calculation that then is used as input for the water quality simulation. This latter approach is computationally more efficient but is limited to the available interactions among the different temporal and spatial scales; it cannot take into account feedback of changes in biota on hydrology. Another disadvantage is the need to store results from the circulation code to be used by the model components that might tax the operating hardware. A direct approach to implementation requires the use of similar-sized grids for flows and suspended constituents. An exception is the model developed by Reyes et al. (2000) where the authors combined a large size (10 km2) mesh for flows with a smaller one (1 km2) for transport. One disadvantage of this approach is the need to use the same time step, which results in long computational times. Callaway et al. (1996) and Rybczyk et al. (1996, 1998) examined the dynamics of plant–soil interactions in a model spatially articulated in the vertical dimension. The amount of surface soil deposited (both organic and inorganic matter) is accounted for during each time step (1 month). Surface sediment is then moved to another storage representing its downward movement to simulate compaction. In total, there are 18 additional storages, each with a different sedimentation rate, representing sedimentation processes from shallow accretion to deep compaction. Although limited to a particular site in a geographical plane, this model is valuable for examining vertical spatial dynamics. Wetland habitat evolution over decades has been the primary concern for understanding system functioning and management. High rates of coastal land loss over the past half century have prompted much research on (1) the root causes of this land loss and (2) management approaches to address this problem. Long-term spatial simulation objectives include the interactions among factors such as sea level rise, subsidence, climate variability, accretion, wetland elevation, wetland health and productivity, and water levels important to determine long-term sustainability of coastal wetlands. Human impacts such as river channeling, wetland impoundment, and canal construction have altered hydrology, salinity, and sediment dynamics of tidal wetlands. An initial long-term scale landscape model was developed as the coastal ecological landscape spatial simulation model (CELSS; Sklar et al., 1985; Costanza et al., 1990; Costanza and Ruth, 1998). Model forcing includes subsidence, sea level rise, changes in river discharge, and climate variability. The modeled region was the southeast portion of the Acadiana Bay in Louisiana. An expanded version of the CELSS model is the Barataria–Terrebonne ecological landscape spatial simulation model (BTELSS; Reyes et al., 2000) that was used to reproduce historical trends in land loss and habitat change from 1956 to 1988 for coastal Louisiana, and then used to predict future trends. This modeling effort implemented an explicit hydrodynamic module, improved ecological algorithms for primary production, and capacity for habitat switching.
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Biomass productivity was expressed as a function of several environmental parameters. dB = pB dt
ð9Þ
p = P ð B þ B þ BÞ c P = P F S J cmax where , , and are translocation, litterfall, and respiration rates, respectively. To represent feedbacks between plant biomass and environmental parameters, the computed biomass amount contributed to the determination of the resulting habitat types after the yearly habitat switching algorithm. The new habitat type in turn modified elevations throughout the model domain. Specific scenarios examined with the BTELSS were different weather patterns and several management alternatives to evaluate environmental impacts analyzed. Several similar efforts have been developed for other watersheds in Louisiana. Martin et al. (2002) built a model to analyze the effects of large river diversions on wetland restoration by implementing, among other things, a sediment accretion module. Vegetation responses to the effects of small diversions, river forcing, and sea level rise also have been examined (Reyes et al., 2003, 2004a,b). The CELSS/BTELSS dynamic simulation models were constructed such that each grid cell would be connected to its four nearest neighbors by the exchange of water, solutes (salts, nitrogen, and phosphorus), and suspended materials (organic and inorganic sediments) (Sklar et al., 1985). The buildup of land or the development of open water in a cell depends on the balance between net inputs of sediments and local organic matter accumulation on the one hand and outputs due to erosion and subsidence on the other. This conceptual approach was further developed (Fitz et al., 1996) and has been implemented for river-influenced watersheds (Voinov et al., 1999), with an economic emphasis (Costanza et al., 2002), and for water quality interactions in South Florida (Fitz et al., 2004). An integrated modeling system, known collectively as the across trophic level system simulation (ATLSS), determines the effects of hydrology (water depth and hydroperiod) and vegetation type on the suitability of habitat for a suite of important species (e.g., snail kites, Florida panthers, and alligators) and functional groups (wading birds), as well as the population dynamics of other taxa (e.g., small fishes). The ATLSS models use a spatial scale of resolution of 100 100 m cells on the Everglades landscape (DeAngelis et al., 1998). In these models, hydrology is the main driver and there are no feedback effects of biota on hydrology. Modeling two-way coupling with the same time and spatial resolutions for two processes may give more robust results but also requires more computational effort. Preliminary
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results indicate the importance of the interactions between hydrodynamic and ecological processes (DeAngelis et al., 2000). The ATLSS project is an ambitious modeling effort that aims to integrate spatially explicit information at different scales for different populations. Among its most innovative approaches is the coupling of hydrological modeling with averaged population models and IBMs (Fleming et al., 1994; Curnutt et al., 2000; Gaff et al., 2000; Comiskey et al., 2002). Each module has been developed independently but all with the aim for incorporation into an overall management system for South Florida.
9. D ESKTOP D YNAMIC M ODELING This type of modeling is a hybrid approach in which some part of the ecosystem modeling effort is based on decision rules rather than mathematical calculations. Mainly oriented for management decisions, this type of modeling approach combines all available modeling information and results, develops a geographical interface, and queries groups of experts as to the most likely outcomes of different mathematically based scenarios. This combination of information is highly useful for specific issues, such as assessing sea level rise vulnerability or longterm river diversions. As long-term assessment evaluation tool, the Coastal Louisiana Ecosystem Assessment and Restoration Program developed a modeling approach to support restoration science needs to rehabilitate the Mississippi River Delta and Chenier Plains (Twilley, 2003; Twilley and Barras, 2003). This modeling approach uses a combination of modules to predict physical and geomorphic processes, water quality conditions, and wetland vegetation succession. For large spatial scales and to examine global patterns of coastal wetlands under diverse scenarios, the wetland change model was developed (McFadden et al., 2007). The goal of this model was to provide with rapid means for assessment of wetland transition and loss. Based primarily on a series of decision rules, model results were corroborated by comparison with other wetland models.
10. CONCLUSIONS Wetland landscape spatial modeling is an active research field in which new technologies and theoretical developments are being used to expand understanding of ecosystem response to disturbance, long-term cumulative impacts, and support management decisions. The diversity of approaches to moving water through a vegetated area and their interaction demonstrates the complexity of this problem. Table 1 identifies some of the important characteristics of each of the different modeling approaches. The quantity of literature demonstrates the scientific and managerial importance of the subject. It was the purpose of this chapter to review some of the recent advances in spatially explicit landscape wetland models, but not
Structural and functional characteristics of wetland dynamic models
Type
Attribute
Mathematical method
Data needs
Computational requirements
Temporal Scale
Physical models Hydraulic
Mass balanced
Discrete mathematics – linear and Boolean algebra Nonlinear differential equations
Medium
Medium
Centuries
High
Highly intensive
Annual, decades
Mass balanced, stochastic
Probabilistic rule based
Medium
Intensive
Decades
Mass balanced, deterministic Mass balanced, deterministic
First order & partial differentials Probabilistic, partial differentials
High High
Intensive Intensive
Centuries Decades
Hybrid approaches, short term or static simulations, and expert systems
Discrete mathematics – linear algebra, probability theory, Boolean algebra
Medium
Low, medium
Centuries
Hydrodynamic Ecological Models Individual-based models Geomorphologic Landscape/ Regional Desktop
Mass balanced, momentum conservation
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Table 1
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be exhaustive in covering the topic. The focus was on models that incorporate ecological dynamics and feedbacks into the landscape to represent ecosystem processes. In most cases, the watershed and local/regional models incorporated new modeling approaches on how to link ecosystems, communities, and populations to the environmental tableau.
ACKNOWLEDGMENTS The author thanks the editors, Don DeAngelis, and an anonymous reviewer for thoughtful comments that improved the clarity of the manuscript.
REFERENCES Abbott, C.A., Berry, M.W., Comiskey, E.J., Gross, L.J., Luh, H.K., 1997. Parallel individual-based modeling of Everglades deer ecology. J IEEE Comput. Sci. Eng. 4, 60–72. Arnold, J.G., Srinivasan, R., Muttiah, R.S., Williams, J.R., 1998. Large area hydrologic modeling and assessment Part 1: model development. J. Am. Water Resour. Assoc. 34, 73–89. Axelsson, O., 2004. Finite difference methods. In: Stein, E., de Borst, R., Hughes, T.J.R. (Eds.), Encyclopedia of Computational Mechanics. Wiley InterScience, Weinheim. Barth, T.J., Ohlberger, M., 2004. Finite volume methods: Foundation and analysis. In: Stein, E., de Borst, R., Hughes, T.J.R. (Eds.), Encyclopedia of Computational Mechanics. Wiley InterScience, Weinheim. Bellomo, D., 2004. Policy for Accepting Numerical Models for use in the National Flood Insurance Program. In: Federal Emergency Management Agency (Ed.), Federal Emergency Management Agency, Department of Homeland Security, Washington, DC. Blumberg, A.F., Mellor, G.L., 1980. A coastal numerical model. In: Sundermann, J., Holz, K. (Eds.), Mathematical Modeling of Estuarine Physics. Springer-Verlag, New York, New York, pp. 202–218. Blumberg, A.F., Mellor, G.L., 1987. A description of a three-dimensional coastal ocean circulation model. In: Heaps, N. (Ed.), Three-Dimensional Coastal Ocean Models, Coastal and Estuarine Sciences. AGU, Washington, DC, pp. 1–16. Brenner, S.C., Carstensen, C., 2004. Finite element methods. In: Stein, E., de Borst, R. Hughes, T.J.R. (Eds.), Encyclopedia of Computational Mechanics. Wiley InterScience, Weinheim, pp. 73–118. Callaway, J.C., Nyman, J., DeLaune, R.D., 1996. Sediment accretion in coastal wetlands: a review and a simulation model of processes. Curr. Top. Wetl. Biogeochem. 2, 2–23. Casulli, V., Cheng, R.T., 1992. Semi-implicit finite difference methods for three-dimensional shallow water flow. Int. J. Numer. Methods Fluids 15, 629–648. Casulli, V., Walters, R.A., 2000. An unstructured grid, three-dimensional model based on the shallow water equations. Int. J. Numer. Methods Fluids 32, 331–348. Ceyhan, E., Basuchowdhuri, P., Judeh, T., Ou, S., Estrade, B., Kosar, T., 2007. Towards a faster and improved ADCIRC (ADvanced Multi-Dimensional CIRCulation) model. J. Coast. Res. 50, 1–6. Chau, K.W., 2006. A review on the integration of artificial intelligence into coastal modeling. J. Environ. Manage. 80, 47–57. Chen, C., Liu, H. Beardsley, R.C., 2003. An unstructured grid, finite-volume, three-dimensional, primitive equations ocean model: application to coastal ocean and estuaries. Am. Meteorol. Soc. 20, 159–186. Chen, R., Twilley, R.R., 1999. A gap dynamic model of mangrove forest development along gradients of soil salinity and nutrient resources. J. Ecol. 86, 37–52.
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SUBJECT INDEX
Aboveground litter, 409 Accommodation space, 45–6, 76, 89–90, 110, 305–6, 875–6 Accretion, 92–3, 310, 410 Acetylene reduction rate, 353 Advection, 169, 174, 216–17, 224, 284, 347–8, 466–7, 470, 897 Aerobic decomposition, 92, 450 Aggregate, 94, 171–2, 322, 324, 434, 595–6, 704 Agricultural crop, 143–4, 719–20 Air–sea interface, 212, 216 Albedo, 148, 215, 222–3 Allochthonous, 48–9, 52, 70, 110, 327, 381, 392–3, 447, 458–9, 549–50, 659–61, 673 input, 673 Allometric, 617–18, 650–1 Alluvial: fan, 79 floodplain, 70 Ammonification, 674–5 Ammonium, 134, 353–4, 357, 384, 392–3, 464, 601–2, 722 oxidation, 350–1, 461f, 462, 669 Anammox reaction, 357 Animal burrow, 47–8, 236, 251, 351, 656 fodder, 721 Annual cycle, 136–7, 144–5, 359 Anthropogenic: activities, 95–6, 296–7, 690 contaminant, 404 pollution, 425–6, 716 pressure, 81, 83 Aquaclude, 771 Aquaculture, 31, 791 shrimp, 787, 792, 794 Aquatic plant, 148, 519–21 Aquifer, 456–7, 493–4, 509
Autecology, 790 Autocompaction, 92, 575, 871, 873 Autoconsolidation, 194 Autotrophic system, 379–80 Backmarsh, 196–7 Bacteria: aerobic, 606–7, 628 anaerobic, 392, 628 denitrifier, 392 nitrifier, 392 Bacterial: abundance, 346 biomarker, 382 marker, 382 substrate, 382 Bank: complex, 177–8, 190 mitigation, 808 overhanging, 190–1 seed, 816, 824 U-shaped, 190 Bar: alternate, 193 point, 191f, 193–4, 311, 415–16 Barrier island, 93, 98–9, 102f, 298, 346, 595, 715, 805 Barrier lagoon, 107 Beach: ridge, 76–8, 107, 127–8, 130, 136, 571, 599 sandy, 69, 187–9, 199, 214, 310–11, 324–6 Bed: elevation, 284–5, 299–302 friction, 298, 310 morphology, 300–2, 307 slope, 48, 296, 298, 299–300, 305, 307, 310–12
909
910
Bedform, 296–7, 299–300, 305 Bedload: motion, 307–8 transport, 303, 305, 311 Bedrock: impermeable, 498 permeable, 498 Belowground root matter, 409 Benthic: boundary layer, 296–7 chamber, 379–80 flux, 352, 357–8 infauna, 330f, 720–1, 738 microalgal, 50–1, 351–2, 354–6, 358, 361–3, 434, 446–7, 454, 630 primary producer, 361, 364–5 Bioaccumulate, 147–8 Biodeposition, 163–4, 333 Biodiversity, 21–2, 65, 133–4, 159–60, 170, 231, 318, 333, 519, 746–7, 816 Bioelement, 345–6 Biofuel, 721 Biogenic stabilisation, 320–1, 333 Biogeochemical: cycle, 364–5, 382, 465, 536, 555–6, 601, 730, 801–2 cycling, 49, 345–6, 375–401, 426, 446 regulation, 333 Biogeochemistry, 31–2, 52, 54, 333, 383, 393, 445–92, 535–62, 600, 641–83, 771, 807 Biogeographic province, 648 Biological production, 92, 133–4, 723 Biomass: aboveground, 134, 169–70, 378–9, 511, 524, 527, 552, 603, 621–2, 724 belowground, 392–3, 446, 511, 524, 724 productivity, 849–50, 899 storage, 221–2 Biomineral liquid, 721 Biostabilization, 263 Biota, 11–12, 45–6, 96, 255, 317–18, 328, 464–5, 570–1, 627, 752, 754, 897–8 Biotic: factor, 1–2, 45–6, 55, 89–91, 144–5, 170, 328–31
Subject Index
interaction, 95, 605, 869–70, 874 Bioturbation, 1–2, 10–11, 199, 236, 300, 333, 346–7, 452, 470, 689 Bioventilation, 347–8 Bivalve, 20–1, 137–8, 331–3, 778–9 Black spot, 327 Bore, 244, 250, 321–2, 568 Boreal forest, 138–40 Bottom roughness, 241 Bowen ratio, 217, 224 Brackish communities, 69 Breakwater, 742 Breeding: area, 136–7, 140–4 ground, 36–7, 140–1, 143–4, 146, 744–5 Burial, 20, 380, 455, 464, 471, 477, 553–4, 772, 848–9 Buttress, 9f, 66 Calcareous mud, 70 Canopy, 18–20, 168–9, 239–41, 617, 626, 729 Carbon: biogenic, 659 biogeochemistry, 536 budget, 50–2, 537f, 663, 671–2 cycle, 380 dioxide, 20, 52, 55, 80, 744 dissolved inorganic, 50–1, 349, 504, 508, 536, 538, 662–3 dissolved organic, 349–50, 352, 358–9, 382, 453, 494, 537–8, 662–3, 671–2 export, 663, 671–2 fixation, 326–7, 616, 621–2, 642–3 input, 536, 663 output, 538 Carbonate bank, 74–5 Cave, 202 submarine, 570–1 Cellular biosynthesis, 352, 359–60 Cellulases, 380–1 Channel: bank, 163–4 bottom, 178, 185, 199, 410–11, 862 head, 198 migration, 78–80, 192–3, 304, 311, 865
Subject Index
navigation, 109, 404 network, 48, 81–2, 159–60, 165, 171, 176, 270–1, 409–10, 846, 896 pan, 201–2 tidal, 9–10, 48, 78–9, 159–63, 171, 177–8, 187, 189, 217–19, 281–2, 299–300, 305, 584, 597–8, 671, 814, 824, 846–7, 849–50, 855–6 wandering, 407 Chemoautotroph, 352, 447–8 Chemoautotrophic process, 350–1 Chemoautotrophy, 50–1, 453 Chenier, 78–9, 310–11, 404f, 861, 900 ridge, 76–8, 81, 569, 582 Chlorophyll a, 346, 353, 356, 362–3 Chloroplast, 584 Cholera, 38 Clay, 69–70, 144–5, 187, 202, 236, 283, 297, 299–300, 346, 347–8, 412, 494, 568, 574, 576, 595, 601–2, 716–17, 719, 814–15, 838 Cliff, 10–11, 108–9, 193, 302–3, 323–4, 409, 599 recession, 297, 303 Climatic change, 32–3, 148, 425–6, 512, 593–4 Clonal reproduction, 134–6 Cloud, 212, 233 Coast: active margin, 594 barrier, 119–21, 416 coastline, 48, 51, 74, 91, 98–9, 105–7, 121, 124, 133–4, 235, 296, 310, 566–8, 594, 598, 641–2, 839, 856, 861, 869–70 collision, 594 convergent, 594 embayment, 297–8, 303, 596, 860–1, 866–7 environment, 96, 148–9, 169, 185, 219, 232, 416, 577, 595, 715–16, 722 erosion, 46, 131–2, 148, 246, 256, 302, 569–70, 616, 695–6, 766–7 marginal sea, 337–8, 594–5 open, 98, 127–8, 236, 244, 298 passive margin, 594–5
911
plain, 45, 65–6, 72, 80–1, 99, 101, 121–2, 127, 131–2, 306–7, 595, 599, 867 progradation, 71, 110, 571 squeeze, 40, 80–1, 777–8 trailing edge, 71, 594–5 Coastal: biogeochemistry, 377–8 defence, 56, 95–6, 716–17, 766–7, 770, 777–8 ecology, 317–18, 699 eutrophication, 327 lagoon, 32, 159–60, 189, 264, 272, 694, 860–1 plain, 45, 65–6, 72, 80–1, 99, 101, 121–2, 127, 131–2, 306–7, 595, 599, 867 protection, 231, 249, 767–8, 770 tundra, 46, 119, 131–2 water, 6–10, 20, 33–5, 38–9, 137–8, 216, 219, 233, 243, 252, 297, 349–50, 362, 501–2, 595–6, 662, 664, 671, 702, 748, 772–3 Coefficient, heat exchange, 215–16 Cohort approach, 842–5 Colonization, 18, 70, 78–9, 98, 162, 200, 220, 333, 382, 508, 523, 627, 723–4, 726, 754, 774, 778, 797, 816, 827, 864–7, 872 Communities: fen woodland, 90, 111 tidal flat, 353 Compaction, 49, 57, 82, 144–5, 190–1, 199, 283, 410, 416, 455, 596, 813, 836, 838, 845, 848, 898 Conductivity, 127, 129–30, 137–8, 215, 251, 458, 771, 814–15 Coral reef, 256, 379–80, 570 Course: blind-ended, 196 development, 189, 196–8, 203 equilibrium, 194 evolution, 203 formation, 198 head, 194 initiation, 186, 196–8 tidal, 9–10, 47, 185–209, 217–19
912
Crab burrow, 9–10, 200, 202, 602, 628–9, 662–3 Creek, 93, 311, 407–8, 769 head, 194–6 tidal, 311, 407–8 Cross bedding, 305 Cross-section: complex, 190 overhanging, 190 U-shaped, 190 V-shaped, 190 Culvert, 36–7, 745–6 Current: asymmetric, 321–2, 574 bidirectional, 516 longshore component, 307 onshore–offshore component, 307 reversing, 890 ripple, 305 shore-normal, 307–8 speed, 307–10 threshold, 307–8 tidal, 9–11, 48, 165, 190–1, 194, 252, 264, 274–7, 296–8, 303, 305, 308, 310–11, 321–2, 326–7, 568, 574, 596, 846 velocity, 253f, 307, 357, 764–5, 780 Cyanobacteria: benthic, 357 biomass, 353–4 heterocystous, 353 mat, 50, 353–4, 434 Cyclone, 28, 583 Cytophaga-flavobacterium, 350–1 Data gap, 847–8, 850 DDT, 38, 110–11 Decomposition: long-term, 848–9 process, 380–1, 845 Degradation: aerobic, 327, 450 microbial, 538 Delta: growth, 311 progradational, 69–70
Subject Index
Dengue, 36–8 Denitrification, 53, 354, 356–7, 459–61, 550–1, 606, 608, 663, 722, 733 Density, bulk, 322, 464–5, 509–10, 774, 844–5, 868 Deposition, atmospheric, 456 Desorption, 602–3, 605–6 Dessication, 859 Detritivore, 33–5, 380–1, 630 Detritus, 50, 255, 351–2, 380–2, 434, 628, 630, 727–9, 745–6, 779 Dewater, 742 Dewatering, 768, 813, 826 Diaspore, 774 Dieback area, 385–9, 869f Diel, 474–5 Diffusion, 168–9, 174, 264, 279, 284, 347–8, 449–50, 654–5, 897 Dike, 36–7, 187, 196–7, 346, 789, 792, 804 Dispersal, 18, 32, 66, 505, 691, 724–6, 738, 773, 816–17, 867–8, 871 Ditching, 196–7 DNRA (dissimilatory nitrate reduction to ammonium), 357, 464, 551 Drag force, 239–41, 244, 245–6, 250, 254–6 Drainage: basin, 311, 788–9 channel, 93–4, 235, 596–7, 717, 719, 746, 776 network, 50, 186, 264–5, 270, 404, 406–9 pattern, 36–7, 108, 196, 406, 510, 577, 744, 750, 770 Driftsand, 347 Driver, 80, 213, 317–18, 328, 333, 508–9, 623, 744, 765, 862–3, 885, 887, 899–900 Drying, 264–5, 463, 771, 891–2 Dryland, 189, 199 Dune, 98–9, 136, 299–300, 305, 720 Ebb, 9–10, 94, 98–9, 189, 192–4, 198, 198–9, 204, 243, 246, 305, 322, 466–7, 715, 814 Ecogeomorphic, 641–83 classification, 648
Subject Index
Ecohydrology, 233, 256, 508–9, 511 Ecological succession, 13–14, 32, 96, 865 Ecology, 245, 255, 317–18, 324–6, 336–7, 515–16, 699, 738, 764–5, 805 Ecosystem: engineer, 319–20, 322, 748 engineering, 319–20, 320t function, 80, 317–43, 511, 643–5, 648, 649–50, 754, 794, 808, 816 health, 319 process, 58, 333–5, 536, 643–5, 738–9, 848, 900–2 service, 35, 39, 317–19, 320t, 326–8, 336, 703 Ecotone, 604, 715, 733, 892–3 Eddy correlation, 213–14 Effective diffusion coefficient, 347–8 Electron: acceptor, 50–2, 350–2, 354, 384, 449–50, 470–1, 542–3, 545, 654 donor, 450, 453, 459, 542–3, 551 El Nin˜o, 33–5, 38, 624, 632–3 Embankment, 45–6, 109, 256, 576, 724–6, 742, 754–5, 763–4, 768, 770, 773–4, 780 Embryo, 596–7, 770 viviparous, 616 Encephalitis, 36–7 Energy: budget, 211–12, 217, 221–2 convection, 273–4 dissipation, 249, 266–9, 273, 281 sink, 213 solar, 212, 217 Environmental management, 255, 696–7 Envirostratigraphic studies, 404–5 Epibenthos, 329t Epibiont, 625–7, 673 communities, 664 Epifaunal biomass, 331–2 Episediment, 673 Erodability, 192, 198–9 Erosion: headward, 17–18, 165, 192–4, 198, 304 upland, 720 Erosional surface, 406
913
Estuaries, 6–8, 33, 66–7, 69, 74, 105, 110, 121, 186, 192, 214, 219, 225, 240, 264, 272, 297, 321–2, 346, 351, 415–16, 425–6, 464, 478, 493–4, 511, 528, 538, 568, 569–70, 595–7, 630–1, 641–2, 669, 720, 745–6, 749, 752, 767–8, 772, 780, 792, 839, 850, 860–1, 863, 866, 892–3 Euryhaline, 429, 464–5, 467, 473 Eurythermal, 429 Eutrophic, 356–7, 501–2, 519, 536 Eutrophication, 6, 18, 51, 379, 394, 501–2, 603, 690, 885 Evapoconcentration, 220–1 Evapotranspiration, 47, 170–1, 213, 426, 465, 508, 887 Exposure: rate, 213 total, 213 Facies, 322–3, 416, 862 Fatty acid, 382, 468, 630, 632–3 Fauna: arboreal, 631 avifauna, 140, 430 brackish fauna, 137 exfauna, 136–7 infauna, 121, 191–2, 199, 450, 738, 795 introduced, 749 mammal fauna, 146 marine fauna, 137 migration area, 140 migratory, 121, 140 mobile, 795–6 nomadic, 136–7 terrestrial fauna, 136–7 Fecundity, 872 Feeder: deposit, 28, 350, 727 filter, 50, 434 Ferran, 129–30 Fertilization, 31, 605, 650–1, 719, 722, 872, 874 Fetch, 204, 281–3, 303, 323, 768–9 Fish: farm, 386, 389, 390–1 feed, 721
914
Fishing, 32, 147, 404, 696, 795 pressure, 425–6 Fixation rate, 350–1, 353–4, 616, 665t Fjord-head delta, 107 Flat: bare, 162, 302–3 mixed sand–mud, 299–300 mud, 6–11, 33–5, 299–300, 303, 721 profile, 305–6 salt, 416 sand, 305, 357, 406, 500–1, 767–8 tidal,10–11,90,161–4,177–8,199,202,204, 213, 215–19, 250, 264, 277, 281–2, 295–7, 305, 306–8, 347–8, 353, 846 Flavonoid, 721 Floc: aggregate, 322 settling, 321–2 Flocculation, 94, 574 Flood control area, 765–6 Flora: brackish, 38, 69, 90, 100, 136, 519 marine, 136, 427t, 505, 507 terrestrial, 616, 771 Flow: asymmetric, 190, 243, 409–10 asymmetry, 193–4 bidirectional, 186, 516 density, 199 meandering, 193 overmarsh, 196, 333 saturated, 199 sheet, 185, 198, 200, 498–500, 769, 770–1 speed, 410 supercritical, 199 transversal, 193 turbulent, 267 undercutting, 190–1 unidirectional, 186 unsaturated, 199 vector, 410 Fluid mud, 94–5, 642–3 Fluvial network, 185–6, 189–90, 196 Flux: advective, 174, 212, 216–19 conductive, 212
Subject Index
moisture, 214 turbulent, 215, 219–20, 224 vertical, 211, 379–80 Flyway, 140–1, 141f Food: chain, , 255, 335–6, 380–1, 752, 817 resource, 53, 136–40, 143–4, 231, 434, 630, 754, 776 web, 18, 22, 28, 147–8, 232, 335, 430–1, 616, 626, 721, 723 Forage, 35, 429–30, 434, 524, 630 Forest: basin, 234–5, 791–2 B-type, 235, 245, 251 fringe, 234–5, 661, 664 riverine, 234–5, 661 R-type, 235–6, 240, 245–6, 252–4, 256 F-type, 235–6, 244, 247–9, 251 Freshwater: area, 138 flood, 17–18, 79–81, 95–6 input, 1–2, 90, 95–6, 100, 425, 493–6, 502–3, 509, 746, 749, 752 marsh, 31, 107, 129–30, 200–1, 447–9, 545, 546–7, 549–52, 602–3, 865 wetland, 13, 22, 45, 74, 81, 90, 449, 515–33, 535–62, 569, 578–9, 605, 801–31, 888–9 Friction coefficient, 164, 268–70 Frost, 3, 29–30, 66, 126, 264, 856–7, 859, 869–70, 873 Froude number, 199, 305 Frustule, 362 Functional trajectories, 805–6 Gas exchange, 66, 536–7, 575–6, 585, 617, 617–18 Genetic response, 133–4 Geobotanical, 593–614 Geochemical: budget, 456–7 cycling, 49, 345–6, 375–401, 446, 476, 547–8, 601, 730, 801–2 tracer, 456–7, 478 Geoform, 597
Subject Index
Geomorphology, 4, 6, 29, 99, 121, 189, 293–316, 321–2, 403–24, 563–91, 593–614, 808, 870 Germination, 505–7, 691, 722, 724–6, 774 Grassland, 81, 214, 508–10, 618, 717, 765–6, 773–4, 873–4 Grazing pressure, 145, 744–5, 750–1 Greenhouse gas, 148, 536, 555, 745, 772 Groundwater: flushing, 220–1 seepage, 2, 199, 251, 718 Habitat: brackish, 51, 128–9, 137, 511 destruction, 425–6, 626 freshwater, 137, 511, 516, 518, 524, 865 recovery, 144–5, 703–4 succession, 144–5 marine, 51, 507–8, 697–8 saltwater-influenced, 137 Halophyte, 70, 111, 263–4, 502, 599 Halophytic vegetation, 46, 65, 104, 161–4, 177–9, 272, 279, 281, 752 Harbor, 110, 147, 219, 241, 247–8, 256, 393–4, 576, 690, 698–9 Harvest, 20, 143–4, 331–2, 346, 430, 524, 536, 617, 690–1, 695, 787–8, 792 Headward: erosion, 17–18, 165, 192–4, 198, 304 retreat, 191–2, 194, 195–6 Heat: balance, 213–15, 217, 220 budget, 212, 216–20 energy, 47, 211–29 exchange, 211, 215, 216–17, 219, 221–2 flux, 47, 213–19, 223, 224–5 gain, 217–19, 223–4 latent, 211–14, 216–17, 221, 223–4 sensible, 47, 214–17, 223–4 specific, 215 storage, 148 Heavy metal, 51, 108, 505, 752, 771–2, 815, 821 Herding, 404 Heterotrophic microorganism, 350
915
Human settlement, 46, 97, 148, 737–8 Humic: acid, 50–1, 359–60, 450, 452–3, 544–5, 552, 603 substance, 543–5, 601–2 Hunting, 31, 146–7, 149, 404, 744–5 Hurricane, 14–15, 28, 57, 407, 448, 583, 623, 651–3, 659, 673–4, 788, 838–9, 892, 896 Hydraulic duty, 406 Hydrodynamic forcing, 303, 323, 361 Hydro-isostatic flexure, 72–3 Hydrology, 13, 18, 20–1, 39, 46, 96, 508, 518, 572, 718, 745, 791–2, 808, 813 Hydropattern, 91 Hydroperiod, 91, 281, 409, 554, 655 gradient, 93, 643–5, 649–50 Hyperbenthos, 329t Hypersaline, 18–20, 30–1, 107, 220, 859, 861 Hypersalinity, 18–20, 33–5, 144–5, 255–6 Hypoxic condition, 601–2, 606 Ice: cover, 124, 377–8 floes, 127–8 pressure ridge, 127–8 push, 127–8 rafting, 134 Immobilization, 462, 550, 663–4 Impoundment, 31, 36–7, 584, 787, 898 Infauna: assemblage, 738 burrow, 191–2, 199 Infaunal: biomass, 331–2 organism, 324–6 Inlet, 66, 170–1, 189, 192, 265, 717, 860–1, 865 Insolation, 322 Inter-ridge swale, 128–9 Intertidal: mudflat, 56, 300, 324, 410–11, 503 stock, 137–8 zone, 65–6, 91–2, 98–9, 105, 107, 110, 160–1, 214, 300, 429, 595–6, 715, 859, 870–1
916
Invertebrate: assemblage, 137 intertidal, 137–8, 778–9 Iron: oxidation, 470 oxyhydroxide, 360–1, 553, 658–9 Isostasy, 72–3, 741–2 Isostatic flexure, 72–3 Kelp forest, 318 Lag: scour, 308–9 settling, 308–9 Laida, 119–21 Lake, 46, 70, 119, 130, 362, 571, 583, 776, 818–19, 866 Laminated mud, 300 Land subsidence, 716, 741–2, 836, 837–8 Landuse, 81 Laser granulometry, 412 Layer: anoxic, 384 oxic, 384 Lead, 110–11, 147, 505, 821 Leaf: area index, 617 node, 618 photosynthesis, 617, 621–2 production, 20, 380–1, 617–18 temperature, 47, 214, 217 turnover, 618 Lenticel, 66, 576, 584 Levee accretion, 192–3 Light: attenuation, 31, 617, 620–1 requirement, 378–9 Lithostratigraphic architecture, 50, 416 Litter, 20–1, 53, 255, 333–4, 548–9, 573–4, 608, 661–2, 671–2, 777, 797–8 Litterfall, 20–1, 605, 617–18, 620–1, 899 Longshore: current, 69–70 drift, 578, 720
Subject Index
Macrobenthic communities, 331–2 Macroconsumer, 50, 434 Macrofauna, benthic, 333, 824 Macropore, 47–8, 236 Malaria, 36–8 Managed realignment, 56, 80–1, 505, 722–3, 742, 763–85 Management, 36–9, 51, 54, 331–2, 509, 696, 739, 744–5, 748, 763–4, 885, 898–900 Manganese-respiration, 544 Mangrove: basin, 67–8 canopy, 236, 240, 617, 626, 664 dwarf, 604 ecosystem, 47–8, 80, 231–3, 241, 255, 629–30, 643–5, 787 forest, 23t, 67–8, 74, 80, 83, 231, 565, 568, 572, 597–8 fringe, 67–8, 569–70, 655–6, 664, 673 fringing, 246–7, 429, 568–70, 795–6 inland, 570, 673 leaves, 20, 629, 663–4 overwash, 67–8 pneumatophore, 9f, 66, 240, 244, 657 productivity, 573–4, 642–3, 649–53, 657, 658–9, 673 prop root, 66, 233, 236, 244, 673 riverine, 67–8, 569–70, 651, 659–61 scrub, 67–8, 651–3, 673 stand, 568, 622–4, 627–8 swamp, 6–8, 76, 79–80, 220, 236, 240, 243–4, 247, 569 topography, 233–4, 245 tree, 33–5, 47–8, 232–3, 240–1, 249, 254–5, 571, 616, 624–5, 628, 792, 896 Manning, 245–6, 267, 268–70 Marker horizon, 575, 814, 819–20, 838 Marsh: creation, 55, 502, 508–9, 511, 715–36, 771, 772–3 ditched, 196–7 edge, 50, 283, 303, 406, 432, 499 freshwater, 31, 95–6, 107, 144–5, 447–9, 471–2, 546–8, 550–2, 602, 865
Subject Index
function, 512, 716, 719–20 island, 196 platform, 46, 50–1, 161–2, 167–8, 178, 410, 412, 446, 850 terrace, 407 value, 55, 716, 723, 754–5 Mass wasting, 194–6 Meadow, 2, 18, 49, 127, 334–6, 377–8, 385–6, 689, 695, 789–90 Mean sea level, 4, 6, 9–10, 239, 241, 303, 572, 788, 845 Mechanical transplanter, 722 Meiobenthic communities, 331–2 Meiofauna, epibenthic, 433–4 Mercury, 110–11, 505 Mesocosm, 318, 334–5, 356, 543–4, 629, 698 Metabolism, microbial, 541 Meteorological station, 214 Methane, 20, 52, 349, 351, 542, 545, 654 Methanogen, 453, 542, 555 Methanogenesis, 351, 450, 453, 545, 732–3 Microaerobic zone, 545–6 Microbial: immobilization, 462–3, 550 mat, 337–8, 352–3, 359 primary producer, 350 respiration, 449–50, 452–3, 536, 538, 541–3, 653 Microclimate, 18–20, 212 Micromorphology, 93–5 Microorganism, 96, 350, 384–5, 523, 542–3, 606–7 Microphytobenthic: assemblage, 331–2 photosynthesis, 326–7 Microzone, 453, 463, 656 Migrant, 429 Migration, 7f, 78–80, 98–9, 136–7, 143, 145, 305, 658–9, 769 Mineralization: aerobic, 449, 462 anaerobic, 393, 450 C, 450–2, 543, 545, 732–3 Mire, 134 Mixing length, 268–70
917
Model: biogeomorphic, 178–9 ecogeomorphic, 641–83, 839, 846, 849–50 ecological, 169–70, 895 finite elements, 272–3, 894 geomorphic, 847 geomorphological, 48, 179, 889 hydraulic, 245, 255, 887 hydrodynamic, 160–1, 164, 272–4, 279, 889 landscape, 58, 840 mineral sediment, 839–40, 849 phase-averaged, 272–3 phase-resolving, 272–3 physical, 187, 886f, 887 point, 281, 839–40 sediment transport, 162, 283–4 surface elevation, 57, 833–53 wave, 264–5, 272–4, 281–2 zero-dimensional, 160, 839–40, 842–3, 850 Modeling: desktop dynamic, 900 dynamic, 886–7, 889, 890–2 eco-geomorphological, 896 ecosystem-level, 897 hydrodynamic, 889 individual-based, 895 Molybdenum, 352 Momentum exchange, 267–9, 271–2 Monoclimax, 317–18 Monocultures, 100, 267–8, 720, 747–8 Monsoon, 104–5, 241, 607–8, 862, 867 Morphodynamic, 161–2, 168, 177–8, 272, 274–7, 283, 296–7, 850 Mud: fluid, 94–5, 642–3 mound, 300–2 Necrosis, 583–4 Nekton, 428, 432, 433–4, 727–9 Neotectonic, 597–9 Network density, 196–7 N-fixation, 50, 352, 434, 604, 667t, 754 Niche construction, 320t
918
Nitrate, 50–1, 134, 354, 359, 384, 464, 500–2, 772 reductase, 604 Nitrification, 354, 392, 459, 463, 606, 664–71 aerobic, 602, 606 Nitrite, 352, 354, 462 Nitrogen: cycle, 352, 606, 608 dissolved, 357, 501–2 exchange, 548 fixation, 50, 352, 434, 604, 667t, 754 mineralization, 602 outwelling, 674–5 Nitrogenase, 352–3, 455–6, 554–5, 602, 663–4 Nonpoint source, 702, 719, 752 Nucleotide, 359–60 Nursery, 31–2, 432, 631, 739–40, 789, 896 Nutrient: cycle, 1–2, 382, 392–3, 434, 556 cycling, 28, 392, 527, 805 limitation, 363, 393, 552, 605 loading, 30–1, 51, 477–8, 501–2, 507–8, 527, 547–8 regeneration, 377–8, 392, 664 regulation, 551 resource, 54, 650–3, 673, 772 uptake, 102, 358–9, 363, 392–3, 643–5, 657 Oceanic current, 148, 565 Oil residue, 753 Oligotrophic, 356–7, 359, 382, 392–3, 536, 570, 627, 672 Opal, 362 Organic: carbon content, 138, 298–9 chromophoric dissolved organic matter, 541, 570 compounds, 359–60, 384–5, 454, 732 matter, 20, 351, 723–4, 732 matter transport, 504 particulate organic matter, 20, 28, 297, 504, 550
Subject Index
Organism: coprophagous, 628–9 filter-feeding, 326–7 Orogenies, 121–2 Outwelling hypothesis, 448, 671–2 Overgrazing, 143–4, 750–1 Overhunting, 147, 404 Overwash, 67–8, 659–61, 888 Oxidized zone, 654, 657, 658–9 Oxygen: content, 214 distribution, 348–9, 361, 653 production, 204, 392 supplies, 392, 576, 653–4, 658, 664–9 Oxygenation, 95, 333, 500–1 Paleochannel, 416 Panne, 100 Pattern: hydraulic, 196–7 network, 197 vegetation, 32, 98, 159–60, 196–7 Peat: swamp forest, 45, 69–70, 80 unchanneled, 406, 408 Peatland, 129–30, 148, 415–16 Pectinase, 380–1 Pelletization, 333 Percoline, 199–200 Persistent Organic Pollutants (POPs), 147–8 Petroleum, 147–8 pH, 129–30, 380, 470–1, 542–3, 603, 649–50, 719–20, 771 Phenotypic plasticity, 133–4 Phosphatase, 552–3 Phosphonate, 359–60 Phosphorus: biogeochemistry, 552 cycle, 359 inorganic, 359–60 organic, 361 particulate, 361 sequestration, 477, 552 Photoautotrophy, 446, 462, 475
919
Subject Index
Photosynthesis, 2, 31, 220, 355, 362–3, 476, 545, 584, 616–17, 627 Pingo, 130–1, 224t Plant: communities, 55, 89–118, 220, 494, 717, 720, 724, 773, 815–16, 848 emergent, 18, 211, 427t, 720 floating-leafed, 211 mortality, 390–1 parasitic, 426–8 pioneer, 47, 204, 300, 406–7, 499 root, 129–30, 179, 190–1, 199, 449, 468, 470–1 submerged, 211 succession, 96, 431 Planting units (PUs), 694 Plate tectonic: active coast, 594 convergent coast, 594 passive coast, 594–5, 598–9 trailing edge, 594–5 Ploughing, 774 Pneumatophore, 47–8, 66, 233, 240, 244, 576, 631, 657 Point bar, 415–16 Polar desert, 119, 125–6, 130 Polder, 770 Pollen record, 79–80, 108, 569, 865–6 Pollution: heavy metal, 752, 771–2, 815 oil, 752–3 Pond: anachaline, 570–1 mangrove, 570–1 non tidal, 571 saline, 570–1 Pool: density, 196–7 dissolved calcium-bound, 393 interconnection, 196–7 porewater, 474 Porewater: flow, 348–9 salinities, 595–7, 599, 600–1 velocity, 348
Porosity, 236–8, 240–1, 247, 494–5, 498, 508–9, 654, 846 Precipitation, 29–30, 47, 111, 148, 212, 214, 221, 394–5, 426, 456, 499–500, 606, 671–2, 887 Predation, 50, 146, 328, 432, 434, 626–7, 868 Prehistoric human, 404 Prey: segregation, 626 size, 626 Primary production: gross, 536–7 net, 379, 382, 434, 446, 536–7, 616, 617–18, 618t, 619t, 723–4 Primary productivity: aboveground, 378–9, 527, 617–18, 845, 848 belowground, 378–9, 381, 848–50 leaf, 20, 380–1, 617–18 net, 658–9, 673–4 total, 617–18 total community, 379–80 Production: aboveground, 378–9, 527, 617–18, 848 belowground, 378–9, 381, 848–50 Propagule: distribution, 790 limitation, 789–90 volunteer, 792–3 Prop root, 47–8, 66, 82, 236, 244, 576, 656, 672 Pyrite, 454, 465, 468–71, 606, 658–9 Pyritization, 394–5, 466–7, 470 Radiation: atmospheric long-wave, 215 environmental, 213 incoming, 213 net, 212–13, 215, 216–17, 221–3 reflected, 213 short wave, 213 terrestrial longwave, 215 total exposure, 213 UV-B, 213
920
Radiocarbon, 76, 416, 575, 862, 865 Radionuclide, 110–11 Radiotracer, 455 Radium budget, 456–7, 457t Rainfall, infiltration, 217 Rainforest, 35–6, 318, 572–3, 615, 624–5 Ramsar site, 2 Realignment, managed, 56, 722–3, 754–5, 763–85 Reclamation, 29, 108–9, 346, 737–8 Recolonization, 144–5, 395–6, 690, 700–1, 789, 795 Recovery, 31–2, 144–5, 331–2, 689–90, 693, 789 Recruitment, 53, 428–9, 623–4, 776, 795 Redox: potential, 377–8, 600–1, 603, 607, 654–5, 821 reaction, 476 zone, 653 Reduction: iron, 384–6, 629, 657–8 manganese, 384, 543, 627–8 sulfate, 350–1, 362, 384, 386, 462, 545, 730 Reed: bed, 214 -gathering, 404 swamp, 95–6 Reef: artificial, 690–1 coral, 2, 52–3, 256, 378–9, 570 platform, 70 Reference site, 56–7, 778–9, 808, 819, 824–6 Reforestation, 247–8 Refugia, 143–4, 776 Regeneration, 82, 360, 377–8, 392, 674–5, 816 Regulatory enforcement, 805 Rehabilitation, 687, 693, 695, 696–7, 702–3, 745, 752–3 Remote sensing, 144–5, 267, 334–5 Reoxidation, 380, 386–9 Residence time, 247–8, 252, 349, 461, 597
Subject Index
Residual circulation, 190 Respiration: aerobic, 380, 384, 386, 449–50, 542–3, 654, 657–8 anaerobic, 352, 384–5, 453, 543, 657–9 microbial, 449–50, 452–3, 536, 653 plant, 538 total community, 379–80 Restoration: ecological mangrove, 790 ecology, 791, 794, 796, 797–8 hydrologic, 790, 793–4, 819 methodology, 56, 788 model, 765 project, 31, 702–3, 740–1, 745 site, 698, 773, 789, 813 Restored site, 699, 739, 795, 805, 808 Resuspension, 264, 277, 300, 894 Reynolds number, 247 Rhizome, 136, 390 Rhizosphere, 380, 452, 627–8, 656 Ria, 596 Ricefield, 104, 109 Ridge: chenier, 76–8, 81, 569, 582 sand, 582 shell, 583 shingle, 570, 576 Ripple, 305, 324–6, 348 River: catchment, 258, 409 discharge, 70, 162, 193, 568 eutrophic, 536 input, 298, 349–50, 641–2, 648 oligotrophic, 536 Root: aerial, 576, 584, 616, 664 exudate, 384, 536–8 horizon, 416 hypocotyl, 81 uptake, 392 Rotational slump, 192–3 Runnel, 38, 354–5 Saline intrusion, 81–2, 502, 578–9 Salinity, gradient, 467, 473, 554, 602
Subject Index
Salinization, 81–2, 393–4 Salt flat, 45, 90–1, 220, 873–4 Salt marsh: creation, 55, 509, 511, 772 functional, 511, 739–40 polar, 121 restoration, 737–61, 773 supratidal, 91, 124, 300 temperate, 105–7, 136 tropical, 856 Salt pan: channel, 596–7 primary, 596–7 Sandflat, 406, 408, 415–16 Saturation index, 773 Scarp, 302 Scour hole, 192 Scroll bar, 78–9 Seabed, 297, 310, 323 Sea defence, 56 Seagrass: artificial, 697 bed, 31, 334–5, 377–9, 383–4, 658–9, 700–1, 796–7 decline, 49, 395–6, 689, 696 detritus, 380–2 geographic distribution, 689 rehabilitation, 687, 693, 696–7 restoration, 685–713 seedling, 695, 698 tissue, 378–9 transplantation, 687–8, 690, 698 Sea ice ecosystem, 146 Sea-level: change, 74, 76, 80 relative, 6–8, 90, 575, 875–6 rise, 71–2, 131–2, 554, 580–1, 871 Secondary producer, 138–40 Sedge, 74, 865–6 Sedgeland, 45, 66–7, 69, 78–9 Sediment: accumulation, 22–8, 100, 264, 774, 841–2 biogenous, 383, 555, 600 budget, 6, 108, 110–11, 574–5 carbonate rich, 384, 393, 394–5 contaminated, 752
921
core, 251, 298–9, 864 deposition, 18, 169, 192, 298–9, 537, 604, 814, 896 dredged, 818–19 dynamic, 49 efflux, 358–9 erosion, 18, 29–30, 179, 277, 298, 688, 846, 849–50 grain size, 93–4, 111, 296–7, 381 lithogenous, 346 mineral, 411, 455, 813, 840 muddy, 48–9, 300, 324–6, 498 organic, 29, 176, 673, 842 permeability, 347–8 resuspension, 264, 281–2, 323 retention index, 311 reworking, 303, 333–4 sink, 310–11 source, 22–8, 362–3, 409 strata, 310–11 supply, 76, 305, 409 temperature, 213, 215 terrigenic, 385–6 transport, 11–12, 163, 192, 272, 284, 409–10, 545, 897 waterlogged, 603, 774 Sedimentary record, 108, 296–7, 568 sequence, 76, 296–7, 300 structure, 48, 296–9, 312 Sedimentology, 293–316, 403–24, 563–91, 593–614 Seed: bank, 56–7, 413, 516, 774, 816, 824 germination, 493–5, 505–6, 691 Seedling: growth, 499, 507 mortality, 506, 871 Seepage, pore water, 499 Semi-enclosed sea, 297 Sensitivity analysis, 838–9, 843–4 Sequestration, 467, 477, 552, 730–2 Settling velocity, 169, 281, 321–2, 405–6 Sewage effluent, 719 Shear stress, bottom, , 165, 167, 199, 266, 277, 283 Shell debris, 297, 411, 494
922
Shielding, 213 Shifting baseline syndrome, 425–6 Shoreline recession, 310–11 Silica: biogenic, 362 lithogenic, 362, 363f mineral, 362, 554 Silicate: dissolved, 362–3, 363f terrestrial, 554 transformation, 554 Silicon: biogeochemistry, 554 cycle, 362 Silt, 69–70, 135f, 187, 283, 346, 403–4, 406–7, 410–12, 416, 494, 568, 576, 719, 814–15, 841, 864 Siltation, 6–8, 10–11, 246, 256, 393–4, 576, 658–9, 721 Sinkhole, 571 Soil: anoxic, 134–6, 775–6 biogenic, 31 buried, 416 degraded, 134–6 disturbed, 134–6 fertility, 138–40, 661–2 formation, 31, 659 organic-rich, 195–6, 547–8 production, 169, 263, 283 properties, 214, 510–11, 672–3, 747 reduced, 426, 584, 604–5, 722, 872 resistance, 198 salinity, 22, 111, 214, 220–1, 426–8, 499–500, 648–9, 744, 825, 870, 873–4 water potential, 90 Sorption, 361–2, 472, 476, 552–3, 601–2, 604, 658–9, 730–2 Southern oscillation index, 873 Spatialization, 847, 850 Species: diversity, 66, 136–7, 174, 324–6, 377–8, 519, 522, 625–6, 720, 751, 774, 795 exotic, 103f, 103–4, 425–6, 756 introduced, 55, 108, 746–9
Subject Index
invasive, 31–2, 39, 336, 529, 721, 747, 749, 816, 827 native, 32, 97, 105, 336, 426–8, 524, 724, 747, 808, 827 nonindigenous, 748 pool, 778–9 target, 773 Spill: impact, 752–3 oil, 147–8, 585, 633, 722–3, 726, 752–4 Spit, 79, 130, 596 Stomatal, 584 Stopover, 136–7, 140–1, 143–4 Storm: stormwater, 747, 752 surge, 15–17, 29, 46, 47–8, 119, 136, 196, 295–6, 409, 583, 716–17, 813, 894 Stromatolitic: fossil, 337–8 system, 337–8 Sublimation, 222–3 Subsidence: deep, 93f, 836t, 837–8, 843f shallow, 575, 836, 836t, 838–9, 848 Substrate stability, 95 Success: compliance, 805, 808, 827 evaluation, 802, 805, 806–7, 825 functional, 805 landscape, 805–6, 827 Sulfate reduction, 346, 350–1, 362, 385–6, 388f, 388f, 394, 449, 462, 468, 542, 544–5, 606, 629, 629t, 657–8, 730 Sulfide: invasion, 389–91, 394–5 toxicity, 95, 390–1, 658–9, 673–4 Surface elevation table, 575, 814, 819–20, 873 Surface emissivity, 215 Surf zone, 298 Survival, 22–8, 66, 337–8, 432, 650, 694, 747, 779–80, 856 Suspended: load, 308
Subject Index
sediment concentration, 92, 94–5, 168–9, 204, 284f, 296–7, 303, 305, 321–2, 447–8, 842 Swale, 127–9, 128f, 279 Temperature, 18–20, 66, 163, 215, 357, 386, 464, 615, 743, 890 Terminal electron acceptor, 352, 354, 450–2, 464, 543, 544–5 Terrace, accretionary, 415–16 TFWs (tidal freshwater wetlands), 22, 51–2, 515–33, 535–62, 801–31 Thermal: conductivity, 215 diffusivity, 215, 219 Thiosulfate, 389, 466–7 Threshold, 18, 74, 161–2, 176, 198, 201f, 264, 303, 308–9, 386–8, 499–500, 808 Tidal: action, 202, 235–6, 243, 297–8, 300–2 bankfull, 409 bedding, 298–9, 303, 413 channel, 10–11, 159–60, 163, 849–50, 885 constituent, 243 course, 9–10, 47, 185–209, 217–19 creek, 9–10, 15–17, 81, 189, 192, 235, 246, 302, 311, 867–8 cycle, 67–8, 204, 240, 308, 324, 770–1, 842 discharge, 407–8 diurnal constituent, 243 embayment, 79, 171, 890 flat, 48, 263–91, 293–343, 345–73, 846, 892 fluctuation, 219, 243, 584–5 flushing, 52, 95, 470, 601, 603, 733, 744–5, 794–5, 870 groove, 188f, 188t, 189 gullies, 187, 188f, 188t, 189 inlet, 298 inundation, 52–3, 205–6, 281, 763–4, 896 macrotidal, 69, 124, 429, 862 meander, 193, 407–8 mesotidal, 11f, 124, 297, 428–9, 846
923
microtidal, 6–8, 110–11, 428–9, 802 networks, 46, 889–90, 897 overmarsh, 409 penetration, 200 prism, 161–3, 171, 176–8, 194, 217–19, 409–10, 499–500, 770–1, 872 pumping, 6–8, 70, 890 range, 52–3, 241, 458–9, 516, 572, 599, 872 regime, 50, 241, 409, 870–1 rill, 187–9, 191–2, 199 semi-diurnal constituents, 243 sinuosity, 190, 206, 407–8, 746 spring–neap tidal cycles, 295–6, 303, 308 stages, 215, 217–19, 247 subtidal, 2, 295–6, 324–6, 354–5, 454, 716, 849–50 subtidal populations, 137–8 supratidal, 107, 295–6, 641–2, 862 undermarsh, 409–10 water level, 295–6, 299–300, 307, 452 waterways, 403–4 waves, 94, 250, 310, 311–12 wetting, 409, 463 Tide: highest astronomical, 90–1, 405–6 mixed, 103 neap, 105, 241–3, 303, 403–4, 429, 505 over tide, 310 shallow, 310 spring, 99, 204, 299–300, 429, 599, 776 Tissue, 28, 389, 527–8, 542–3, 874 Topsoil, 221, 510–11, 719–20 Tortuosity, 347–8 Tourism, 378–9, 393–4, 721 Trace metal, 346, 352, 470–1 Transport, advective, 174, 326–7, 348, 356 Trawling, 31, 331–2, 690–1 Tree: mortality, 28, 585, 626 ring, 624 Trophic: food web, 147–8 structure, 335
924
Tropical rainforest, 35–6, 318, 615, 624–5 Turbidity maximum, 94, 537 Turbulence, tsunamis, 47–8 Typhoon, 239, 244, 247–8, 298, 788 Ultraviolet: light penetration, 541 radiation adsorption, 541 Undergrazing, 750–1 Ungrazing, 750–1 Urbanization, 29, 377–8, 740–1, 801–2, 808 Urban stormwater, 752 Valley, drowned bedrock, 596 Vegetated platform, 177, 409 Vegetation: artificial, 280–1 canopy, 263 disturbance, 48, 96, 128–9, 281, 430 emergent, 18, 219, 280, 430 fringing, 428 layer, 280–1 Velocity: distribution, 268–9 field, 168–9, 268–9 profile, 94, 163–4, 279, 280–1 settling, 169, 281, 321–2, 405–6, 410–11 Vitamin, 359–60 Vivipary, 82 Wastewater treatment, 690–1 Water: table, 65–6, 69–70, 96–7, 510–11, 771 potential, 47, 90, 217 stress, 214, 519 vapor, 212 Waterbird assemblage, 779 Waterlogging, 91, 600–1, 606, 655, 814–15 Watershed divide, 174, 176 Wave: action, 48, 147, 273–4, 298, 570, 694 breaking, 272–3, 298, 310–11 damage, 583
Subject Index
damping, 404 energy, 10–11, 280, 298, 595–6, 615, 730, 861–2 energy attenuation, 249 energy density, 273–4 energy dissipation, 249, 273, 281 field, 264, 272, 274, 281–2, 285 frequency, 273 gravity, 272 group celerity, 273 height, 273–4, 596, 846–7 impact, 407 incident, 249 monochromatic, 264–5, 273–4, 280 number, 274–7, 280 period, 250, 272–7, 280 propagation, 272–4 ripple, 305 significant, 249, 273–4 solitary, 243 swell, 239, 272 Wave–current action, 305 Wave-cut platform, 407 Weed: invasive, 747 propagule, 752 Wet grassland, 214 Wetland:, brackish, 38, 529, 748 coastal, 1–155, 185–6, 211–29, 600–1, 606, 848 creation, 716 fauna, 124, 136–7, 146 flora, 4, 523, 797–8 freshwater, 13, 45, 81, 515–33, 578–9, 605 nontidal, 808, 823–4 ombrogenous, 70 perimarine, 90–1, 93, 108–9 polar coastal, 119–55 prairie, 888–9 restoration, 44, 54, 738, 805, 899 restored, 738, 805, 815–16, 825 subarctic, 99 temperate, 225, 605 tropical coastal, 29, 45, 63–88
925
Subject Index
Wetting, 29–30, 252–3, 265, 409, 463, 892 Wild-fowling, 404 Wind: dehydration, 322 speed, 215 Wintering area, 140–1, 145–6
Woody peat, 69 Wrack deposition, 13, 95 Yellow fever, 36–8 Zonation pattern, 68–9, 255, 328–31, 519–21, 622–3, 657, 673–4
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GEOGRAPHIC INDEX
Abbotts Hall, 776, 778 Adelaide, 695 Adriatic Sea, 405, 412 Africa, 67, 83, 346, 405, 427, 507, 594, 623, 625, 747, 857 Aira-Gawa, 238 Alaska, 3, 42–3, 121–2, 127, 131, 137–9, 142–3, 145, 147, 223–4, 518, 529, 741, 858 Peninsula, 145 Albany, 695 Albemarle-Pamlico Sound, 105 Alert, 138–9 Aleutian Islands, 121 Anse d’Aiguillion, 206 Antarctica, 695 circle, 121 Arctic Coastal Plain, 121–2, 131, 137, 144 Circle, 121, 124 Ocean, 121, 123–5, 134, 138, 223, 858 Argentina, 21, 107, 110, 186, 193, 200, 206, 215, 518, 858 Asia, 44, 69–70, 73, 76–8, 80, 83, 104–105, 109, 126, 248, 346, 427, 518, 623, 626, 656, 696, 787, 791 Asia Pacific region, 248 Australasia, 105, 625, 689, 876 Australia, 9, 11, 18, 39, 44, 53, 57, 66, 68–70, 72, 76, 78–9, 81–3, 105–6, 109, 145, 186, 205, 221, 236, 238, 245–6, 251–2, 256, 377, 379, 387, 405, 429, 568–9, 571, 575–6, 578, 582–4, 594, 596, 600, 603, 620, 624, 628, 655, 657, 665–7, 669, 670, 672, 688, 693–6, 701–2, 744–5, 748, 755, 856, 858–60, 862–4, 866–7, 870–3, 875 Awase Tidal Flat, 698
Baeksu Tidal Flat, 301 Baffin Island, 124, 139 Bahamas, 67, 526, 571 Bahı´a Blanca, 21, 95, 107–108, 110, 162, 189, 192–3, 200, 202–203, 206, 215 Baja California, 103, 857, 875 Bangkok, 82 Bangladesh, 40, 619, 626 Barataria Basin, 540, 843–4 Barbuda, 571 Barrow, 223 Bashita-Minato, 251, 256 Beaufort, 130, 700 Beaulieu, 427, 692 Bele´m, 598 Belgium, 518, 522, 529, 549, 717, 766, 842 Belize, 68, 72, 76, 605, 627, 659–62, 668, 670, 673–4, 658, 865 Bermuda, 332, 570, 573, 577–9, 857 Bolinao, 378, 697 Borneo, 583 Bowen, 133, 217, 224, 576 Brazil, 107, 598, 600, 608, 624, 626, 857–8, 871 Brisbane, 576 Britain, 97, 407, 414, 417, 496, 751, 844 British Columbia, 518, 802–804, 806, 846, 859 Brittany, 722, 753 Brooks Range, 121–2 Brunei-Daressalam, 583 Burdekin, 567 Bylot Island, 139, 145 Cairns, 569, 575 Caleta Brightman, 197 California, 31, 103, 447, 529, 721, 724, 727, 729, 747, 749, 755–6, 802–804, 857, 875 927
928
Cambodia, 697 Canada, 3, 11, 21, 43, 98–9, 109, 120–1, 126–7, 131–3, 138, 142–3, 186, 213, 219, 296, 301, 322, 460, 802–803, 807, 822–3, 825–6, 858 Canadian Arctic Islands, 121, 125 Can Gio, 235, 238, 241–2 Canale S. Felice, 272 Cape Churchill Peninsula, 136 Cape Cod, 217 Cape Harlett, 133 Cararma Inlet, 865 Carr’s Island, 540 Cayman Islands, 74, 577, 579 Cedar Keys, 861, 868 Chile, 107, 138, 333, 741, 858 China, 28, 104–106, 109, 296, 298, 300–302, 304–305, 310, 312, 405, 425, 619, 620, 624, 667, 670, 695, 697–8, 721, 748 Chioggia Inlet, 265, 274–5, 278 Christchurch Harbour, 303 Christmas Island, 579 Chukotka Peninsula, 139, 142 Churchill, 136, 139, 144 Cockburn Sound, 694–5 Cocoa Creek, 236, 251 Colne Pt. 460 Colombia, 575–6, 584, 626 Coral Creek, 238, 245, 572 Corner Inlet, 66, 858, 861, 867 Coromandel Penninsula, 861 Costa Rica, 626, 794, 797 County Down, 498 Cuba, 576, 619 Currambene Creek, 865 Cyprus, 387, 390 Darwin Harbour, 9, 572 Denmark, 97, 109, 382, 385, 387, 526, 691–2 Dominican Republic, 620 Dueling Creek Marsh, 819 Dungeness, 110 East Bintan, 697 Ecuador, 238, 241–2, 626, 660, 858 El Salvador, 567
Geographic Index
Ellesmere Island, 138–9 England, 94, 99–101, 214, 296, 299, 301, 303, 404, 413, 416, 460, 495, 497, 501, 502–504, 510, 518, 528, 740, 741, 765, 771, 775, 840 English Channel, 416 Essex, 95, 495, 497, 501, 504–505, 510, 722 Eurasia, 142–3, 745 Eurasian Arctic, 142 Europe, 31, 50–1, 53, 81, 92, 97–8, 109, 121, 144, 186, 196, 338, 346, 404–405, 416, 427–9, 456, 460, 515, 518–22, 524, 528–9, 687–94, 702, 716–17, 722, 724, 726, 738, 742, 750–1, 754, 763, 765–8, 773, 775–6, 778, 780–1, 841 Eve Street Wetland, 741, 743 Everglades, 67, 70, 74, 82, 222, 603, 651, 665–6, 865, 893, 899 Fanga’uta Lagoon, 580–1 Federated States of Micronesia, 576, 651 Fenland, 109 Fennoscandinavia, 122, 124, 134 Finland, 122 Florida, 16, 42, 68, 73–5, 82, 98, 101, 213, 222, 387, 393, 518, 570, 579, 583–4, 595–6, 601, 603, 605, 623, 630, 647, 650–1, 660–1, 664–6, 673, 699–701, 753, 787, 793–4, 796–7, 857, 859, 861, 865, 868, 870, 875, 893, 896, 899–900 Coastal Everglades, 651, 665–6, 675 Keys, 579 Lagoon, 651 Folaha, 581 Fort Matanzas, 857 Foxe Basin, 120–1, 124, 131–3, 139, 142 France, 11, 98, 109, 385, 387, 429, 522, 630, 692–3, 722, 726, 753, 765 Freiston Shore, 764, 770, 774, 776–7 French Channel, 415 Friesian Islands, 97 Fury Strait, 120, 124 Gambia, 619 Ganges-Brahmaputra-Meghna system, 70
929
Geographic Index
Georgia, 102, 357, 363–4, 447, 453, 465, 552, 699 Marsh, 450 Germany, 102, 357, 363–4, 447, 450, 453–4, 465, 552, 699 Gordon Creek, 236, 251 Great Bahama Bank, 70 Great Barrier Reef, 70, 72, 568, 573, 575 Great Lakes, 120 Great Sippewissett Marsh, 455–6, 466 Greenland, 121, 123, 138–9, 142, 379, 836 Grijalva, 83, 567 Grouper Gardens, 68 Guadeloupe, 620 Guinea-Bissau, 576 Gulf Coast, 98, 101–102, 425, 429, 446, 460, 465, 518, 538 Gulf of Bothnia, 122 Gulf of Maine, 99–100 Gulf of Normandy, 97 Gulf of Papua, 568 Gulf St. Vincent, 867 Guyana, 858 Hamilton, 464 Harbour, 332 Township, 804 Hawaii, 620 Hecla Strait, 120, 124 Hepu, 697 Himalaya, 71, 122, 595, 598 Hinchinbrook, 39, 238, 245–6, 252, 572–3, 603, 635, 669–70, 874 Channel, 39, 252, 670, 874 Island, 238, 245–6, 572–3, 603, 669 Hiroshima, 698 Ho-Chi-Minh, 238 Hokkaido Island, 518 Hollywood, 796 Hong Kong, 29, 34, 42, 626, 629, 697 Hudson Strait, 120, 124 Iberian Peninsula, 97 Iceland, 125, 137, 142–3 Igloolik Island, 131 Inagua, 571
India, 584, 598, 619, 625, 630 Indonesia, 68–9, 71, 73, 78, 575, 619–20, 626, 697, 790 Ipoh, 76 Ireland, 301, 498, 751 Iriomote Island, 19, 20, 236, 238, 251, 256, 518, 666 Ishigaki Harbor, 241–2 Isle of Skye, 496 Italy, 186, 332, 387, 390, 412, 691 Jamaica, 74, 387, 667, 670, 865 Jansand, 349 Japan, 238, 241, 247, 251, 256, 405, 427, 509, 518, 688, 695–8 Java, 337, 584 Jiangsu, 105–106, 296, 298, 300–303, 305–307, 310–12 Jug Bay Marsh, 539 Kamchatka Peninsula, 121 Kampo, 697 Karrendorfer Wiesen, 773 Kenilworth Marsh, 529, 818–19, 823 Kenya, 623, 665, 795–6 King Creek, 68 King Sound, 9, 862 Kingman Marsh, 818–19, 822–3 Klong Ngao, 788 Kola Peninsula, 124 Korsae, 651 Koukdjuak, 120, 133 Kuwait, 216 La Camargue, 104 Lake Argyle, 70 Lake MacLeod, 571 Lido Inlet, 265, 274–5, 277–8 Limfjord, 691–2 Little Sound, 580 Loch Beag, 496 Long Hoa, 235, 242 Long Island Sound, 100 Louisiana, 432, 511, 659, 802–804, 836, 843–6, 859–62, 868–9, 875, 898–900
930
Low Isles, 573 Lower Ob Basin, 147 Maera-Gawa, 238 Mai Po, 29, 34, 42, 629 Maine, 98–100 Malamocco Inlet, 265, 274–5 Malaysia, 69, 72, 78, 618–19, 624–5, 667, 670, 699, 787, 792, 795 Manabi, 238 Mandora, 571 Manukau Harbour, 698 Maranha˜o, 598 Martigues, 692 Maryland, 453, 517, 520, 522, 525–6, 551, 716, 818 Mashapaquit, 460 Massachusetts, 100, 447, 455, 508, 528, 548, 753 Matang Forest, 660, 792 Mexico, 575, 603, 651, 670–2, 699, 803, 857–9, 861, 867–8, 870–1, 873, 875 Micronesia, 33, 576, 619, 651, 655 Minas Basin, 109, 201, 219 Mississippi Sound, 102, 699, 894 Mud Island, 575, 585 Nakama-Gawa, 238 Natal, 583, 584 Nauset Marsh, 457 Nee Soon, 80 Negril Swamp, 866 Nelson Lagoon, 139, 145 The Netherlands, 28, 109, 186, 301, 387, 448, 518–19, 522, 529, 690, 702, 717, 765–6 New Caledonia, 576 New England, 99, 100–101, 460, 465, 518, 528 New Guinea, 16, 71, 568–9, 571, 577, 581–2, 620 New Jersey, 100–101, 523, 804, 818 New South Wales, 81, 567, 693, 860, 863, 874 Newtownards, 301 Nigeria, 790
Geographic Index
Noard-Fryslaˆn Buˆtendyks, 770 Norfolk, 94, 404, 841 North America, 3, 21, 31, 51, 55, 73, 81, 92, 98–9, 103, 107–108, 122–4, 143–4, 147, 196, 222, 425, 427, 429–30, 449, 515, 518–20, 524–6, 529, 538, 545, 605, 607, 699, 738, 745, 750, 751, 773, 776, 805, 858–9, 873 North Atlantic Drift, 124 North Carolina, 101, 508, 700, 719, 727, 729, 732, 802, 804 North Fambridge, 510, 701 North Inlet, 170, 174, 448, 457, 469 North Norfolk, 94, 841 North Point, 120–1, 129 North Sound, 580 Northeast Passage, 147 Northern Ireland, 301, 498, 751 Northern Marshalls, 571 Northern Territory, 9, 68, 569, 578 Northey Island, 765, 778 Northwest Passage, 147, 673 Norway, 122, 125, 143 Nunuvut, 131 Oceania, 695–6 Okinawa, 242, 666, 698 Oregon, 802 Orplands, 779 Otter Point Creek, 539 Oude Maas, 519 Pacific: Islands, 651, 696, 698 Ocean, 67, 70, 123, 137, 145, 581, 643, 859 Pantai Remis, 76 Papua New Guinea, 16, 71, 568, 620 Para´, 598 Patagonia, 107, 200, 204 Patagonian Coast, 107 Paull Holme, 768, 778–9 Pauatahanui Inlet, 219 Paulina Marsh, 447, 468 Pechora Basin, 147 Pennsylvania, 806, 818
Geographic Index
Persian Gulf, 405 Peru, 576, 857–8 Philippines, 378, 696–7, 787, 795–7 Philips Creek, 457, 460 Phonpei, 651 Pipon, 573 Point Samson, 576 Poland, 691 Pope’s Nose Creek, 584 Portugal, 690, 692, 772 Puck Lagoon, 691 Puerto Rico, 583–4, 619, 667, 670 Qatar, 795 Queen Maud Gulf, 139, 142 Queensland, 40, 79, 252, 569, 572, 575, 693, 862–3, 866 Quintana Roo, 866 Rasmussen Lowlands, 139, 142 Ringfield, 457, 460 Romney Marsh, 109–110, 415 Russia, 3, 121–5, 131–2, 134, 137, 139, 142, 146, 518 St Lucia Lake, 583 San Tau Beach, 697 Sa˜o Luis, 598 Sapelo Island, 451 Saudi Arabia, 576 Scotland, 496 Siberia, 122, 143, 147, 529 Sieperda Marsh, 774, 779 Singapore, 8, 576, 697 Sippewissett, 451, 455, 457, 466, 471 Skagerrak, 334 Skidaway Is., 451 Solomon Islands, 576 Somerset Levels, 109 South America, 67, 70–71, 76, 104, 107, 110, 140–1, 346, 427, 518, 594, 750, 857 South Carolina, 102, 170, 448, 469, 504, 509, 528, 552, 602, 699, 845 South Korea, 301 Southampton Island, 139, 146 Spain, 387, 390, 405, 693, 727, 740
931
Sri Lanka, 620, 790 Stiffkey, 404 Strait of Magellan, 138 Suez Canal, 689 Sundarban, 40, 70, 595 Suriname, 600 Sweden, 122 Sweet Hall Marsh, 536–9, 541, 547–8, 550 Sydney, 567, 743–4, 864, 871 Harbor, 567 Taimyr, 137 Taylor Slough, 672–4 Terminos Lagoon, 651, 653, 659–61, 670–2 Texas, 98, 447, 675, 699, 727, 859, 861–2 Thai Thuy, 247 Thailand, 16, 19, 29, 40, 387, 393, 575, 598, 619, 626, 660, 665, 667, 670, 787–8, 790, 795 Three Gorges Dam, 109 Three Isles, 573 Thuy Hai, 249 Tibet, 595 Tierra del Fuego, 107, 138, 141 Timika, 571 Tjilatjap, 584 Tobacco Range, 865 Tollesbury, 504, 505, 722, 768–72, 774–5, 778–9 Tonga, 577, 580 Tongatapu, 581 Torres Strait, 568 Torridge Marsh, 460 Toulon, 690 Townsville, 236, 577 Twin Cays, 68, 668, 670, 674, 865 United Kingdom, 81, 95, 108–110, 692, 712, 765, 767, 770, 773 USA, 16, 25, 27, 31, 33, 39, 42, 51, 100, 170, 174, 213, 273, 303, 357, 359, 363, 385, 447–8, 456–7, 465, 476, 503, 509, 517, 520–3, 525–6, 528–9, 536–7, 541–3, 547–8, 551–2, 554, 595, 605, 619, 665–6, 753, 793, 795–6, 817, 857, 860, 868–70
932
Geographic Index
Venezuela, 598, 600 Venice Lagoon, 162, 165–6, 170, 173–4, 265, 269, 270, 274, 278–9, 282–3, 332, 412, 460, 692–3, 845–6 Victoria, 66, 105, 693, 860, 866 Vietnam, 12, 235, 238, 241–2, 247, 249, 608, 618–20, 667, 670, 787, 790 Virginia, 33, 39, 51, 247, 447, 525, 536, 549, 552, 603, 701, 727, 729, 813–14, 818
West Indies, 44, 66–7, 70, 73 West Papua, 571, 573, 581 Western Australia, 9, 69, 76, 78, 379, 571, 693–5, 701, 744 Western Scheldt, 772 Western Siberian Plain, 147 Westham Island, 846, 847 Whangarei Harbour, 698 Wrangel Island, 139, 142 Wynnum, 655, 657
Wales, 81, 324, 567, 693, 860, 863, 874 Wallasea, 767 The Wash, 296, 301, 322, 413 Washington, 56, 123, 528–9, 721, 802, 817–18, 820, 826 Wentlooge Flats, 324
Yakutat Forelands, 139, 145 Yamal, 123 Yenissei, 123 Yukon, 121, 139, 143, 145 Zackenberg, 138–9
TAXONOMIC INDEX
Acanthiza iredalei rosinae, 430 Acanthus ilicifolius, 633 Acer rubrum L., 522 Acorus calamos, 521, 823 Acrostichum aurem, 67 Aegicerus corniculatum, 620, 633, 667, 670 Aeluropus littoralis, 69, 105–6, 568, 633 Aequipecten irradians concentricus, 699 Aeschynomene virginica (L.) BSP, 522 Agalinis maritima, 100 Agelaius phoeniceus L., 525 Alauda arvensis, 799 Algae, 2, 19, 38, 52, 93, 98, 255, 320–1, 352, 356, 358–9, 379, 390, 428, 436–7, 523, 537, 552, 630–1, 727, 773, 808 Alle alle (little auks), 146 Allenrolfea patagonica, 107 Alligator, 76, 80, 569, 578, 862, 867, 899 Alopex lagopus, 146 Alosa alosa L., 524 Alosa fallax Lace´pe`de, 524 Alosa sapidissima Wilson, 524 Ambrosia trifida L., 520 Ammodramus caudacutus, 430 Ammodramus maritimus, 430 Ammophila arenaria, 136 Amphibolis spp., 695 Amphibolis antarctica, 695 Amphipod, 138, 329, 333, 434, 779 Amphipods, talitrid, 434 Anas cracca (teal), 779 Anas platyrhynochos, 779 Anas rubripes Brewster, 526 Anquillids, 428 Anser brachyrhynchus (pink-footed geese), 143 Anseriformes (waterfowl), 430 Anthus pratensis (meadow pipit), 779 Apium sellowianum, 107
Apodasmia similis (jointed rush), 861 Arctosa fulvolineata, 777 Ardea herodias L, 523 Arenaria interpres, 138 Arenicola marina, 320, 330 Arrow-arum, 520 Arrowgrass, seaside, 431 Artemisia galla, 104 Arthrocnemum macrostachyum, 104 Arthrocnemum perenne, 427 Arthrocnemum subterminale, 427 Aster spp., 95, 98, 506, 724, 773 Aster tripolium, 98, 773 Atherinosoma microstoma, 429 Atriplex patula, 99, 136 Atriplex portulacoides, 98, 431, 723, 726 Audouinella sp. (Dillwyn), 319, 320 Auk, 146 Avicennia africana, 619 Avicennia alba, 667, 670 Avicennia germinans (black mangrove), 67, 101, 599, 619, 650, 665–8, 670, 857, 860, 869, 896 Avicennia marina, 66, 105, 568, 629, 665–8, 670, 795, 858, 861 Avicennia marina var. Australasica, 66, 858 Avicennia marina var. Eucalyptifolia, 568, 858 Avicennia marina var. Marina, 858 Avicennia officinalis, 796 Baccharis halimifolia, 729 Bacteria, 320, 329, 351–2, 355–6, 382, 384–6, 389, 392, 434, 436–7, 450, 454, 464, 466, 468, 472, 475–6, 538, 545, 552, 555, 625, 627–8, 630, 633, 653–4, 657, 716, 740 Baldcypress (Taxodium distichum), 896 Barnacle, 138, 144, 145 Bartsia alpina, 136 Bass (Dicentrarchus labrax), 776 933
934
Batis maritima, 67, 101, 859 Bear, brown, 146 Bear, grizzly, 146 Bear, polar, 43, 146–8 Beaver, 523–5 Benthic invertebrates, 54, 334, 523, 630, 724, 726–7, 778–9, 812 Bidens laevis (L.) BSP, 520 Birds, long-legged wading, 430 Birds, wading, 430, 724, 726, 729, 741, 779, 899 Bittern, least, 525 Bivalve, 20, 137–8, 329, 331, 338, 350, 436, 437, 779 Blackbird, red-winged, 525–6 Blue crab, 31, 431, 434–5, 727 Blysmus rufus, 773 Bobolink, 526 Bolboschoenus spp., 507, 520 Bolboschoenus maritimus (L.) Palla, 520 Borrichia frutescens, 859, 861, 870 Brant, Atlantic, 143 Branta bernicla (Atlantic brant), 143 Branta canadensis L., 525 Branta leucopsis (barnacle goose), 144 Bruguiera conjugata, 633 Bruguiera parviflora, 568 Bruguiera sexangula, 619, 633 Bull tongue, 101 Bulrush, 431 Bunting, reed (Emberiza schoeniclus), 779 Burguiera gymnorrhiza, 568, 571, 583–4, 619, 667, 670 Burmarigold, 520 Calidris alpina (dunlin), 42, 142, 145, 779 Calidris canutus, 138 Calidris mauri (western sandpipers), 145 Calidris ptilocnemis (rock sandpipers), 145 Calidris pusilla, 145 Callinectes sapidus, 131, 727 Callitriche heterophylla Pursh, 525 Caltha palustris L. (var. araneosa), 522 Carex spp., 105, 134–5, 427, 520, 522, 773, 859 Carex glareosa, 427
Taxonomic Index
Carex maritima, 135 Carex ramenskii, 134 Carex salina, 134 Carex scabriflora, 105 Carex serotina, 773 Carex subspathacea, 134 Carex ursina, 135 Caribou, 43, 146–8 Castor canadensis Kuhl, 523 Cattail, 520, 811, 816 Cattle, 19, 430–1 Caulerpa taxifolia, 26, 31, 336 Centrarchidae, 524 Ceriops spp., 667 Ceriops decandra, 568, 619, 667, 670 Ceriops tagal, 584, 665, 667, 863 Charadriiformes (shorebirds, gulls, and terns), 430 Chasmagnathus granulatus, 194, 330 Chemoautotrophy, 50, 351–2, 448, 453–4, 462, 465 Chen caerulescens atlantica, 145 Chen caerulescens caerulescens, 142 Chen canagica, 142 Chen rossii, 142 Chironomidae, 138 Chrysanthemum articum, 136 Ciconiiformes (long-legged wading birds), 430 Cinna arundinacea L., 522 Cistothorus palustris Wilson, 525 Cladocerans, 138 Clupea harengus (herring), 776–7 Clupeids, 428 Cochlearia officinalis, 135 Conocarpus erectus, 583, 857, 866 Copepods, 138, 329, 631 Coral reefs, 2, 32–35, 52, 70, 318, 378, 393, 570, 631, 634, 658 Cordgrass, 429–34, 520, 860 Cordgrass, big, 520 Cordgrass, salt-meadow, 433 Cordgrass, smooth, 429, 431–4, 860 Cordylanthus maritimus ssp maritimus, 749 Corophium arenarium, 328
935
Taxonomic Index
Corophium salmonis, 138 Corophium volutator, 219, 328, 330, 778–9 Cortaderia selloana, 107 Cotula coronopifolia, 107 Crab, 19, 31, 34, 42, 46, 162, 194–5, 200, 202–3, 247, 330, 431, 434, 598, 602, 626, 629–30, 663, 753, 795 Crab, blue, 31, 431, 434–5, 727 Cressa truxiliensis, 107 Crustacean, 49, 138, 330, 430, 433, 616, 625, 631, 727, 778 Cyanobacteria, 49, 337, 352–5, 359, 428, 434, 456 Cymodocea nodosa, 381, 387, 689 Cymodocea rotundata, 387 Cymodocea serrulata, 387 Cyperaceae, 427, 599–600, 744 Cyprinidae, 524 Cyprinus carpio L., 524 Cytophaga-Flavobacterium (CFB) consortium, 351 Deschampsia wibeliana (Sond.) Parl, 522 Diatoms, 22, 24, 96, 204–5, 329, 362–4, 428, 625, 754 Dicentrarchus labrax (bass), 777 Dipteran, 138, 727 Distichlis scoparia, 427 Distichlis spicata, 99, 102, 427, 446, 749, 859 Dolichonyx oryzivorus L., 526 Duck, 140, 146, 526–7, 804 Duck, American black, 526 Dunlin (Calidris alpina), 142, 779 Earthworm, 139 Ectoprocta, 138 Eelgrass (Zostera marina), 53, 137, 385, 387, 688 Eleocharis geniculata, 599 Elymus spp., 98–99 Elymus arenarius, 99 Elytrigia atherica, 751 Emberiza schoeniclus (reed bunting), 779 Enhalus spp., 377–8, 387, 696 Enhalus acoroides, 377–8, 387 Enoplognatha mordax, 377
Epilobium, 529 Epiphyte, 379, 381, 383, 434, 475, 626–7 Erigone longipalpis, 777 Eteone longa, 138 Eudocimus ruber, 626 Excoecaria spp., 69, 582, 633 Excoecaria agallocha, 633 Fen, 90, 96, 111, 128, 130 Festuca rubra, 98, 135, 427, 751 Fimbristylis spadicea, 599 Finfish, 631, 724, 726–7, 729 Fish, 20, 21, 25, 32, 37–38, 42, 49, 56, 103, 137, 200, 225, 251, 320, 326, 329, 336, 378, 386, 389–90, 429, 433, 436, 516, 523–4, 555, 616, 625, 630–2, 696, 701, 716–7, 721, 727, 740, 776–7, 792, 794–7, 805, 807–8, 810, 812, 815, 817, 825 Flies, 138 Foraminifera, 96, 138 Fox, Arctic, 146 Fox, red, 146 Frankenia pulverulenta, 104 Frankenia salina, 428 Fraxinus pennsylvanica Marsh, 522 Fundulus spp., 432, 434, 524, 727 Fundulus diaphanous Lesueur, 524 Fundulus heteroclitus L., 524 Fungi, 434, 436–7, 450, 523, 538, 625, 627–8, 630, 633, 716 Gastropod, 138, 330, 333, 435–7, 631 Geese, 21, 134, 136, 142–6, 431, 525, 692, 745, 822–6 Geothlypis trichas L., 525 Glasswort, 428 Glaux maritima, 99 Gobiopterus semivestitus, 429 Goby, glass, 429 Godwit, bar-tailed (Limosa lapponica), 145 Goose, barnacle, 144–5 Goose, Canada, 525 Grackle, common, 526 Graminoid, 134, 136, 144–5
936
Grass, 10, 21, 44, 53, 67, 69, 74, 75, 77–78, 98, 101, 134, 138, 379, 431, 433, 688, 690, 696, 727, 749 Gratiola neglecta Torrey, 520 Gull, 42, 140, 146, 430, 526 Halimione spp., 98, 429, 723 Halimione portulacoides, 98 Halodule beaudetti, 387 Halodule wrightii (shoal grass), 53, 394, 688, 700 Halodule, 53, 387, 394, 688, 695–6, 700 Halophila spp., 377, 380, 689 Halophila ovalis, 378, 379, 387 Halophila stipulacea, 689 Hardyhead, small-mouth, 429 Hare, 146, 431, 750 Hare, Arctic, 146 Hediste diversicolor, 330, 336, 779 Heleidae, 138 Hemigrapsus spp., 727 Heritiera littoralis, 69, 568, 633 Heron, great blue, 523 Herring (Clupea harengus), 523 Heterostachys ritteriana, 103 Heterozostera tasmanica, 695 Hibiscus moscheutos, 520 Hierochloe odorata, 99 Hippuris tetraphylla, 136 Hippuris vulgaris, 136 Hydrilla verticillata, 823 Hydrobia minuta, 138 Hydrobia ulvae, 330 Ictaluridae, 524 Impatiens capensis Meerb, 520 Imperata cylindrical, 105–6 Insect, 24, 138, 147, 433, 538, 625–6, 627, 631, 727 Isopod, 138, 333, 627, 749 Iva frutescens, 99, 729 Ixobrychus exilis Gmelin, 525 Jaumea carnosa, 103 Jewelweed, 520
Taxonomic Index
Juncus spp., 89, 98, 101–2, 104–5, 107, 427, 446, 520, 722, 724, 743, 823, 859, 864 Juncus acutus, 107, 743 Juncus articulatus, 107 Juncus balticus, 89, 859 Juncus effusus, 823 Juncus gerardii, 98 Juncus kraussii, 105, 427, 864 Juncus maritimus, 105 Juncus roemerianus, 101–2, 427, 446, 724 Juncus setchuensis, 105 Kandelia, 12, 239, 249, 618–20, 667, 670 Kandelia candel, 249, 618–20, 667, 670 Killifish, 49, 432, 434, 524, 727 Kite, snail, 899 Knot, red, 138, 141–3, 145 Koenigia islandica, 136 Lagodon rhomboides, 727 Laguncularia racemosa (white mangrove), 583–5, 619, 650 Lampranthus tegens, 747 Lanice spp., 333 Lapwing (Vanellus vanellus), 780 Lark, sky (Alauda arvensis), 779 Larus argentatus, 526 Larus atricilla, 526 Larus delawarensis, 526 Lavender, sea, 428 Leersia oryzioides (L) Swartz, 521 Leishmania donovani, 633 Lemming, 146, 148 Lemna minor, 823 Leptocarpus similis (jointed rush), 861 Leymus mollis var. arenarius, 136 Limonium spp., 99, 101, 104, 107, 427–8, 506 Limonium brasiliense, 107 Limonium californicum, 428 Limonium girardianum, 104 Limonium nashii, 99 Limonium virgatum, 104 Limonium vulgare, 427, 506 Limosa haemastica, 145
937
Taxonomic Index
Limosa lapponica (bar-tailed godwit), 145 Limpet, 138 Lindernia dubia (L.) Pennell, 520 Littoraria irrorata, 431, 434 Loon, 140 Ludwigia palustris (L.) Ell., 520, 823 Ludwigia peploides, 823 Lumnitzera spp., 69, 236, 582, 584 Lumnitzera racemosa, 584 Lutra canadensis Schreber, 523 Lythrum spp., 521, 819, 823–4 Lythrum salicaria, 819, 823–4 Macoma baltica, 779 Macroalgae, 24, 25, 50, 351, 371, 378, 381–2, 434, 446, 504, 692 Macrobenthos, 138, 140, 329–50 Macroinvertebrate, 432–4, 630, 824 Macrophyte, 31, 90, 170, 279, 445–7, 456, 463, 475–6, 536, 691, 724 Maidencane, 101 Malaclemys terrapin, 430 Mallard (Anas platyrhynochos), 779 Manatee grass (Syringodium filiform), 53, 688 Mangrove, black, 67, 101, 860 Mangrove, red (Rhizophora mangle), 67, 583, 599, 619, 624, 650, 665–8, 670, 857, 896 Mangrove trees, 9, 34, 47, 232–3, 235–6, 240–1, 249–50, 254–6, 571, 616, 626, 628, 671, 896 Marenzelleria cf. viridis, 336 Melaleuca cajuputi, 69 Melampus bidentatus, 434 Microalgae, 24, 50, 351, 354–6, 358–9, 361–3, 434, 446, 448, 536, 547, 630–2, 740 Microbe, 20, 329, 350, 352, 450, 453, 466, 468, 538, 542, 549, 555, 615, 625, 627–8, 630, 720 Microbial mats, 337, 352–3, 359 Microheterobenthos, 329 Microphytobenthos, 163, 166, 179, 264, 326, 329, 337 Mink, 146 Mites, 139
Mora oleifera, 67 Morella (Myrica) cerifera, 520 Morone saxitalis Walbaum, 520 Mosquito, 20 Mosquito, saltmarsh (Ochlerotatus taeniorrhynchus, Ochlerotatus sollicita), 36 Mouse, salt marsh harvest, 430 Murres, thick-billed (Uria lomvia), 146 Muskox (Ovibos moschatus), 146 Muskrat (Ondatra zibethicus), 523–5, 814 Mussel, 49, 138, 395, 434 Mussel, blue, 395 Mycorrhizae, 536 Myrica (Morella) cerifera, 520 Myriophyllum spicatum, 520 Mytilus edulis, 395 Naididae, 138 Nekton, 428–9, 432–3, 729 Nematoda, 778 Nematode, 138–9, 329, 625, 639 Neohelice (formerly Chasgmanathus) granulatus, 194, 202, 431 Nereis diversicolor, 779 Nerodia clarkia, 429 Nuphar lutea (L.) Sm., 519–21, 525, 823 Nyctanassa violacea, 626 Nypa fruticans, 69, 568, 620 Nyssa sylvatica, 520 Ochlerotatus camptorhynchus, 749 Ochlerotatus sollicita (saltmarsh mosquito), 36 Ochlerotatus taeniorrhynchus (saltmarsh mosquito), 67 Oenanthe conioides Lange, 522 Oenanthe lachenalii, 733 Oligochaeta, 778 Oligochaetes, 138, 824–5 Ondatra zibethicus L. (muskrat), 523 Ononis repens, 773 Orontium aquaticum, 823 Oryzomys palustris Harlan, 524 Oscillatoria spp., 353 Osier, 519 Osmunda cinnamomea L., 522
938
Osmunda regalis var. spectabilis (Willd.) Gray, 522 Osmunda regalis, 522 Osprey, 523 Ostracod, 138 Otter, river, 523 Ovibos moschatus (muskox), 146 Owl, 140 Oyster, 30–31, 42, 49 Palaemonetes pugio, 727 Pandion haliaetus L., 523 Panicum hemitomon, 101 Panther, Florida, 899 Paranais spp., 138, 778 Paranais litoralis, 778 Parapholis incurva, 749 Pardosa purbeckensis (halophilic wolf spider), 777–8 Parnassus palustris, 136 Passerculus sandwichensis (Belding’s savannah sparrow), 740 Passerine, 140, 148, 779 Pelliciera spp., 67 Peltandra virginica (L.) Schott (arrow arum), 520–1, 543, 823 Penaeus aztecus, 727 Penaeus japonicus, 796 Penaeus setiferus, 727 Perca flavescens Mitchill, 523–4 Perch, yellow, 471, 523–4, 633, 729 Periwinkle, 431, 434–5 Phalaris spp., 521, 823–4 Phalaris arundinacea, 823–4 Phalarope, red (Phalaropus fulicarius), 142 Phalaropus fulicarius (red phalarope), 142 Phoca hispida (ringed seals), 146 Phoca vitulina (harbour seals), 147 Phragmites australis (Cav.) Trin. ex Steudel, 521 Phragmites communis, 31, 497 Phyllospadix spp., 699 Phyllospadix torryi, 378 Pickerelweed, 520 Picoeukaryote, 337 Picoheterobenthos, 329
Taxonomic Index
Picophytobenthos, 329, 337 Picophytoplankton, 337 Pilea pumila (L.) A. Gray, 520 Pipit, meadow (Anthus pratensis), 779 Pirata piraticus, 777–8 Plagianthus divaricatus, 107 Plankton, 36, 37, 49, 434, 436, 502, 537 Plantago eripoda, 99, 135 Plantago maritima, 99 Plover, golden (Pluvialis apricaria), 780 Pluvialis apricaria (golden plover), 780 Poa subcaerulea, 773 Poaceae, 427, 600 Polychaete, 138, 329, 332, 335–6, 630, 727, 779 Polygonum arifolium L., 520–1, 823–4 Polygonum punctatum Elliott, 520 Polygonum sagittatum L., 521 Polygonum vivipara, 136 Polypogon monspeliensis, 749 Pontederia cordata L. (pickerelweed), 520 Posidonia, 379, 387–8, 390, 689, 691–2, 694–5 Posidonia australis, 694–5 Posidonia coriacea, 694 Posidonia oceanica, 379, 387–8, 390, 689, 692 Potamogeton filiformis, 136 Potamogeton pectinatus, 136 Potentilla egedii, 135 Primula stricta, 136 Puccinellia spp., 95, 107, 134, 200, 723–4, 726, 773, 775, 859 Puccinellia biflora, 107 Puccinellia fasciculata, 773 Puccinellia magellanica, 107 Puccinellia maritima, 95, 723–4, 726, 773, 775 Puccinellia phryganoides, 859 Purslane, sea, 429 Quiscalus quiscula L., 526 Rabbit, 431 Rail, king, 525 Rail, Virginia, 525
Taxonomic Index
Rallus elegans Audubon, 525 Rallus limicola Vieillot, 525 Rallus longirostris levipes, 729 Rangifer tarandus, 147 Ranunculus cymbalaria, 135 Raptor, 140 Rat, marsh rice, 524 Redshank (Tringa tetanus), 779 Reef, 2, 31–35, 52, 67–68, 70, 72–73, 198, 250, 256, 318, 333, 378, 380, 393, 474, 568, 570, 573, 575, 579, 580–1, 585, 594, 631, 634, 643, 658, 662, 673, 691 Reithrodontomys raviventris, 430 Rhizophora apiculata, 40, 568, 618, 624, 665, 667, 670, 792 Rhizophora mangle (red mangrove), 67, 583, 599, 611, 624, 650, 665, 667–8, 670, 857, 896 Rhizophora mucronata, 238, 624, 633, 665, 667, 795 Rhizophora stylosa, 568, 670 Rice, wild, 520, 524, 526 Rose-mallow, 520 Rotifers, 139 Ruppia cirrhosa, 691 Rush, jointed (Apodasmia similis), 861 Rush, jointed (Leptocarpus similis), 861 Sagittaria lancifolia, 101 Salicornia spp., 98, 100–3, 136, 220, 427–8, 724, 740, 752, 755, 773–5, 859, 864, 870 Salicornia australis, 864 Salicornia bigelovii, 740, 755 Salicornia borealis, 136 Salicornia dolichostachya, 98 Salicornia europaea, 100, 774 Salicornia fragilis, 98 Salicornia ramosissima, 98 Salicornia virginica, 103, 220, 427, 428, 724, 859 Salix, 519, 823 Salix nigra, 823 Salmonids, 428 Saltbush, 431, 875
939
Samolus repens, 427, 864 Samphire, 69 Sandpiper, calidrid, 142 Sandpiper, rock (Calidris ptilocnemis), 145 Sandpiper, western (Calidris mauri ), 145 Sarcocornia fruticosa, 104 Sarcocornia perennis, 21, 107, 202, 749 Sarcocornia quinqueflora, 105, 427 Schoenoplectus spp., 507, 520–1, 823 Schoenoplectus fluviatilis, 823 Schoenoplectus lacustris (L.) Palla, 520–1 Schoenoplectus tabernaemontani, 823 Schoenoplectus triqueter (L.), 520 Scirpus americanus, 100 Scirpus karuizawensis, 105–6 Scirpus lacustris, 543 Scirpus mariqueter, 105 Scirpus maritimus, 431 Scirpus triquiter, 105 Scylla olivacea, 795 Sea heath, 428 Seagrass, 2, 7, 8, 10–15, 18, 20–21, 24, 26, 30, 32, 33–34, 40, 48, 53, 65, 212–22, 320, 334–6, 377–89, 391–3, 395–6, 446, 537, 570, 580, 631, 658, 687–91, 693, 695–704, 789, 797 Seal, harbor (Phoca vitulina), 147 Seal, ringed (Phoca hispida), 146 Sedum rosea, 136 Selliera radicans, 107 Seston, 382 Sesuvium portulacastrum, 599, 859, 875 Shad, 524 Shad, Allis, 524 Shad, American, 524 Shad, Twait, 524 Shellfish, 616, 631, 699, 726–7, 729 Shoal grass (Halodule wrightii ), 53, 394, 688, 700 Shorebirds, 42, 137–8, 140–3, 145–6, 430 Shrimp, 19, 25, 28–29, 32, 247, 429, 433–4, 436–7, 631–2, 657, 716, 727, 787, 792, 794, 796, 896 Shrimp, mud (Corophium volutator), 219, 328, 330, 778–9
940
Shrubs, 22, 24, 66, 69, 128, 148, 403, 519, 522, 729, 779, 804 Smartweed, 520 Snail, mud, 49, 434 Snail, saltmarsh coffee-bean, 434 Snake, salt marsh, 429–30 Sonneratia spp., 69, 79, 146, 236, 239, 521, 568, 571–2, 574, 576, 582, 619–20, 624, 633, 667–8, 791, 795 Somateria mollisima (common eider), 146 Sonneratia alba, 568, 624, 667–8, 795 Sonneratia apetala, 619 Sonneratia caseolaris, 69, 571 Sonneratia lanceolata, 69, 568, 620 Sparganium spp., 521, 823 Sparganium eurycarpum, 823 Sparrow, saltmarsh sharp-tailed, 430 Sparrow, seaside, 430 Spartina spp., 21, 31, 49, 98–99, 101, 103–4, 170, 427, 433, 446, 449, 469, 520, 718–9, 721–4, 747–8, 751, 775–6, 838, 859 Spartina alterniflora, 21, 49, 98 Spartina anglica, 98, 427, 449, 721, 747, 775–6 Spartina cynosuroides (big cordgrass), 101, 520, 719, 723 Spartina densiflora, 104, 427 Spartina foliosa, 103, 427, 721, 747, 859 Spartina maritima, 98, 427, 747 Spartina patens, 31, 99, 427, 433, 446, 469, 718, 838, 859 Spartina spartinae, 859 Spartina xtownsendii, 98 Spatterdock, 520 Spergularia canadensis, 99 Sphaeroma quoyanum, 749 Spider, 137, 777–8 Spider, halophilic wolf (Pardosa purbeckensis), 777 Sporobolus virginicus, 105, 427, 599, 859, 864 Sprat (Sprattus sprattus), 776–7 Sprattus sprattus (sprat), 776–7 Springtail, 137, 139 Stellaria humifusa, 135
Taxonomic Index
Stercorarius pomarinus (pomarine jaegers), 146 Streblospio shrubsolii, 779 Striped bass, 524 Suaeda spp., 98, 100, 105–7, 427, 773, 775, 859, 864 Suaeda australis, 864 Suaeda glauca, 105–6 Suaeda linearis, 100 Suaeda maritima, 98, 427, 773, 775, 859 Suaeda patagonica, 107 Surfgrass, 699 Swan, 140 Symphytum spp., 521 Syringodium spp., 53, 688, 695–6 Syringodium filiforme (manatee grass), 54, 688 Tapes phillippinarum, 332 Tardigrade, 138 Taxodium distichum (bald cypress), 520 Teal (Anas cracca), 779 Tearthumb, 520, 824 Tecticornia spp., 875 Tern, 140, 430 Terrapin, diamondback, 430 Thalassia spp., 380, 387, 389, 695, 699 Thalassia hemprichii, 387 Thalassia testudinum, 380, 387, 699 Thelypteris palustris Schott, 520 Thornbill, slender-billed, 430 Tipulidae, 138 Triglochin maritima, 99, 103, 427, 431, 859 Triglochin palustris, 135 Tringa tetanus (redshank), 779 Turnstone, ruddy, 138, 142 Typha spp., 100, 520–1, 811, 816, 819, 823–4, 842 Typha latifolia (broad-leaved cattail), 842 Uca spp., 430, 451–2, 629, 727, 753 Uca pugnax, 451, 753 Uca vocans, 629 Uhlorchestia spartinophila, 434 Ulva spp., 330, 726, 779 Uria lomvia (thick-billed murres), 146 Ursus maritimus, 146
941
Taxonomic Index
Valeriana spp., 521 Vallisneria americana, 520 Vanellus vanellus (lapwing), 780 Vascular plant, 24, 95, 130, 147, 199, 212, 447, 504, 627, 632, 739, 740 Vertebrate, 140–7, 625 Viburnum dentatum L., 522 Viola cucullata Aiton, 522 Vole, 146, 750 Warbler, common yellowthroat, 525 Waterbird, 104, 146, 779 Waterfowl, 137, 140, 142–3, 146, 148, 430–1, 526, 555, 699, 726, 729, 750, 807 Weasel, 146 Weed, 109, 744, 747–8, 752
Wolverine, 146 Wolves, 146 Worm, 319 Wren, marsh, 525, 729 Xylocarpus granatum, 525, 729 Xylocarpus mekongensis, 624 Xylocarpus moluccenis, 633 Zizania aquatica L., 520 Zizaniopsis miliacea, 101 Zostera japonica, 697 Zostera marina (eelgrass), 53, 688 Zostera noltii, 385, 387, 689, 692 Zoysia macrostachya, 105 Zoysia sinica, 427
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