Estuarine Nutrient Cycling: The Influence of Primary Producers: The Fate of Nutrients and Biomass

Estuarine Nutrient Cycling: The Influence of Primary Producers

Volume 2

The titles published in this series are listed at the end of the volume

Estuarine Nutrient Cycling: The Influence of Primary Producers The Fate of Nutrients and Biomass

Edited by

Søren Laurentius Nielsen Department of Life and Sciences & Chemistry, Roskilde University, Roskilde, Denmark

Gary T. Banta Department of Life and Sciences & Chemistry, Roskilde University, Roskilde, Denmark and

Morten Foldager Pedersen Department of Life and Sciences & Chemistry, Roskilde University, Roskilde, Denmark

KLUWER ACADEMIC PUBLISHERS DORDRECHT / BOSTON / LONDON

A C.I.P. Catalogue record for this book is available from the Library of Congress.

ISBN 1-4020-2638-2 (HB) ISBN 1-4020-3021-5 (e-book)

Published by Kluwer Academic Publishers, P.O. Box 17, 3300 AA Dordrecht, The Netherlands. Sold and distributed in North, Central and South America by Kluwer Academic Publishers, 101 Philip Drive, Norwell, MA 02061, U.S.A. In all other countries, sold and distributed by Kluwer Academic Publishers, P.O. Box 322, 3300 AH Dordrecht, The Netherlands.

Cover illustration: Seagrass with epiphytic algae by Lars Nejrup

Printed on acid-free paper

All Rights Reserved © 2004 Kluwer Academic Publishers No part of this work may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, microfilming, recording or otherwise, without written permission from the Publisher, with the exception of any material supplied specifically for the purpose of being entered and executed on a computer system, for exclusive use by the purchaser of the work. Printed in the Netherlands.

AQUATIC ECOLOGY SERIES

Aquatic ecology is an extraordinarily broad and diverse discipline. Aquatic ecology is the study of the functional relationships and productivity of organisms and communities of waters as regulated by their physical, chemical, and biotic environment. The marine environment extends broadly from the complex land-water coastal environments to the open ocean. Limnology encompasses all inland aquatic environments, including streams, rivers, lakes, reservoirs, and wetlands. Research has accelerated in certain areas and been less active in others. Reassessments and syntheses are stimulating to the discipline as a whole, as well as enormously useful to students and researchers in ecological sciences. A series of succinct monographs and specialized evaluations in aquatic ecology has been developed. Subjects covered are topical (e.g., lake hydrodynamics, microbial loop in aquatic ecosystems) rather than broad and superficial. The treatments must be comprehensive and state-of-the-art, whether the topic is at the biochemical, mathematical, population, community, or ecosystem level. The objectives are to advance the topics by the development of arguments, with documented support, that generate new insights, concepts, theories to stimulate thought, ideas, directions, controversies. The books are intended for mature as well as emerging scientists to stimulate intellectual leadership in the topics treated. Receipt of manuscripts approximately 18 months after an agreement is desired, for publication within 10 months thereafter. Electronic submission is essential with hardcopy. Format and manuscript guidance will be provided. For further information and book proposal details please contact: Prof. Robert G. Wetzel, Series Editor Department of Environmental Sciences and Engineering The University of North Carolina Chapel Hill, North Carolina 27599-7431 USA Email: [email protected] Phone: 919 + 843-4916

Dr. Anna Besse, Publishing Editor Aquatic and Biogeosciences Springer Van Godewijckstraat 30 P. O. Box 17 3300 AA Dordrecht The Netherlands Email: [email protected] Phone: 31 (0) 651 33 86 01

v

CONTRIBUTORS

CARL L. AMOS Southampton Oceanographic centre, University of Southampton, Southampton SO14 3ZH, Hants, England GARY THOMAS BANTA Department of Life Sciences and Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark ALESSANDRO BERGAMASCO CNR, Ist. Sperimentale Talassograf, Spinata S. Raineri, I-98122 Messina, Sicily, Italy MORGENS R. FLINDT Institute of Biology, University of Southern Denmark, Campusvej 55, DK-5230 Odense M, Denmark JENNIFER HAUXWELL Wisconsin Department of Natural Resources, Department of Natural Resources Research Center, 1350 Femrite Drive, Monona, WI 53761, USA JOAO NETO Institute of Marine Research, Department of Zoology, University of Coimbra, P-3004517 Coimbra, Portugal SOREN LAURENTIUS NIELSEN Department of Life Sciences and Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark vii

viii

MIGUEL A. PARDAL Institute of Marine Research, Department of Zoology, University of Coimbra, P-3004517 Coimbra, Portugal MORTEN FOLDAGER PEDERSEN Department of Life Sciences and Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark KAJ SAND-JENSEN Freshwater Biological Laboratory, University of Copenhagen, Helsingørsgade 51, DK-3400 Hillerød, Denmark IVAN VALEILA Boston University, Marine Program, Marine Biology Laboratory, Woods Hole, MA 02543, USA CATHRINE B. PEDERSEN Freshwater Biological Laboratory, University of Copenhagen, Helsingørsgade 51, DK-3400 Hillerød, Denmark FREDE O. ANDERSEN Institute of Biology, University of Southern Denmark, Campusvej 55, DK-5230 Odense M, Denmark JENS KJERULF PETERSEN Department of Marine Ecology, National Environmental Research Institute, P.O. Box 358, DK-4000 Roskilde, Denmark

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JUST CEBRIAN Dauphin Island Sea Laboratory, Department of Marine Science, University of Southern Alabama, 101 Bienville Boulevard, Dauphin Island, AL 36528, USA JACK J. MIDDELBURG Netherlands Institute of Ecology, Centre for Estuarine and Coastal Ecology, P.O. Box 140, NL-4400 AC Yerseke, Netherlands KARLINE SOETAERT Netherlands Institute of Ecology, Centre for Estuarine and Coastal Ecology, P.O. Box 140, NL-4400 AC Yerseke, Netherlands’ PETER M.J. HERMAN Netherlands Institute of Ecology, Centre for Estuarine and Coastal Ecology, P.O. Box 140, NL-4400 AC Yerseke, Netherlands HENRICUS T.S. BOSCHKER Netherlands Institute of Ecology, Centre for Estuarine and Coastal Ecology, P.O. Box 140, NL-4400 AC Yerseke, Netherlands CARLO H.R. HEIP Netherlands Institute of Ecology, Centre for Estuarine and Coastal Ecology, P.O. Box 140, NL-4400 AC Yerseke, Netherlands KAREN J. MCGLATHERY Department of Environmental Sciences, University of Virginia, P.O. Box 400123 Charlottesville, VI 22904, USA

x

KRISTINA S. Gothenburg University, Department of Marine Ecology, P.O. Box 461, S-40530 Gothenburg, Sweden IRIS C. ANDERSON School of Marine Science, Virginia Institute of Marine Science, College of William and Mary, Gloucester Point, VI 23602, USA NILS RISGAARD-PETERSEN Department of Marine Ecology, National Environmental Research Institute, Vejlsøvej 25, DK-8600 Silkeborg, Denmark

PREFACE We have written this book in order to gather existing information on how, and to what extent, different types of marine primary producers affect the nutrient dynamics of coastal marine ecosystems. Like many of our colleagues, we have been deeply involved in research related to coastal eutrophication over the last 10-15 years. We were initially more interested in how changes in nutrient richness (i.e. eutrophication) affected the performance of individual plants, plant populations and the structure of plant communities, but this interest has changed over time. Plants and algae are often highly abundant in shallow coastal marine systems and large amounts of nutrients must therefore be channelled through the autotrophic compartment. We therefore became increasingly interested in how marine plants may modify the processes involved in coastal marine nutrient cycling. Plant communities in coastal ecosystems are often made up of a number of very different plant types (i.e. microalgae, macroalgae and rooted macrophytes) and it is an open question whether variations in community structure and dominance patterns are of any significance for nutrient cycling processes. We had the opportunity to host a special session on this subject at the ASLO (American Society of Limnology and Oceanography) 2000 summer-meeting in Copenhagen, Denmark. Many of our colleagues contributed with very interesting presentations, but it became obvious that most of us tend to study the effect of one specific plant type on one aspect of nutrient cycling (e.g. the effects of seagrasses on benthic nitrification-denitrification). A more holistic systems approach seemed lacking. Most of the ways that plants can affect nutrient cycling in coastal marine ecosystems have been described over the last 20-30 years but new findings continue to appear. Furthermore, relatively little is known about the quantitative importance of these effects relative to each other and to other regulating factors. Few, if any, have to our knowledge attempted to compare the effects of different plant types on overall nutrient dynamics in shallow coastal waters. We have therefore invited a number of colleagues – each a specialist in their own field – to provide a review on the role of different primary producers on specific processes involved in nutrient cycling of shallow coastal ecosystems. This book is the result of their efforts. We would like, first of all, to thank all the authors for their contributions to this book – without you this project would not have been possible. We would also like to thank many of our colleagues and good friends for their inspiration and ideas over the years. Many deserve to be mentioned, but the following persons have played a special role for us because they have stimulated a wealth of good ideas through endless discussions over the years: Jens Borum, Kaj Sand-Jensen, Carlos M. Duarte, Just Cebrián, Karen McGlathery, Frede Ø. Andersen, Marianne Holmer, Erik Kristensen, Daniel Conley, Ole Geertz-Hansen, Dorthe Krause-Jensen, Birgit Olesen, Michael W. Kemp and Rolf Karez. We would also like to thank Kluwer and especially editor Anna Besse-Lototskaya and Judith Terpos from the Kluwer staff for making this book possible and for their patience during the completion of this long overdue work.

xi

CONTENTS AQUATIC ECOLOGY SERIES AIMS & SCOPE

v

CONTRIBUTORS

vii

PREFACE

xi

INTERACTIONS BETWEEN VEGETATION AND NUTRIENT DYNAMICS IN COASTAL MARINE ECOSYSTEMS: AN INTRODUCTION

1

by M.F. Pedersen, S.L. Nielsen and G.T. Banta 1. Introduction

1

2. Coastal Plant Communities and Nutrient Dynamics

2

3. Direct Effects of Plants on Nutrient Dynamics

5

4. Indirect Effects of Plants

9

5. Hypothesis

10

6. References

11

7. Affiliations

15

ESTUARINE PRIMARY PRODUCERS

17

by K. Sand-Jensen and S.L. Nielsen 1. Introduction

17

2. How can estuarine plant communities be characterized?

19

3. What are the geometric consequences of variations in organism size and shape?

23

4. What is the importance of organism size and suspended/attached life form for water motion and solute transport?

25

5. What is the significance of plant size and shape for light utilization, nutrient uptake and growth?

29

6. What are the basic functional properties and interrelations of different photosynthetic communities?

40

7. Summary

49 xiii

xiv

8. Acknowledgements

50

9. References

50

10. Affiliations

57

EFFECTS OF NUTRIENT LOADING ON SHALLOW SEAGRASS-DOMINATED COASTAL SYSTEMS: PATTERNS AND PROCESSES

59

by J. Hauxwell and I. Valiela 1. Introduction

59

2. Nutrients: background, nutrient limitation, and increased loading to coastal zones

60

3. Components of seagrass ecosystems

63

4. Establishing patterns: effect of nutrient loading on assemblages of coastal primary producers

65

5. Understanding processes: how are seagrasses lost?

74

6. Notes

86

7. References

86

8. Affiliations

92

PLANT BOUND NUTRIENT TRANSPORT. MASS TRANSPORT IN ESTUARIES AND LAGOONS

93

by M.R. Flindt, J. Neto, C.L. Amos, M.A. Pardal, A. Bergamasco, C.B. Pedersen and F.Ø. Andersen 1. Introduction

93

2. Plant transport patterns measured in the field

94

3. Macroalgae erosion thresholds and settling rates

98

4. Attached macroalgae sloughing rates

101

5. Macroalgae settling rates

103

6. Relating plant transport to growth and loss rates

106

7. Field studies on plant bound nutrient transport

113

8. Conclusion

123

xv

9. Acknowledgements

125

10. References

125

11. Affiliations

128

GRAZING ON PELAGIC PRIMARY PRODUCERS – THE ROLE OF BENTHIC SUSPENSION FEEDERS IN ESTUARIES

129

by J.K. Petersen 1. Introduction

129

2. Physical forcing

134

3. Eutrophication and benthic suspension feeding

139

4. Where to go from now?

146

5. References

147

6. Affiliations

152

GRAZING ON BENTHIC PRIMARY PRODUCERS

153

by J. Cebrián 1. Introduction

153

2. Grazing on marine benthic producers under pristine conditions: extent, control and consequences

155

3. Towards an understanding of how eutrophication-induced changes in benthic producers assemblages may affect herbivory

162

4. Conclusions and suggestions for future research

174

5. References

177

6. Affiliation

185

DECOMPOSITION OF MARINE PRIMARY PRODUCERS: CONSEQUENCES FOR NUTRIENT RECYCLING AND RETENTION IN COASTAL ECOSYSTEMS

187

by G.T. Banta, M.F. Pedersen and S.L. Nielsen 1. Introduction

187

2. Decomposition

188

xvi

3. Comparing patterns of decomposition and nutrient release during decomposition among marine primary producers

191

4. Changes in C, N and P composition during decomposition

199

5. Nutrient ratios – implications for mineralization and immobilization

203

6. Decomposition patterns – implications for nutrient retention

207

7. Case study – detritus dynamics in two small estuaries dominated by different primary producers

209

8. Conclusion

214

9. References

214

10. Affiliations

216

BURIAL OF NUTRIENT IN COASTAL SEDIMENTS: THE ROLE OF PRIMARY PRODUCERS

217

by J.J. Middelburg, K. Soetaert, P.M.J. Herman, H.T.S. Boschker and C.R. Heip 1. Introduction

217

2. Burial defined

217

3. Sediment accumulation

219

4. Nutrient concentrations

222

5. Importance of primary producers on organic matter burial

225

6. Conclusions

226

7. References

226

8. Affiliations

230

THE IMPORTANCE OF PRIMARY PRODUCERS FOR BENTHIC NITROGEN AND PHOSPHORUS CYCLING

231

by K.J. McGlathery, K. Sundbäck and I.C. Anderson 1. Introduction

231

2. Benthic nutrient cycling processes

232

3. Sediment – water column fluxes

241

4. Competition between primary producers

245

xvii

5. The role of primary producers in the estuarine ‘filter’ – fate and retention of assimilated nutrients

247

6. Changes in nutrient cycling with shifts in primary producer dominance

250

7. Conclusion

251

8. References

252

9. Affiliations

261

DENITRIFICATION

263

by N. Risgaard-Petersen 1. Introduction

263

2. Nitrification and denitrification in sediments

264

3. Denitrification of nitrate from the water column

265

4. Denitrification based on NO3 produced by nitrification

267

5. Alternative NO3 reduction pathways: DNRA

269

6. Denitrification in sediments colonized by benthic microalgae

270

7. Seagrasses and the nitrogen cycle

274

8. Conclusions

276

9. References

277

10. Affiliation

280

ATTEMPTING A SYNTHESIS – PLANT/NUTRIENT INTERACTIONS

281

by S.L. Nielsen, M.F. Pedersen and G.T. Banta 1. Primary effects of eutrophication

281

2. Eutrophication the other way around – effects of plants on nutrient dynamics

283

3. Conclusion

290

4. References

291

5. Affiliations

292

SUBJECT INDEX

293

AQUATIC ECOLOGY SERIES PUBLISHED TITLES

303

MORTEN FOLDAGER PEDERSEN, SØREN LAURENTIUS NIELSEN AND GARY T. BANTA

INTERACTIONS BETWEEN VEGETATION AND NUTRIENT DYNAMICS IN COASTAL MARINE ECOSYSTEMS: AN INTRODUCTION

1. INTRODUCTION The intent of this chapter is to provide a brief background for the ideas and hypotheses that led to the making of this book. Primary producers are quantitatively important in most shallow coastal ecosystems, and although these areas represent less than 2% of the oceanic surface they produce about 20% of the global marine primary production (Charpy-Robaud and Sournia 1990). Autotrophic communities in coastal ecosystems are complex in nature, i.e. they are typically made up of various forms of microalgae (benthic, epiphytic and pelagic), macroalgae (ephemeral and persistent, sensu Littler and Littler 1980) and rooted macrophytes, representing a wide range of life strategies, morphological features and, physiological, functional as well as ecological properties (Littler and Littler 1980; Sand-Jensen and Borum 1991; Duarte 1995; Schramm 1996). These inherent differences may influence the way that different plant types respond to environmental changes (for example eutrophication), but we also expect that they affect the fate of organic matter produced during photosynthesis and, thus, modify major pathways of energy, carbon and plant nutrients (especially N and P). The composition of plant communities that inhabit coastal marine areas may thus play an important role for the functioning of these ecosystems (Duarte 1995). Marine vegetation includes various plant types ranging from unicellular algae to angiosperms. The physiological, functional and ecological properties of the most important types of marine primary producers are presented in detail by Sand-Jensen and Nielsen (chapter 2 this book), but the most distinctive differences between major plant types need a short introduction here. The physiological, functional and ecological differences among microalgae, macroalgae and rooted macrophytes are largely related to size (i.e. thickness and/or relative surface area), shape and structural complexity. Physiological properties, such as intracellular nutrient levels, nutrient uptake rates and demands, photosynthetic capacity and inherent growth rate are all scaled to size so that microalgae and ephemeral macroalgae generally are richer in nutrients, utilize light better and grow faster than large macroalgae and rooted macrophytes (e.g. Duarte 1995, Valiela et al. 1997). The combination of high cellular nutrient levels and fast growth of small algae and plants leads, however, to high requirements for nutrients per unit biomass and time, and smaller plant types often are more sensitive to low nutrient availability than larger plants (e.g. Pedersen and Borum 1996, 1997). A number of functional and ecological properties are also related 1 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 1-15. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

M.F. Pedersen, S.L. Nielsen and G.T. Banta

2

to size and/or morphological complexity. For example, susceptibility to sedimentation, advective transport, grazing and decomposition vary systematically with size and morphological complexity so that small, morphologically simple plants (i.e. unicellular algae and macroalgae with low differentiation) tend to have lower sinking rates, lose a higher proportion of their production through advective transport and grazing and are decomposed faster compared to larger, morphologically more complex plants (Sand-Jensen and Borum 1991; Duarte 1995; Schramm 1996; Valiela et al. 1997; Cebrián 1999).

2. COASTAL PLANT COMMUNITIES AND NUTRIENT DYNAMICS Autotrophic communities of shallow coastal marine ecosystems are highly productive. Total plant biomass ranges typically from 100 to 700 g C m-2 (Fig. 1) while estimates of total primary production range from 150 to more than 700 g C m-2 year-1 (Fig. 2), ranking these systems among the most productive biomes on earth. The high productivity is a result of the combined effect of relatively high nutrient inputs and light penetration to most of the bottom (Sand-Jensen and Borum 1991; Duarte 1995; Borum 1996). Tempelkrogen, Isefjord, DK Childs River, MA, USA Vellerup Cove, Isefjord, DK Roskilde Fjord, Outer Broad, DK Roskilde Fjord, Inner Broad, DK Quashmet River, MA, USA Green Hill. RI, USA Bass Harbor, ME, USA Potter, RI, USA Ninigret Pond, RI, USA Point Judith, RI, USA St. Margareths Bay, CAN Rottnest Island, WA, AUS Sage Lot Pond, MA, USA

0

200

400

600

Total autotrophic biomass (g C m-2)

800

0,0

0,5

1,0

Contribution to total biomass

Figure 1. Total plant biomass (left) and the contribution of different plant groups to total biomass (right) in shallow coastal ecosystems. The systems are ranked after increasing Nloading along the y-axis. Colour codes in right-hand figure: perennial macrophytes (black), annual macroalgae (grey) and, microalgae (white). References are in table 1.

Plant biomass, productivity and the contribution of different plant types to total biomass and productivity are determined by a complex combination of the physical settings (e.g. bathymetry, substrate conditions, tides, wave exposure), prevalent conditions (e.g. salinity, temperature), resource availability (e.g. light conditions, nutrient richness) and biological interactions (e.g. competition and herbivory) (e.g. Sand-Jensen and Borum 1991). The strength of these factors differs substantially

3

Interactions between vegetation and nutrient dynamics

from system to system and total plant biomass, production and composition therefore also vary when compared across systems (figures 1 and 2). Scheldt Estuary, HOL Dollard Estuary, HOL Wadden Sea, HOL Tempelkrogen, Isefjord, DK Childs River, MA, USA Vellerup Cove, Isefjord, DK Roskilde Fjord, Outer Broad, DK Roskilde Cove, Roskilde Fjord, DK Veerse Meer, HOL Roskilde Fjord, Inner Broad, DK Bass Harbor, ME, USA Buttermilk Bay, MA, USA Rhode Island ponds, RI, USA Grevelingenmeer, HOL St. Margareths Bay, CAN Rottnest Island, WA, AUS Corpus Christi Bay, TX, USA Sage Lot Pond, MA, USA

0

200

400

600 -2

800 -1

Total NPP (g C m year )

0,0

0,5

1,0

Relative contribution to NPP

Figure 2. Total primary production (left) and the relative contribution of different plant groups to total production in shallow coastal ecosystems (right). The systems are ranked after increasing N-loading along the y-axis. Colour codes in right-hand figure: perennial macrophytes (black), annual macroalgae (grey) and, microalgae (white). References are in table 1.

The structuring role of major plant nutrients (especially N and P) has received increasingly more attention over the last 2-3 decades because the availability of N and P may affect the biomass, productivity and composition of marine plant communities (e.g. Ryther and Dunstan 1971; Hecky and Kilham 1988; Howarth 1988; Borum and Sand-Jensen 1996). Hauxwell et al. (chapter 3 in this book) discuss the relationship between nutrient richness and autotrophic biomass, productivity and composition in coastal marine ecosystems in detail, but one important point needs to be mentioned here. Plant biomass and production are often expected to increase with eutrophication, but Borum and Sand-Jensen (1996) showed that this is not necessarily the case because large, slow-growing macrophytes with a high biomass per unit area tend to be replaced by small, fast-growing algae with a relatively low biomass per unit area during nutrient enrichment. Hence, total primary production seems unaffected by nutrient richness, whereas the induced changes in composition causes a change in the contribution of different plant groups to total primary production so that microalgae and ephemeral macroalgae contribute more and, perennial macrophytes less, to total plant biomass and primary production with increasing nutrient richness. Plants are however not only affected by the availability of major nutrients, they may also themselves modify processes involved in nutrient cycling and thus potentially affect overall transport, transformation and retention of nutrients. Coastal marine ecosystems receive large amounts of N and P from terrestrial and atmospheric

4

M.F. Pedersen, S.L. Nielsen and G.T. Banta

sources through riverine inputs, run off, dry and wet deposition and N-fixation (Nixon 1995). Table 1. References from which data for figures 1-3 were extracted. Nitrogen loading rates for Rottnest Island (Australia) and St. Margareths Bay (Canada) were obtained from Borum and Sand-Jensen (1996). Conversion factors from chlorophyll to carbon and from dry weight to carbon were 125 and 0.4, respectively.

Coastal ecosystems

Sources from which data were extracted.

Waquoit Bay Estuaries, MA, USA. Valiela et al. 1997; Hauxwell et al. 2003; (Sage Lot Pond, Childs River and Quashmet Hauxwell et al. 1998. River). Corpus Christi Bay, TX, USA.

Flint 1984; Flint 1985.

Rottnest Island, WA, Australia.

Walker et al. 1988; Borum and SandJensen 1996.

St. Margareths Bay, Canada.

Mann 1972a; 1972b; Borum and SandJensen 1996.

Dutch Estuaries, Holland. (Grevelingen, Veerse Meer, Wadden Sea, Dollard, Scheldt Estuaries), Holland.

Nienhuis 1992.

Rhode Island Ponds, RI, USA. (Point Judith, Ninigret Pond, Potter and Green Hill).

Thorne-Miller et al. 1983; Lee and Olsen 1985.

Buttermilk Bay, MA, USA.

Valiela and Costa 1988.

Bass Harbor, ME, USA.

Kinney and Roman 1998.

Roskilde Fjord, Denmark. (Roskilde Cove, Inner and Outer Broads).

Jensen et al. 1990; Borum et al. 1991.

Isefjorden, Denmark. (Vellerup Cove and Tempelkrogen).

Pedersen, Banta and Nielsen, unpublished.

Nutrient inputs vary greatly with size of catchment area, geology, land use, climatic conditions etc., but average 45 g N and 4 g P m-2 year-1 when compared across a number of European and north-American systems (Kaas et al. 1996; Nixon et al. 1996). Coastal marine ecosystems are often conceived as dynamic “filters” because physical, chemical and biological processes affect the composition and amounts of nutrients once they have entered the ecosystem (Schubel and Kennedy 1984). Hence, major nutrients are repeatedly exchanged between dissolved and particulate phases

Interactions between vegetation and nutrient dynamics

5

and between inorganic and organic forms, and elements with a non-conservative behaviour (e.g. N and P) may be “stripped” from the water phase and become temporarily or permanently stored in biota or sediments within the system. The retention of N and P is highly variable when compared across systems but averages 43% and 12% of the inputs, respectively (based on data from Kaas et al. 1996 and Nixon et al. 1996). Retention of N and P is correlated to the input of nutrients and to water residence time (Nixon et al. 1996). Denitrification (in the case of N) and burial of nutrients bound in slowly decomposable organic matter, insoluble salts and complexes or adsorbed to particle surfaces are thought to be the main processes responsible for the observed retention (e.g. Howarth et al. 1995; Nixon et al. 1996). 3. DIRECT EFFECTS OF PLANTS ON NUTRIENT DYNAMICS It is well established that marine plants can affect some of the processes involved in coastal nutrient cycling, but it is presently less well known whether or not this impact is large enough to significantly affect overall nutrient dynamics at the ecosystem level and thus slow down the horizontal (seaward) transport of nutrients, increase the residence time and stimulate the retention of major nutrients. Little is also known about the relative contribution of different plant types to the total effect that plants may have on nutrient cycling. Plants may affect nutrient dynamics directly, through uptake and subsequent immobilization of dissolved nutrients, or indirectly by modifications of the physical and chemical properties of the environment in which they live. Nutrients that are assimilated by plants are temporarily immobilized and may therefore become unavailable for other biogeochemical processes that consume nutrients. Thus, plants may not only compete with other plants for nutrients but also with microbes (e.g. nitrifiers and denitrifiers) and chemical processes for dissolved inorganic nutrients. The direct effects of plants on nutrient dynamics are closely coupled to the fate of the primary production and thus to the susceptibility of various plant types to loss processes such as grazing, decomposition and horizontal export. The uptake of nutrients by plants must be substantial relative to the nutrient inputs if the vegetation of coastal marine ecosystems is to modify overall nutrient dynamics significantly through their mere presence. Unfortunately, few studies have attempted to compare total nutrient assimilation by plants to nutrient inputs, but the high productivity encountered in most coastal ecosystems combined with the fact that the major plant nutrients are assimilated along with carbon in approximate C:N:P-ratios of 106:16:1, 800:49:1 and 435:20:1 for microalgae, macroalgae and rooted macrophytes, respectively (Redfield et al. 1963; Duarte 1992) suggest that nutrient uptake by plants may be substantial. Estimates of total nutrient uptake in a number of systems for which data are available show that uptake of N and P by plants may exceed nutrient inputs from external sources as long as these inputs remain below about 50 g N m-2 and 5-10 g P m-2, respectively (Fig. 3). Most of the nutrients entering a coastal marine ecosystem will thus have to pass through the autotrophic component, and the fate of the produced plant material may therefore play an important role for the subsequent fate of these nutrients. We will therefore now consider how the fate of primary production varies among different plant groups and discuss how such differences may

M.F. Pedersen, S.L. Nielsen and G.T. Banta

6

100 -1

P uptake (g P m year )

20

75

15

-2

-2

-1

N uptake (g N m year )

affect overall nutrient dynamics in coastal marine systems. A conceptual model for this approach is shown in Fig. 4. The storage of nutrients in plant biomass is only temporary since nutrients are lost from the plant component (and from the ecosystem) through export of plant matter or, alternatively, they become remineralized through grazing or decomposition within the system (Duarte and Cebrián 1996; Cebrián 1999).

50 25 0 0

100

200 -2

-1

N loading (g N m year )

10 5 0 0

5

10 15 20 25 30 35

P loading (g P m-2 year-1)

Figure 3. Total N and P uptake by plants in shallow coastal systems with different loading of N and P. Total nutrient uptake was estimated from reports on total annual primary production, the contribution of important autotrophic components to total production and average nutrient concentrations for specific plant types (Duarte 1992). References are in table 1.

Nutrient assimilation by plants therefore only represents net retention of nutrients under non-steady state conditions, for example when biomass accumulates seasonally or when plant biomass or detritus accumulates over longer time scales. Temporary storage of plant-bound nutrients may nevertheless be important because it can reduce rates of seaward transport of nutrients and affect the timing of nutrient availability, delaying the availability of nutrients relative to the optimal growth season. Little is known about the amounts of plant-bound nutrients that are lost from coastal ecosystems through advective transport, but the few studies that are available indicate that these losses can be substantial and therefore potentially important for the nutrient balance in some systems (e.g. Flindt et al. 1997; Salomonsen et al. 1999). The importance of horizontal export of plant material depends largely on the physical conditions of the system (i.e. tides, currents, flushing time etc.) but may also, to some extent, depend on the functional properties of the plants that dominate the system.

Interactions between vegetation and nutrient dynamics

7

Figure 4. Conceptual model showing the possible fate of plant-bound nutrients; inorganic nutrients are taken up and immobilized temporarily in living plant biomass. These nutrients can be lost from the plant community and the ecosystem through advective export or they may become mineralized through grazing. Plant-bound nutrients that are neither exported nor mineralized through grazing enter the detritus compartment where they may be mineralized during decomposition or, alternatively, become buried and thus remain immobilized over long time scales (i.e. decades – centuries).

Large, perennial plants with roots or hold-fasts (i.e. seagrasses, kelps and fucoids) are attached to the substrate and typically have higher sinking rates than free-floating macroalgae and phytoplankton (e.g. Bergamasco et al. 2003). Nutrients bound in large, attached macrophytes should therefore be less susceptible to horizontal transport than dissolved nutrients or nutrients bound in free-floating plants. The assimilation of nutrients by plants should thus reduce horizontal transport and export and, therefore, increase the residence time of nutrients in most cases. The importance of this effect is suspected to increase with increasing dominance of large, perennial and attached macrophytes. Plant-bound nutrients remaining within coastal systems will ultimately be mineralized and recycled through grazing or they will enter the detritus pool. Marine plants are grazed by a wide range of herbivorous animals, and losses of plant matter through herbivory are generally substantial in marine ecosystems (e.g. Cyr and Pace 1993). However, the proportion of net primary production that is lost through herbivory depends partly on the morphological and physiological properties of the plants (e.g. nutrient content and concentrations of phenolic compounds and other defence chemicals; Mattson 1980; Ragan et al. 1986; Hay and Fenical 1988). Losses through herbivory vary therefore systematically among different plant groups (Duarte 1995). Hence, small, nutrient-rich plants tend to lose a much higher proportion of their net primary production to herbivory than large, nutrient-poor plant types, which often contain many structural components and/or have higher concentrations of defence chemicals such as phenolic compounds (e.g. kelps, fucoids, seagrasses; Cebrián and Duarte 1994; Cebrián et al. 1998; Griffin et al. 1998). Assuming that most of the

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plant-bound nutrients consumed by herbivores are lost rapidly through excretion, we expect the amount of mineralized and recycled nutrients to be strongly correlated to grazing losses. Nutrients that are bound in small, nutrient-rich plant types should be recycled faster than when bound in large, nutrient-poor, perennial plants. The role of mineralization through grazing of pelagic and benthic plant types is discussed further by Petersen and Cebrián in chapters 5 and 6, respectively. Plant-bound nutrients may finally reach the detritus pool and become mineralized through decomposition. Decomposition rates and the degree to which detritus is decomposed differ significantly among different plant types and, like grazing, depend on properties such as tissue nutrient levels and the amount of structural compounds and defence chemicals (Valiela et al. 1979; Buchsbaum et al. 1991, Enriquez et al. 1993). Detritus from slow-growing, perennial macrophytes is typically nutrient-poor and contains large amounts of structural components, and detritus from these plants decomposes much more slowly and less completely than detritus originating from small, nutrient rich plants (Enriquez et al. 1993, Cebrián et al. 1998). The anaerobic conditions that often appear in the sediments where most decomposition occurs, may further slow down rates of decomposition. This is especially true for slowly decomposable matter originating from nutrient-poor plants (e.g. Benner et al. 1984) whereas decomposition of easily decomposable matter seems less affected by redox conditions (e.g. Andersen 1996). The turnover time of detritus-bound nutrients may therefore increase considerably when they originate from large, perennial macrophytes. The combination of higher inputs of detritus and lower turn-over rates should lead to accumulation and high steady state stocks of detritus-bound nutrients when plant communities are dominated by large, perennial macrophytes. Banta et al. compare rates of decomposition and mineralization among different marine primary producers and discuss the possible implications of these variations in chapter 7. Burial of nutrients is one of the major mechanisms behind nutrient retention in coastal marine ecosystems, but whether or not marine plants contribute significantly to these nutrient losses is less well known. Slow turn-over and accumulation of detritus should theoretically increase the chance for long term burial of organic matter and associated nutrients, and detritus from large, nutrient-poor plants should, other things being equal, have a greater chance of being buried than detritus originating from small, nutrient-rich plants. Hence, sediments below dense seagrass meadows are often much richer in organic matter and nutrients than neighbouring areas without perennial vegetation (Kenworthy et al. 1982; Pedersen et al. 1997). It is however not clear whether dominance by slow-growing, perennial macrophytes actually leads to a larger burial of organic matter and nutrients at the ecosystem level and, if so, whether such differences are related to inherent variations in decomposition rates among different plant types or to other effects that these plants may have on their environment (e.g. influence on current speed and sedimentation rates; effects on redox potentials - see below). Another question is, whether the amount of nutrients contained in autochonous derived detritus that is buried is of any significance when compared to the burial of nutrients contained in organic matter of allochthonous origin. Middelburg et al. provides an interesting discussion on these subjects in chapter 8. In summary, plants typically assimilate large percentages of the nutrients received by coastal marine ecosystems and should therefore be able to affect overall nutrient dynamics of the system in which they reside. The strength of these effects is

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determined, in part, by the fate of the plant material, and is thus linked to inherent properties of various plant types such as size and shape (see chapter 2 for a further discussion). 4. INDIRECT EFFECTS OF PLANTS In addition to the direct effects plants have on nutrient dynamics through uptake and more or less temporary immobilization of nutrients, they have indirect effects on nutrient dynamics as well through their influence on the physico-chemical environment. Many marine plants are situated at the sediment-water interface (e.g. rooted macrophytes and benthic microalgae) and may therefore potentially be able to affect the benthic-pelagic exchange of nutrients. Dense populations of rooted macrophytes and large macroalgae may thus slow down water movement and reduce the chance of resuspension events and increase net sedimentation rates (e.g. Fonseca et al. 1982; Short and Short 1984). As mentioned above, these effects often result in an increased input of particulate nutrients of allochthonous and autochonous origin to the sediments in macrophyte dominated areas, which may stimulate both the turnover and the burial of nutrients within the sediment (Kenworthy et al. 1982). Seagrasses and other large macrophytes may thus stimulate conservation of nutrients through effects on the physical environment. Dense populations of benthic microalgae and/or extensive mats of free-floating ephemeral macroalgae may also stimulate a downward flux of dissolved inorganic nutrients through uptake of nutrients from the overlying water and, at the same time, intercept the efflux of nutrients from the sediment through efficient assimilation of nutrients originating from the sediment porewater. Such “filter-effects” may however be strongly variable in space and time because mats of both benthic microalgae and ephemeral, free-floating macroalgae are susceptible to water movement and resuspension and tend to be heterogeneously scattered over the bottom (KrauseJensen et al. 1996). Plants may also interfere with benthic nutrient cycles through modification of oxygen levels, redox potentials and pH-levels within the sediment because changes in these factors may affect the solubility of chemical compounds and influence processes such as nutrient adsorption to particles, ammonification, nitrification, denitrification and N-fixation (e.g. Pomeroy et al. 1965; Howarth et al. 1995; Wigand et al. 1997). Benthic microalgae and rooted macrophytes modify oxygen levels within sediments through leakage of oxygen directly or from roots (of seagrasses) during photosynthesis (e.g. Revsbech et al. 1981; Sand-Jensen et al. 1982, Kemp and Murray 1986). The aerobic zones in the uppermost regions of the sediment and in the rhizosphere of seagrasses may stimulate the formation and maintenance of insoluble P-compounds (especially metal-phosphorus compounds; Pomeroy et al. 1965; Howarth et al. 1995) and, thus, reduce the efflux of nutrients to the water phase. Increasing the oxygen levels in the sediment may further stimulate nitrification (Henriksen and Kemp 1984; Rysgaard et al. 1994) and possibly even coupled nitrification-denitrification (e.g. Caffrey and Kemp 1990; 1992). The net effect on nitrification-denitrification is however not clear since benthic microalgae and rooted macrophytes may compete with nitrifying bacteria for ammonia (Henriksen and Kemp 1984; Risgaard-Petersen and Ottosen 2000). Rooted macrophytes and benthic

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microalgae may thus serve to conserve nutrients within the sediment and may even, under certain circumstances, stimulate a permanent loss of nitrogen through denitrification. Rooted macrophytes may, however, also stimulate mobilization of otherwise immobilized nutrients by releasing organic acids from their roots, lowering pH. Adsorption of ammonia to particle surfaces depends partly on pH of the media, and these ions become more mobile at low pH (Rosenfeld 1979). Seagrasses can also mobilize phosphorus under nutrient poor conditions. Jensen et al. (1998) showed that seagrasses in P-limited systems are able to increase the bio-availability of P bound in insoluble Ca-complexes by lowering the pH of the rhizosphere. Whether or not these mobilized nutrients are acquired by the plants or diffuse to the water column is presently unknown. Marine plants seem thus able to affect nutrient cycling indirectly in a large number of ways. Although our knowledge of how different plant groups affect benthic-pelagic exchange of nutrients and benthic nutrient cycling is rapidly growing, we know little about the net effects of these processes at the ecosystem level. McGlathery et al. and Risgaard-Petersen discuss the importance of these indirect effects in much more detail in chapters 9 and 10, respectively. 5. HYPOTHESES This brief introduction has shown that plants are able to modify processes involved in nutrient cycling of coastal marine ecosystems in a variety of ways. Are such effects of any significance relative to other factors and does dominance of one or the other plant group play any role in the degree to which plant communities affect nutrient cycling? Unfortunately, no single study have yet, to our knowledge, aimed to compare the effect of different plant types on the cycling of N and P and to compare these effects to other biological, chemical and physical effects. Work by Duarte and Cebrián (1996) and Cebrián et al. (1998; 1999) showed, however, that dominance of certain plant types can play a significant role for the way that carbon is cycled in marine ecosystems. Hence, most of the carbon fixed through photosynthesis is channelled through grazers and decomposers while less is lost through export or permanent burial in systems dominated by microalgae and ephemeral macroalgae. In contrast, marine systems dominated by slow-growing, perennial macrophytes appear to be sinks for carbon because a substantial fraction of the primary production is buried permanently and less is channelled through herbivores and decomposers. We hypothesize that the composition and dominance patterns of coastal plant communities may influence cycles of nitrogen and phosphorus in much the same way, since living and dead plant material contains both these elements along with carbon. Thus, we expect overall nutrient dynamics (i.e. transport, export and retention) of coastal marine ecosystems to be controlled by the primary producers when large, perennial and slow-growing macrophytes are dominating the plant community because nutrients are immobilized in living biomass over long time scales due to the slow turnover of biomass in these plants. Losses of plant-bound nutrients through export are expected to be low because horizontal transport of plant material is much slower than for suspended particulate and dissolved nutrients. Mineralization through grazing is also expected to be low, and most of the plant-bound nutrients that

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were assimilated during growth should ultimately enter the detritus compartment where they remain immobilized over relatively long time scales due to slow and incomplete turnover of detritus. The presence of below-ground tissues (roots and rhizomes in seagrasses) combined with slow decomposition due to low nutrient content, high concentrations of refractory components and anaerobic conditions are expected to increase the chance of long-term burial of plant-bound nutrients within the sediments. The presence of attached macrophytes may additionally stimulate sedimentation of suspended, particle-bound nutrients. Rooted macrophytes may further stimulate physical and chemical immobilization of dissolved inorganic nutrients and coupled nitrification-denitrification in the sediments through leakage of oxygen from the roots. We hypothesize, in contrast, that the potential to control overall nutrient dynamics is much lower for small, fast-growing plants. Nutrients assimilated by these plants are only immobilized for short time periods due to the fast turnover of living biomass. The export of plant-bound nutrients may only be marginally slower than for dissolved inorganic nutrients because most of these plant types are free-floating or suspended in the water and subject to almost the same advective transport as dissolved nutrients. The turnover of plant-bound nutrients is increased by efficient grazing and mineralization, and a relatively low proportion of the plant-bound nutrients may reach the detritus pool. Decomposition and mineralization are rapid, and almost complete and long-term burial of detritus-bound nutrients should therefore be insignificant. Hence, the potential impact of fast-growing plants on overall nutrient cycles should be relatively small and overall transport and retention should not be severely affected by dominance of these plant types. In the following chapters the expectations outlined above will be evaluated. In chapter 11, we will give a synthesis with an emphasis on any modifications necessary when these general hypotheses are tested in complex ecosystems. 6. REFERENCES Andersen, F.Ø. (1996). Fate of organic carbon added as diatom cells to oxic and anoxic marine sediment microcosms. Marine Ecology Progress Series, 134, 225-233. Benner, R., MacCubbin, A.E. & Hodson, R.E. (1984). Anaerobic biodegradation of the lignin and polysaccharide components of lignocellulose and synthetic lignin by sediment microflora. Applied Environmental Microbiology 47: 998-1004. Bergamasco, A., De Nat, L., Flindt, M.R. & Amos, C.L. (2003). Interactions and feedbacks among phytobenthos, hydrodynamics, nutrient cycling and sediment transport in estuarine ecosystems. Continental Shelf Research 23, 1715-1741. Borum, J., Geertz-Hansen, O., Sand-Jensen, K. & Wium-Andersen, S. (1991).Eutrophication: Effects on marine plant communities. In: Nitrogen and phosphorus in fresh and marine waters (pp. 40-54). Miljøstyrelsen, Denmark. Borum, J. (1996). Shallow waters and Land/Sea Boundaries. In B.B. Jørgensen & K. Richardson (Eds.), Eutrophication in Coastal Marine Ecosystems (Vol. 52, pp. 179-203). Washington DC: American Geophysical Union. Borum, J. & Sand-Jensen, K. (1996). Is total primary production in shallow coastal marine waters stimulated by nitrogen loading? Oikos 76, 406-410.

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Buchsbaum R., Valiela, I., Swain, T., Dzierzeski, M. & Allen, S. (1991). Available and refractory nitrogen in detritus of coastal vascular plants. Marine Ecology Progress Series 72, 131-143. Caffrey, J.M. & Kemp, W.M. (1990). Nitrogen cycling in sediments with estuarine populations of Potamogeton perfoliatus and Zostera marina. Marine Ecology Progress Series 66, 147-160. Caffrey, J.M. & Kemp, W.M. (1992). Influence of the submerged plant, Potamogeton perfoliatus, on nitrogen cycling in estuarine sediments. Limnology and Oceanography 37, 1483-1495. Cebrián, J. (1999). Patterns in the fate of production in plant communities. The American Naturalist 154, 449-468. Cebrián, J. & Duarte, C.M. (1994). The dependence of herbivory on growth rate in natural plant communities. Functional Ecology 8, 518-525. Cebrián, J., Williams, M., McClelland, J., & Valiela, I. (1998). The dependence of heterotrophic consumption and C accumulation on autotrophic nutrient content in ecosystems. Ecology Letters 1, 165-170. Charpy-Robaud, C. & Sournia, A. (1990). The comparative estimation of phytoplanktonic, microphytobenthic, and macrophytobenthic primary production in the oceans. Marine Microbial Food Webs 4, 31-57. Cyr, H. & Pace, M.L. (1993). Magnitude and patterns of herbivory in aquatic and terrestrial ecosystems. Nature 361, 148-150. Duarte, C.M. (1992). Nutrient concentration of aquatic plants: Patterns across species. Limnology and Oceanography 37, 882-889. Duarte, C.M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia 41, 87-112. Duarte, C.M. & Cebrián, J. (1996). The fate of marine autotrophic production. Limnology and Oceanography 41, 1758-1766. Enriquez, S., Duarte, C.M. & Sand-Jensen, K. (1993). Patterns in decomposition rates among photosynthetic organisms: The importance of detritus C:N:P. Oecologia 94: 457-471. Flindt, M.R., Salomonsen, J., Carrer, M., Bocci, M. & Kamp-Nielsen, L. (1997). Loss, growth and transport dynamics of Chaetomorpha aerea and Ulva rigida in the Lagoon of Venice during an early summer field campaign. Ecological Modelling 102, 133-141. Flint, R.W. (1984). Phytoplankton production in the Corpus Christi Bay estuary. Contributions in Marine Science 27, 65-85. Flint, R.W. (1985). Long-term estuarine variability and associated biological response. Estuaries 8, 158169. Fonseca, M.S., Fisher, J.S., Zieman, J.C. & Thayer, G.W. (1982). Influence of the seagrass, Zostera marina L., on current flow. Estuarine and Coastal Shelf Science 15, 351-364. Griffin, M.P.A., Cole, M.L., Kroeger, K.D. & Cebrián, J. (1998). Dependence of herbivory on autotrophic nitrogen content and on net primary production across ecosystems. Biological Bulletin 195, 233-234. Hauxwell, J., Cebrián, J. & Valiela, I. (2003). Eelgrass Zostera marina loss in temperate estuaries: relationship of light limitation imposed by algae. Marine Ecology Progress Series 247, 59-73. Hauxwell, J., McClelland, J., Behr, P. & Valiela, I. (1998). Relative importance of grazing and nutrient controls of macroalgal biomass in three temperate shallow estuaries. Estuaries 21, 347-360.

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Hay, M.E. & Fenical, W. (1988). Marine plant-herbivore interactions: The ecology of chemical defence. Annual Review of Ecology and Systematics 19, 111-145. Hecky, P.E. & Kilham, P. (1988). Nutrient limitation of phytoplankton in freshwater and marine environments: A review of recent evidence on the effects of enrichment. Limnology and Oceanography 33, 796-822. Henriksen, K. & Kemp, W.M. (1988). Nitrification in estuarine and coastal marine sediments. In T.H. Blackburn and J. Sørensen (Eds.), Nitrogen cycling in coastal marine environments (SCOPE 33, pp. 207-250). U.K.: John Wiley & Sons. Howarth, R.W. (1988). Nutrient limitation of net primary production in marine ecosystems. Annual Review of Ecology and Systematics 19, 89-110. Howarth, R.W., Jensen, H.S., Marino, R. & Postma, H. (1995). Transport to and processing of P in nearshore and oceanic waters. In H. Tiessen (Ed.), Phosphorus in the global environment - transfers, cycles and management (SCOPE 54, pp. 323-346.). U.K.: Wiley. Jensen, H.S., McGlathery, K.J., Marino, R. & Howarth, R.W. (1998). Forms and availability of sediment phosphorus in carbonate sand of Bermuda seagrass beds. Limnology and Oceanography 43, 799-810. Jensen, L.M., Sand-Jensen, K., Marcher, S. & Hansen, M. (1990). Plankton community respiration along a nutrient gradient in a shallow Danish estuary. Marine Ecology Progress Series 61, 75-85. Kaas, H., Møhlenberg, F., Josefson, A., Rasmussen, B., Krause-Jensen, D., Jensen, H.S., Svendsen, L.M., Windolf, J., Middelboe, A.L., Sand-Jensen, K. & Pedersen, M.F. (1996). Marine områder. Danske fjorde – status over miljøtilstand, årsagssammenhænge og udvikling. Vandmiljøplanens Overvågningsprogram 1995. Danmarks Miljøundersøgelser. 205 pp. Faglig rapport fra DMU nr. 179 (in Danish). Kemp, W.M. & Murray, L. (1986). Oxygen release from roots of the submerged macrophyte, Potamogeton perfoliatus L. regulating factors and ecological implications. Aquatic Botany 26, 271-283. Kenworthy, W.J., Zieman, J.C. & Thayer, G.W. (1982). Evidence for the influence of sea grasses on the benthic nitrogen-cycle in a coastal-plain estuary near Beaufort, North-Carolina (USA). Oecologia 54, 152-158. Kinney, E.H. & Roman, C.T. (1998). Response of primary producers to nutrient enrichment in a shallow estuary. Marine Ecology Progress Series 163, 89-987. Krause-Jensen, D., McGlathery, K., Rysgaard, S. & Christensen, P.B. (1996). Production within dense mats of the filamentous macroalga Chaetomorpha linum in relation to light and nutrient availability. Marine Ecology Progress Series 134, 207-216. Lee, V. & Olsen, S. (1985). Eutrophication and management initiatives for the control of nutrient inputs to Rhode Island coastal lagoons. Estuaries 8, 191-202. Littler, M.M. & Littler, D.S. (1980). The evolution of thallus form and survival strategies in benthic marine macroalgae: Field and laboratory tests of a functional form model. American Naturalist 116: 25-44. Mann, K.H. (1972a). Ecological energetics of the seaweed zone in a marine bay on the Atlantic coast of Canada I. Zonation and biomass of seaweeds. Marine Biology 12, 1-10. Mann, K.H. (1972b). Ecological energetics of the seaweed zone in a marine bay on the Atlantic coast of Canada II. Productivity of seaweeds. Marine Biology 14, 199-209. Mattson, W.J. Jr. (1980). Herbivory in relation to plant nitrogen content. Annual Review of Ecology and Systematics 11, 119-161.

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Nienhuis, P.H. (1992). Eutrophication, water management, and the functioning of Dutch estuaries and coastal lagoons. Estuaries 15, 538-548. Nixon, S.W. (1995). Coastal marine eutrophication: a definition, social causes, and future concerns. Ophelia 41, 199-219. Nixon, S.W., Ammerman, J.W., Atkinson, L.P., Berounsky, V.M., Billen, G., Boicourt, W.C., Boynton, W.R., Church, T.M., Ditoro, D.M., Elmgren, R., Garber, J.H., Giblin, A.E. Jahnke, R.A., Owens, N.J.P., Pilson, M.E.Q. & Seitzinger, S.P. (1996). The fate of nitrogen and phosphorus at the land-sea margin of the North Atlantic Ocean. Biogeochemistry 35, 141-180. Pedersen, M.F. & Borum, J. (1996). Nutrient control of algal growth in estuarine waters. Nutrient limitation and the importance of nitrogen requirements and nitrogen storage among phytoplankton and species of macroalgae. Marine Ecology Progress Series 142, 261-272. Pedersen, M.F. & Borum, J. (1997). Nutrient control of estuarine macroalgae: growth strategy and the balance between nitrogen requirements and uptake. Marine Ecology Progress Series 161, 155-163. Pedersen, M.F., Duarte, C.M. & Cebrián, J. (1997). Rates of changes in organic matter and nutrient stocks during seagrass Cymodosea nodosa colonization and stand development. Marine Ecology Progress Series 159, 29-36. Pomeroy, L.R., Smith, E.E. & Grant, C.M. (1965). The exchange of phosphate between estuarine waters and sediments. Limnology and Oceanography 10, 167-172. Ragan, M.A. & Glombitza, K.W. (1986). Phlorotannins, brown algal polyphenols. In F.E. Round & D.J. Chapman (Eds.), Progress in Phycological Research (Vol. 4, pp 129-210). U.K.: Biopress Ltd. Redfield, A.C., Ketchum B.A. & Richards, F.A. (1963). The influence of organisms on the chemical composition of sea-water. In M.N. Hill (Ed.), The Sea (pp 26-77). U.K.: Wiley. Revsbech, N.P., Jørgensen, B.B. & Brix, O. (1981). Primary production of microalgae in sediments measured by oxygen microprofile, H14CO3 fixation, and oxygen exchange methods. Limnology and Oceanography 26, 717-730. Risgaard-Petersen, N. & Ottosen, L.D.M. (2000). Nitrogen cycling in two temperate Zostera marina beds: seasonal variation. Marine Ecology Progress Series 198, 93-107. Rosenfeld, J.K. (1979). Ammonium adsorption in nearshore anoxic sediments. Limnology and Oceanography 24, 356-364. Rysgaard, S., Risgaard-Petersen, N., Sloth, N.P., Jensen, K. & Nielsen, L.P. (1994). Oxygen regulation of nitrification and denitrification in sediments. Limnology and Oceanography 39, 1643-1652. Ryther, J.H. & Dunstan, W.H. (1971). Nitrogen, phosphorus, and eutrophication in the coastal marine environment. Science 171, 1008-1013. Salomonsen, J., Flindt, M.R., Geertz-Hansen, O. & Johansen, C. (1999). Modelling advective transport of Ulva lactuca (L) in the sheltered bay, Møllekrogen, Roskilde Fjord, Denmark. Hydrobiologia 397, 241-252. Sand-Jensen, K. & Borum, J. (1991). Interaction among phytoplankton, periphyton, and macrophytes in temperate freshwaters and estuaries. Aquatic Botany 41, 137-176. Sand-Jensen, K., Prahl, C. & Stokholm, H. (1982). Oxygen release from roots of submerged aquatic macrophytes. Oikos 38, 349-354. Schramm, W. (1996). Conclusions. In W. Schramm & P.H. Nienhuis (Eds.), Marine benthic vegetation – recent changes and the effects of eutrophication (pp. 449-458). Berlin: Springer Verlag.

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Schubel, J.R. & Kennedy, V.S. (1984). The estuary as a filter: an introduction. In V.S. Kennedy (Ed.), The estuary as a filter (pp. 1-11). U.S.A.: Academic Press. Short, F.T. & Short, C.A. (1984). The seagrass filter: Purification of estuarine and coastal waters. In V.S. Kennedy (Ed.), The estuary as a filter (pp. 395-413). U.S.A.: Academic Press. Thorne-Miller, B., Harlin, M.M., Thursby, G.B., Brady-Campbell, M.M. & Dworetzky, B.A. (1983). Variations in the distribution and biomass of submerged macrophytes in five coastal lagoons in Rhode Island, U.S.A. Botanica Marina 26, 231-242. Valiela, I. & Costa, J.E. (1988). Eutrophication of Buttermilk Bay, a Cape-Cod coastal embayment – concentrations of nutrients and watershed nutrient budgets. Environmental Management 12, 539-553. Valiela, I., Koumjian, L. & Swain, T. (1979). Cinnamic acid inhibition of detritus feeding. Nature 280, 5557. Valiela, I., McClelland, J. Hauxwell, J. Behr, P. Hersh, D. & Foreman, K. (1997). Macroalgal blooms in shallow estuaries: Controls and ecophysiological and ecosystem consequences. Limnology and Oceanography 42, 1105-1118. Wigand, C., Stevenson, J.C. & Cornwell, J.C. (1997). Effects of different submersed macrophytes on sediment biogeochemistry. Aquatic Botany 56, 233-244. Walker, D.I., Masini, R.J. & Pailing, E.I. (1988). Comparison of annual production and nutrient status of the primary producers in a shallow limestone reef system (Rottnest Island), Western Australia. In Proceedings from the 25th Australian Marine Sciences Association Conference (pp. 183-187). University of Sydney.

7. AFFILIATIONS M.F. Pedersen, S.L. Nielsen & G.T. Banta: Department of Life Sciences and Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark.

KAJ SAND-JENSEN AND SØREN LAURENTIUS NIELSEN

ESTUARINE PRIMARY PRODUCERS 1. INTRODUCTION The marine primary producers exist in a very large variety of sizes and shapes and live in many different habitats (Sand-Jensen and Borum 1991, Hemminga and Duarte 2000). It may therefore seem almost unmanageable to build an overview and to formulate general relationships between organism size, shape and habitat on the one hand and functional properties of species and plant communities on the other. The risk is that precise statements regarding behaviour and function can only be established for selected species in a given habitat, while the behaviour and function of the diverse assemblage of species in several estuarine habitats will remain uncertain or unknown. If this was so only a small fraction of the complex world could be described and understood. We have today, however, good possibilities of analyzing many species, enabling us to reach general assessments. It has become possible to formulate a number of common principles in broad-scale analysis, covering differences in organism size, shape and habitat (Duarte 1995, Duarte et al. 1995, Niklas 1994, 1997). Differences in organism size and shape have a number of general consequences both for the environmental conditions that the organisms experience and respond to and for their ability to photosynthesize, grow and survive (Nielsen and Sand-Jensen 1990, Agusti et al. 1994, West et al. 1997, Enquist et al. 1998). In addition, different habitats create characteristic conditions regarding light availability, water movement, nutrient supply and physical stability. We are certainly far from having a unified overview of the environmental conditions, plant traits and responses and the relative roles of the various marine primary producers in ecosystems. These shortcomings will be clear from this chapter and have several reasons. One reason is the historical lack of methods to provide an accurate and appropriate description of the physical and chemical conditions with sufficient temporal and spatial resolution to cover the organisms in their diverse habitats in the water column, on the rocks, in the sediments and on the surfaces of other plants. Many of these problems can now be handled with the use of equipment that continuously measures light, temperature, water movement, oxygen, pH, sulphide, nitrate, etc. in macro- and microhabitats, but the potential of this new technology has so far not been fully exploited (e.g. Kühl and Revsbech 2001). Methodological problems still remain in measuring plant function under natural field conditions, but newly developed fluorescence methods hold promising possibilities for frequent measurements of photosynthesis at different scales (Maxwell and Johnson 2000). Another reason for the lack of robust predictions of the behaviour of estuarine primary producers is the complexity involved in their regulation which by far surpasses that of primary producers in lakes (Cloern 2001, Nixon 2001). While 17 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 17-57. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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phosphorus loading models incorporating external phosphorus input, water renewal rate, water depth and sediment phosphorus retention are commonly capable of accounting for 75-90% of the variability in phytoplankton biomass among the entire suite of oligo-, meso- and eutrophic lakes (Vollenweider 1976, Dillon and Rigler 1974, Schindler 1987), nitrogen loading models for estuaries can usually account for much less of the variability in phytoplankton biomass or productivity; for example about 35% in fifty estuaries compared by Borum (1996). The open character of estuaries can lead to profound horizontal losses of phytoplankton (Lucas et al. 1999a,b), the variable tidal regime of estuaries strongly influences vertical stability and particle concentrations of the water column and, thereby, the light climate and growth of phytoplankton (Monbet 1992, Cloern 1999, 2001), the large benthic suspension feeders (e.g. bivalves) can have a very strong, but highly variable, grazing impact on phytoplankton biomass and turbidity of estuaries (Kaas et al. 1996; Meuwig et al. 1998), and drift macroalgae can be important primary producers, animal food and shading components in estuaries (Geertz-Hansen et al. 1993, Nixon 2001, chapter 4 Flindt ). Major investement in comparative analyses among systems to synthesize the lessons from many site-specific studies and simultaneously deal with several obviously important regulators of the different primary producers (e.g. nutrients, light availability, substratum stability, grazing, transport losses) has been fruitful in the past and remains a promising tool also for the future. Establishing an overview has been delayed by flaws in traditional research strategies. It has remained a problem than many researchers prefer to concentrate on one out of many species or on a single plant community out of the three to four communities that dominate in a particular ecosystem. For a long time, the understanding of the environment and growth conditions of the attached microalgae and macrophytes remained biased by the misconception that they live under comparable conditions and regulating principles to those of the phytoplankton (Sand-Jensen 1989). The overview of aquatic plant communities has mainly been established by comparisons of resource utilization, photosynthesis, respiration and growth for a large number of species, sizes and shapes of a given plant type under controlled conditions in the laboratory, allowing many environmental parameters to be kept constant (e.g. Markager and Sand-Jensen 1994, 1996, Enriquez et al. 1996). Or in some cases it has been established by measurements under standardised conditions in nature, where various manipulations in chambers, cages etc. have been used (Havens et al. 2001). Many empirical relationships have therefore been established between the function of photosynthetic organisms and their size, shape and life form under given environmental conditions and resource supply (Duarte 1995). For the mixed communities of benthic and epiphytic microalgae this type of information is still very sparse. Another largely unresolved problem – both conceptually and methodologically – in the study of photosynthetic organisms and communities is the significance of loss processes for their abundance and biomass (Cebrian and Duarte 1994, 1995; Duarte and Cebrian 1996). Organism biomass is a net result of growth minus loss processes, and it is particularly the growth that can be predicted from the environmental conditions and the size, shape and life form of the organisms. The loss processes through grazing, parasite attack, decomposition, sinking and horizontal transport which are complex and very difficult to quantify. Not only are plants engaged in intricate interactions with each other, but animals are also part of these interactions,

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because they consume different species and plant groups with variable intensity and effect. The grazing loss, for example, depends on the number of grazers relative to the grazed plant, the grazing potential of the grazers, their food demand and the presence of alternative food sources (Geertz-Hansen et al. 1993). The complex regulation of food webs is part of the reason why reliable predictions of the abundance and role of various primary producers in coastal areas have been so difficult to reach. On the other hand, judging from recent progress in the understanding of food web regulation in freshwater lakes (e.g. Meijer et al. 1994), this is a promising research theme where original discoveries can be made and predictions established by controlled experiments and cross-system comparisons. In this chapter we focus on photosynthetic organisms. We present a brief characterisation of the various types of estuarine primary producers, their habitats, the differences between species regarding size and shape and the consequences for nutrient utilization, photosynthesis, growth and biomass accumulation. We have organized the chapter around five questions: − How can estuarine plant communities be characterized? − What are the geometrical consequences of variations in organism size and shape? − What is the importance of organism size and suspended vs. attached life habits for water motion and solute transport? − What is the significance of organism size and shape for light utilization, nutrient uptake, and growth? − What are the basic functional properties and interrelations between different photosynthetic communities? 2. HOW CAN ESTUARINE PLANT COMMUNITIES BE CHARACTERIZED? 2.1 Taxonomic affiliation, species richness and its consequences Estuarine plant types comprises many and taxonomically very diverse groups of algae and a group of vascular plants among the rooted, monocot angiosperms – including seagrasses and a few species of Potamogeton, Zannichellia etc. (Table 1). Photosynthetic organisms are grouped into four kingdoms: Eubacteria, Protozoa, Plantae and Chromista. The microscopic and macroscopic algae are found in 11 groups distributed over all kingdoms with approximately 17,000 known species all together, while the marine vascular plants only comprise about 55 species of seagrasses (Hemminga and Duarte 2000). The diversity within the collective group algae is enormous regarding cellular structure (prokaryotes vs. eukaryotes), pigmentation (red, green and brown algae), chloroplast structure, biochemistry, reproduction and size. In comparison, all seagrasses are monocot angiosperms, and more similar to each other, though of polyphyletic origin (Les et al. 1997). Some algal groups only contain unicells (e.g. diatoms), others contain both unicells and multicellular forms (e.g. green algae) while others still only contain multicellular forms (e.g. brown algae). The unicells can grow attached or suspended in the plankton (e.g. dinoflagellates and chrysophytes). The multicellular forms have difficulties remaining in suspension and primarily grow attached or in some cases as drift algae.

K. Sand-Jensen and S.L. Nielsen

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Table 1. Overview of the main taxonomic groups of photosynthetic organisms and the number of known species within them in marine, freshwater and all environments together. Groups with only a few known species (< 30) are not included. The term division is equivalent to phylum. Adapted from Falkowski and Raven (1997).

Taxonomic group Kingdom: Eubacteria Subdivision: Cyanobacteria Kingdom: Protozoa Division: Euglenophyta Division: Dinophyta Kingdom: Plantae Division: Rhodophyta Division: Chlorophyta Division: Bryophyta: Division: Lycopsida Division: Filicopsida Division: Magnoliophyta Subdivision: Monocotyledoneae Subdivision: Dicotyledoneae Kingdom: Chromista Division: Cryptophyta Division: Haptophyta Division: Heterokonta Class: Bacillariophyceae Class: Chrysophyceae Class. Fucophyceae Class: Synurophyceae Class: Tribophyceae

Marine

Freshwater

All aquatic and terrestrial environments

150

1350

1500

30 1800

1050 200

1080 2000

5880 1300

120 14920 1000 70 94

6000 16220 22000 1228 8400

55

455 391

52000 188000

100 100

100 400

200 500

5000 800 1497

5000 200 3 250 550

10000 1000 1500 250 600

50

The species richness among the photosynthetic organisms of the sea is relatively limited compared to the terrestrial environment with most known species among red algae, diatoms, dinoflagellates, brown algae and green algae (Table 1). About twothirds of the marine species are macroscopic and an even larger fraction grows attached, while a smaller fraction consists of planktonic species. The high density and high dispersal ability of microscopic planktonic species and the fact that the marine environment is continuous without strong physical boundaries promotes the flow of genes and constrains the evolution of new species (Fenchel 1993, Rapoport 1994, Fenchel et al. 1997). Density and dispersal ability are less among the attached algae, and their growth habitats are distributed as a mosaic along the coastlines, contributing

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to a larger diversity of habitats, a stronger genetic isolation and, thereby, a greater likelihood of species evolution. It is striking that species richness is so much larger among the macroscopic flowering plants on land than among all photosynthetic organisms in the sea. One explanation is that terrestrial plants exist in lower densities and have weaker dispersal potentials than marine species in general and planktonic species in particular. A second explanation is that terrestrial plants grow in very diverse and physically disconnected environments with strong dispersal barriers between them – for example mountains on land and oceans between islands and continents. A third explanation is that the specialised relationships between insects and terrestrial flowering plants promote reproductive isolation and species formation. These principal differences between marine and terrestrial environments should also have important consequences for the extinction risk of species following profound anthropogenic perturbations such as global climate changes. Because of stronger physical barriers, lower dispersal potentials, specialised plant-insect pollination systems and more restricted distribution of terrestrial plants many more terrestrial plant species than marine photosynthetic species are likely to succumb in the face of altered global temperatures. On land additional problems are the changes in precipitation, drought and storm events which should make it even more difficult for species to adjust their distribution patterns to the future climate. Local species richness in high-saline areas often comprises more than 100 species of planktonic microalgae, benthic microalgae and attached macroalgae (Middelboe and Sand-Jensen 1998, Sand-Jensen 2000). In most places species differences are so extensive with respect to morphology, physiology and ecology that some of them are likely to tolerate or even benefit from new environmental conditions caused by anthropogenic impact. Mass occurrences of certain species in the plankton and among benthic macroalgae are obvious examples of how they can benefit from environmental changes in nutrient status. In brackish areas of low species diversity, however, the entire ecosystem function can be at stake if a key species is suddenly severely hampered by pollution or climate change. In the Baltic Sea proper, Fucus vesiculosus is the only common representative of the large leathery macroalgae, forming the structural backbone of macroalgae communities (Wallentinus 1991). If this species declines significantly or even disappears, it would affect all the macroalgal communities as well as the entire Baltic Sea ecosystem. Fucus vesiculosus is presumably very sensitive to environmental changes because it relies on sexual reproduction between separate male and female plants. The sexual reproduction is only effective over short distances and is highly susceptible to physical and chemical stress (e.g. low salinity, heavy metals, etc.). The marine seagrasses form a particularly sensitive group. Seagrasses often grow in monocultures, and the local species richness often consists of just one or very few species (Hemminga and Duarte 2000). The genetic diversity and thereby the morphological, physiological and ecological diversities are small among species, and there is a low probability that other species can take over if the dominant seagrass species disappears due to anthropogenic impacts. For example, all seagrass communities are severely affected by poor light conditions caused by mass occurrences of phytoplankton and drift macroalgae, and no seagrasses can grow drifting in the water, thereby overcoming the effect of overshadowing.

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K. Sand-Jensen and S.L. Nielsen

2.2 Plant communities classified according to organism size and habitat Plant communities are usually classified according to organism size i.e., small microalgae, large macroalgae and large angiosperms, and according to growth habitat i.e., suspended/drifting in the water column or attached to various substrata (e.g. rocks, sediments or larger macroalgae or angiosperms, Table 2). The classification according to size has three important implications: 1) Size influences the acquisition of resources (e.g. light, nutrients and inorganic carbon) and the metabolism and growth of the organisms; 2) Large size facilitates the development of plant communities with a three-dimensional architecture, influencing water motion, light availability and concentration of dissolved matter; and 3) Large size of attached species makes it possible to overgrow and overshadow smaller species. The classification according to the three main types of substrata – rock, sediment and larger plants – has important consequences regarding temporal stability and nutrient supply. Large stones and rocks are physically very stable substrata, which remain in place during storms, so that large slow-growing macroalgae can colonize, grow to large size and overshadow the smaller species. However, on stones and rocks, no fine-grained organic particles settle to decompose and release nutrients. So while the physical stability is high, the nutrient supply is often low, because it derives from the water only. Sediments are physically unstable, but contain fine-grained organic particles under decomposition, beneficial for microalgae and small opportunistic macroalgae of high nutrient requirements and fast growth. Large macroalgae and plants constitute a special substratum, because they have a limited longevity (weeks to few years) and because the host plant can inhibit the epiphytes by consuming inorganic carbon and nutrients and by producing oxygen (Borum 1985, Sand-Jensen et al. 1985). During senescence of old host tissue the release of dissolved organic matter and nutrients can be utilized by the epiphyte community (Moeller et al. 1988). The epiphytic algae are mostly fast-growing microalgae, but if the host plant lives for a long time slow-growing algae will have sufficient time to colonize and grow. A traditional classification with eight plant communities includes (Table 2): Four types of microalgae: 1. Phytoplankton living suspended in the water. 2. Epilithic microalgae growing attached to rocks and stones. 3. Epipelic and epipsammic microalgae growing in the surface layers of muddy or sandy sediments, either attached to the particles or free-living between them. 4. Epiphytic microalgae living on the surfaces of macroalgae or angiosperms. All types include a large number of species, and some species are found on more than one type of substratum. The macroalgae comprise three plant types: 5. Drift macroalgae are usually eroded from a solid substrate, but can continue to grow while drifting in shallow water. 6. Attached macroalgae live on rocks and stones. 7. Rhizoid-bearing macroalgae (e.g. stoneworts) live on sheltered soft sediments.

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Table 2. Schematic categories of aquatic photosynthetic communities according to growth habitat, taxonomy (algae vs. vascular plants) and organism size (microalgae vs. macroalgae).

GROWTH HABITAT SUSPENDED IN WATER ATTACHED TO ROCKS ETC. ATTACHED TO SEDIMENT ATTACHED TO MACROPHYTES

ALGAE MICROSCOPIC PHYTOPLANKTON EPILITHIC MICROALGAE EPIPELIC / EPIPSAMMIC MICROALGAE EPIPHYTIC MICROALGAE

MACROSCOPIC DRIFT MACROALGAE ATTACHED MACROALGAE RHIZOIDBEARING MACROALGAE EPIPHYTIC MACROALGAE

VASCULAR PLANTS

SEAGRASSES

While only a few species belong to types 5 and 7, type 6 contains many species. The vascular plants only comprise: 8. Macroscopic, rooted species, growing in soft sediments of sand, silt or mud. Vascular plants are usually defined as a particular group due to their taxonomic affiliation and their ability to root in soft sediments. Macrophyte is a common term for macroalgae and vascular plants. The classification is primarily applied because it is manageable, but it is not very precise, as nature is continuous rather than discrete. For hypothesis-testing it can be necessary to operate with several more classes (e.g. many size classes of micro- and macroalgae, different life forms and more types of substrata) or preferably measure organism size, substratum and their characteristics directly in the habitats thereby providing continuous quantitative data on their variability. Thus, linear dimensions of microalgae and macroalgae can vary 103 – 104-fold. Very fine-grained and coarsegrained sediments can vary more than 102-fold in particle diameter (< 0.1 mm to > 1 cm). Also, the 3-dimensional arrangement of sediment particles affects the degree of packing and thereby their erodibility and permeability to water, gases and microorganisms. Stability of fine-grained sediments is not a simple function of particle diameter and shear stress generated by water flow because particle size distribution, armouring of the surface with coarser bed material than below, presence of cohesive forces (e.g. between clay and organic particles) and secreted mucopolysaccharides all influence bed stability (Holland et al 1974, Gordon et al. 1992). 3. WHAT ARE THE GEOMETRIC CONSEQUENCES OF VARIATIONS IN ORGANISM SIZE AND SHAPE? The linear dimensions (L) of algae vary over seven orders of magnitude from about 1 µm in small microalgae to 10 m in large brown algae. If the shape was the same in small and in large algae, i.e. isometric shape, volume would increase over 21 orders of magnitude (L3) and the ratio between algal surface and volume (A/V ≈ L2/L3 ≈ L-1) would drop over seven orders of magnitude (Schmidt-Nielsen 1984, Niklas 1994). A drop in algal A/V-ratio this large would imply that large algae would be unable to acquire sufficient light and nutrient resources from their surroundings.

K. Sand-Jensen and S.L. Nielsen

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Indeed, photosynthetic organisms are not isometric. On the contrary, their photosynthetic tissue clearly becomes flatter with increasing size (Niklas 1994, SandJensen 2000). While small, spherical microalgae are common, large spherical microalgae or macroalgae are exceptions (Reynolds 1984, 1987). When they exist, they are not massive, but hollow with the photosynthetic tissue as a surface layer in direct contact with the surrounding water (e.g. species of Volvox and Codium). The consequences for the A/V-ratio of various sizes and shapes are illustrated in Table 3. The A/V-ratio for any given volume is smallest for a sphere, twice as large for a short cylinder (length = 10 times the diameter) and four times as large for a long cylinder (length = 100 times the diameter). A flat plate (length x width = 100 times the thickness) has an A/V-ratio twice that of a sphere of the same volume, while a very flat plate (length x width = 104 times the thickness) has the largest A/V-ratio, nearly 10 times that of the sphere. Both short and long cylindrical forms and thin and very thin thalli are common in macroalgae, and leaves of seagrasses resemble geometrically flat to very flat forms. For any shape, an increase in linear dimension by a factor of 10 will cause the volume to increase 103-fold and the A/V-ratio to decrease 10-fold. With this variation in size and shape, it is possible to maintain a constant A/V-ratio with a ten-fold increase in linear dimension and 103-fold increase in volume by shifting from a spherical shape to a very thin plate. With an increase in linear dimension of 100, the A/V-ratio decreases 100-fold, and then it is no longer possible to maintain an unaltered A/Vratio by changing shape (Table 3). Over the variation in linear dimension by seven orders of magnitude from the smallest microalgae to the largest macroalgae, the A/V-ratio will therefore decrease significantly because large algae need to have a reasonable thickness to attain a sufficient strength and prevent them from being torn into pieces by the strong drag forces impinging on their large surfaces (Vogel 1994). Table 3. Surface area: volume ratio (A/V) and flatness index (A3/V2) for typical shapes of aquatic photosynthetic organisms. Spheres: thallus thickness, T = diameter. Short cylinders: length, L = 10 times thallus thickness, T (= cylinder diameter). Long cylinder: L = 100 x T. Thin thallus: length, L times width, W = 100 times thallus thickness, T. Very thin thallus: L x W = 10,000 T. Absolute A/Vratios also shown for three representative volumes (1, 106 and 1012 µm3).

General

Flatness

A/V-ratio for different volumes

A/V-ratio

Index

1 µm3

106 µm3

1012 µm3

Sphere

6 T-1

113

4.84

4.84 x 10-2

4.84 x 10-4

Cylinder, short

∼ 4 T-1

502

7.95

7.95 x 10-2

7.95 x 10-4

Cylinder, long

∼ 4 T-1

5024

17.14

17.14 x 10-2

17.14 x 10-4

Flat thallus

∼2 T-1

800

9.30

9.30 x 10-2

9.30 x 10-4

Very flat thallus

∼ 2 T-1

80000

43.06

43.06 x 10-2

43.06 x 10-4

Shape

The variability from small microalgae to large macroalgae can be illustrated with an example. A spherical microalga with a volume of 1 µm3 has an A/V-ratio of 4.8 µm-1. A 3 m long, 0.3 m wide and 3 mm thick macroalga, corresponding in shape to a very

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thin plate, has a volume of 3 x 1015 µm3 and an A/V-ratio of 6.7 x 10-4 µm-1; 104 times smaller than the small sphere. Therefore the macroalgae will unavoidably have a less favourable A/V-ratio for uptake of limiting nutrients from the environment than the microalgae. This limitation will be strengthened, because thicker diffusive boundary layers exist over the surfaces of macroalgae than microalgae. The A/V-ratio has the dimension L-1, showing its inverse dependence on organism size, which makes it difficult to compare species of variable shape and size (Table 3). To compare flatness independent of size, it is necessary to use a dimensionless index such as A3/V2 (or A3/2/V, Niklas 1994). This flatness index increases in the order: sphere < short cylinder < thin plate < long cylinder < very thin plate for the shapes listed in Table 3. The precise value of the index obviously varies with the ratio between length and thickness of cylinders and plates, while it is constant in spheres (36 π) regardless of volume. With increasing volume it is possible to increase the flatness of cylinders and plates if they become relatively thinner (Table 3). In the example, the flatness index was 500 in the short cylinder and 5000 in the long cylinder. For any given cell volume, the surface area was twice as large in the long as in the short cylindrical form. This explains a general pattern, observed in the internodal cells in Chara corallina, which become relatively thinner relative to their length with increasing volume (Niklas 1994). 4. WHAT IS THE IMPORTANCE OF ORGANISM SIZE AND SUSPENDED/ATTACHED LIFE FORM FOR WATER MOTION AND SOLUTE TRANSPORT? Water motion around plant surfaces is of paramount importance for the exchange of matter between plants and the surrounding water and therefore for the risk of transport limitation of their metabolism and growth. In this regard both organism size and suspended or attached life form are important; size because it affects thickness of the diffusive boundary layers, and life form because suspended microorganisms experience a turbulent environment while attached microorganisms are buried under the viscous and diffusive boundary layers generated by the solid substratum. Attached macrophytes protrude into the turbulent layers above the substratum, but are surrounded by thick diffusive boundary layers over their surfaces. 4.1 The significance of microalgal size and mobility Microorganisms live in a world with low Reynolds numbers dominated by viscous forces. They also live in a microworld, where diffusion occurs so rapidly that it exceeds the advective mass transport of matter. While turbulent flow is important on larger scales (> few mm), the vortices are dampened by viscosity at small scales and disappear below the 1-mm scale. So water motion is simpler and more predictable on the microscale, where water sticks to any firm surface, and parcels of water flow in laminar and well-ordered pattens around surfaces (creeping flow, Vogel 1994). Transport of dissolved substances from the water to the surface of organisms is by molecular diffusion, resulting in clear relationships between metabolism, organism size and substrate concentration in the water (C). If nutrient uptake (the flux, J) in a spherical microalga (diameter, d) is limited only by diffusion (diffusion coefficient,

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K. Sand-Jensen and S.L. Nielsen

D), and the organism is able to lower substrate concentration to zero at the cell surface, the flux is given by:

J = 2π DCd

(1)

(Karp-Boss et al. 1996). The flux is proportional to the linear dimension of the cell, the external substrate concentration and the diffusion coefficient. But the volume of the cell is proportional to the linear dimension cubed, so the volume-specific nutrient uptake (E) will decrease with the diameter squared according to:

E = 12 DCd −2

(2)

This relationship implies that bacteria and small microalgae are not limited by diffusion, but are able to maintain an extremely efficient substrate uptake even at very low external substrate concentrations. Every second they can take up small dissolved molecules from a volume of water that is several hundred to thousands times larger than their cell volume (Fenchel et al. 1998). A microalga with a diameter of 1 µm, that effectively can reduce the concentration of dissolved phosphate to zero at the cell surface, can saturate its uptake at external nutrient concentrations below 0.1 mg P m-3 (Sand-Jensen 2000). A large microalga of 10 µm in diameter requires a higher external P-concentration of about 1 mg P m-3 to ensure an optimal P-supply by diffusion to saturate growth. This estimate supports the hypothesis that very small phytoplankton species can achieve a maximal growth rate in very oligotrophic oceans, whereas the total pool of nutrients limits algal biomass (Raven 1999). However, matters may be more complicated because the cells may not be able to reduce the solute concentration to zero at the cell surface and the largest resistance to nutrient utilization may not be diffusive transport, but can alternatively be the transport across the cell surface or the internal conversion of inorganic cellular Ppools to organic P-compounds for cell growth. For bacteria and small microalgae diffusion is so efficient that the transport of dissolved substances to and from the cell is unaffected by cell movement relative to the surrounding water by swimming or sinking (Lazier and Mann 1989, Kiørboe 1993, Karp-Boss et al. 1996). However, for larger cells in the size range 40 – 100 µm or larger the total transport of dissolved substances will be increased by advective transport by swimming or sinking. Most microalgae also benefit from active swimming because it permits them to seek out optimal microenvironments with respect to light, nutrients or grazing, or allows them to avoid sinking out of the photic zone to deeper water or to the sediment. The surface sediments have steep gradients in light, nutrients, oxygen and pH in the top few millimetres (Revsbech and Jørgensen 1986). It can therefore be a great advantage for microscopic bacteria, algae and animals to be able to orient themselves in the gradients and seek out the best layers. As the swimming speed of microorganisms, as “a rule of thumb”, is about 10 times their length per second (Denny 1993), it will only take them from seconds to minutes to move to the preferred layers, while the movement itself only has significance for exchange of matter for the largest organisms and for those substances that constrain metabolism and growth.

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4.2 The significance of attachment for microalgae Firm surfaces create their own boundary layer with increased viscosity and transport resistance for exchange of matter between the substratum and the surrounding water. According to the theory of diffusion, the vertical component of the eddy diffusivity (Ez) decreases with an exponent of 3 – 4 with the height over the substratum and proportional to the kinematic viscosity (ν) according to:

Ez = kν z 3 − 4

(3)

(Boudreau 2001, Dade et al. 2001). In a thin layer over the substrate (typically 0.5 – 2 cm) the vertical eddy diffusivity is smaller than the kinematic viscosity (approx. 1 – 2 x 10-6 m2 s-1) and water movement is dominated by viscous forces in the so-called viscous sublayer (VSL). Even closer to the surface (typically 0.2 – 2 mm) eddy diffusivity decreases to less than the diffusion coefficient (D, approx. 0.5 – 2 x 10-9 m2 s-1 for small molecules) in the diffusive boundary layer (DBL), where diffusion is the fastest mode of transportation (Jørgensen 2001). The outer limits of VSL and DBL are not fixed, as they undulate due to transmission of pressure waves or particularly strong turbulent vortices through the layers towards the firm substratum (Gundersen and Jørgensen 1990). The thickness of the diffusive boundary layer depends on the intensity of turbulence, the roughness of the substrate, the exact position on the surface and on the presence of animals, creating their own current (Jørgensen 2001). Increased intensity of turbulence is often coupled to higher macrocurrent velocity which decreases the thickness of DBL. In flow aquaria in the laboratory, the thickness of DBL decreased over a sediment surface from 560 µm at 0.3 cm s-1 to 140 µm at 7.7 cm s-1 (Jørgensen and Des Marais 1990) and over an eelgrass leaf DBL decreased from 470 µm at 0.22 cm s-1 to 160 µm at 11.2 cm s-1 (Larsen, unpubl.) An increased macroturbulence, especially at low current velocity will also reduce the thickness of DBL, even though the mean velocity remains constant. Increased roughness of the substratum surface will usually decrease the thickness of DBL by inducing microturbulence closer to the surface, but in addition the increased roughness will increase the spatial variation in DBL. The DBL is very thin on exposed sides of protruding objects facing the current, but thicker on the lee side of the objects (Jørgensen and Des Marais 1990). Over a rough sediment surface, the DBL will be thin over the top of the sediment grains, but thick in the depressions between the grains. These variations in the thickness of DBL are very important for the exchange of substances between the substratum and the water, because the flux is inversely proportional to the thickness of DBL. A rough sediment surface will also be able to increase the total transport of matter between the water and the sediment, because it effectively increases the surface area through which diffusion can take place (Gundersen and Jørgensen 1990) Finally, animals that move or pump water for food ingestion and respiration, will be able to decrease the thickness of DBL and the surface area for diffusion across animal tubes and burrow structures, thus increasing the exchange of dissolved substances (Aller 2001). Because DBL is defined as the layer where the eddy diffusion coefficient for mass decreases below that of the diffusion coefficient, the thickness of DBL will change with the diffusion coefficient of the given dissolved gas or solute molecule being

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K. Sand-Jensen and S.L. Nielsen

thinner for large, slowly diffusing molecules and thicker for small, rapidly diffusing molecules and also thicker at increasing temperatures (Jørgensen 2001). Since the thickness of DBL also depends on the eddy diffusivity close to the substratum, it should change with the third to fourth root of the diffusion coefficient. Thus, a tenfold variation in the diffusion coefficient from protons to small organic molecules should result in an approximate two-fold variation in the thickness of DBL. Most DBL-thicknesses have been measured by oxygen microelectrodes which will yield values close to those for nitrate and ammonium of similar diffusion coefficients but slightly higher than those for small organic molecules. For oxygen the diffusion coefficient approximately doubles from 0 to 20 °C and the kinematic viscosity decreases about 1.7-fold resulting, in theory, in a 40% thicker DBL (Jørgensen 2001). The temperature dependence of DBL-thickness is about the same for other small molecules as for oxygen. The presence of VSL and DBL will impede the uptake of nutrients and inorganic carbon for microalgal metabolism from the water and likewise impede the removal of waste products. On the other hand, the impeded exchange will facilitate the retention and recirculation of important nutrients in the surface layers Reuter et al. 1980). Sediments often receive a significant input of small organic particles, creating very high nutrient concentrations in the sediment pore-water through mineralization. Nutrient concentrations are often 100 – 1000 times higher in the sediment than in the overlying water, and a prerequisite for maintaining this concentration gradient is the presence of VSL and DBL over the sediment, so that the nutrients are only slowly released to the water. The benthic microalgae are situated on top of a large and concentrated nutrient pool, whereas the phytoplankton live in a diluted environment, where nutrient concentrations are small and often limiting for biomass development. While the presence of VSL and DBL can be beneficial for the supply of nutrients to microalgae in the sediment, because they facilitate the retention of nutrients that are released by remineralization in the sediment, the situation is different for microalgae living attached to inactive substrata such as stones or rocks, or other very coarse substrata with no significant sedimentation of nutrient-rich, organic or inorganic particles (Fairchild and Everett 1988). Here, the nutrients need to be taken up from the water and the thickness of VSL and DBL can constrain the nutrient flux. On sandy sediments, for example, growth and biomass accrual of microalgae can be nutrient limited because sandy sediments are frequently percolated by the water or rebedded by currents due to their physical instability. 4.3 The significance of the flow environment for macrophytes The flow environment and nutrient supply are rather complicated for macrophytes (macroalgae and seagrasses), because they are large, form dense populations and grow attached to stones, rocks or soft sediments (Jumars et al. 2001). As already mentioned, the solid surfaces create a benthic boundary layer with decreased turbulence and mass transport of substances. The individual macrophyte also creates its own boundary layer, the thickness of which depends on the surrounding flow environment and the position on the macrophyte surface (Hurd et al. 1997). The boundary layers are variable due to a variable macro-flow and movement of the macrophyte in the current (Ackerman and Okubo 1993). Increasing macroflow leads to enhanced nutrient uptake (Gerard 1987, Hurd et al. 1996). However, many of the

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details are not known. The complexity increases because the macrophytes form dense populations, reduce current velocity and rescale the turbulence. It is far from clear if the intensity of turbulence is decreased, enhanced or remains unchanged (Jumars et al. 2001). That depends partly on the turbulence in the free water masses, the type and density of the macrophyte population and to a very large extent on the distances from the lateral edge of the population and the canopy surface to the measurement position within the canopy (Gambi et al. 1990, Sand-Jensen and Mebus 1996). Far from the lateral edge and canopy surface the macro-current velocity and the intensity of turbulence will be significantly dampened, but at the very edge and canopy surface the turbulence can be significantly increased where currents and waves suddenly encounter an increased resistance (Sand-Jensen and Pedersen 1999). 5. WHAT IS THE SIGNIFICANCE OF PLANT SIZE AND SHAPE FOR LIGHT UTILIZATION, NUTRIENT UPTAKE AND GROWTH? Photosynthetic organisms with thin thalli have a shorter pathway for light through the tissue, high A/V-ratios and, in general, a more efficient light utilization and nutrient uptake per biomass unit than organisms with thick thalli (Kirk 1994, Niklas 1994). Thin photosynthetic tissues also contain higher pigment concentrations and enzyme capacity per cell volume. These differences yield the potential for a higher metabolism and a faster growth in small, thin species than in large, thick species (Nielsen and Sand-Jensen 1990, Duarte 1995, Duarte et al. 1995). Thus, the geometric differences have significant eco-physiological implications. This is not the whole story, however. Evolutionary aspects are involved as well, when small thin species are adapted (r-selected) to have an efficient resource utilization and to be able to photosynthesize, respire and grow very fast (Littler 1980, Reynolds 1987). These adaptations have a cost in the form of large expenses to maintain the tissue and large losses to grazing and senescence. The thick species are adapted (Kselected ≈ C-selected, Grime 1979) to a lower resource utilization, metabolism and growth, but on the other hand have a longer longevity of the tissue due to smaller maintenance expenses, small grazing losses and slow senescence. There are clear trade-offs between resource utilization and growth on one side and grazing, senescence and longevity on the other (Cebrián and Duarte 1994, 1995). The adaptive element also means that some organisms (small as well as large), are adapted (S-selected sensu Grime 1979) to grow and survive in extremely resourcepoor environments, lacking light or nutrients. Crust-forming macroalgae are an example (Littler and Littler 1985). For these S-selected organisms it is particularly important to protect themselves against losses, retain nutrient resources and survive (Markager and Sand-Jensen 1994). S-selected organisms will therefore deviate from the general relationships describing resource utilization, metabolism and growth of rand K-selected species to organism size. 5.1 Physiological rates scaled to tissue thickness or A/V-ratio To scale resource utilization and metabolism to size and shape of the photosynthetic tissue the thickness and the A/V-ratio are often used. Tissue thickness is intuitively easy to understand, but has the inherent weakness that there is no unequivocal relation to the A/V-ratio. The relationship between tissue thickness and A/V-ratio

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varies two to three-fold between spherical, cylindrical and plate shape (Table 3). Using the A/V-ratio as a scaling parameter has the advantage that it describes the exposed surface through which external resources are received, relative to the volume in which the resources are consumed. However, volume can be more or less densely packed with organic matter, causing variations in mass density and carbon concentration. It is therefore better to use carbon content per tissue volume or tissue surface area as a scaling parameter (e.g. Markager and Sand-Jensen 1996) also because photosynthesis and respiration measured in carbon units per unit time can then be expressed in relation to carbon biomass (i.e. carbon (cell carbon)-1 time-1) resulting in relative rates per unit time analogous to the relative growth rate. Surface area/cell carbon content has therefore been recommended as a scaling parameter (e.g. Markager and Sand-Jensen 1996), but as cell carbon content is often unknown, tissue thickness or A/V-ratios are often used instead. 5.2 Light absorption as a function of size and shape Irradiance decreases exponentially with distance through an optically homogenous material (Kirk 1994). The absorbance of the tissue (Abs) is a product of the distance that light has to travel (L) and the light extinction coefficient (η) given as:

Abs = 0.434 Lη

(4)

where the constant 0.434 derives from expressing absorption in log10-, rather than lnunits. The cells in thick tissues therefore on average receive a lower irradiance, relative to the incoming irradiance, than cells in thin tissues as a consequence of self-shading. This effect can partly be counteracted by concentrating the strongly pigmented photosynthetic cells close to the surface of the organism, and letting cells deeper in the tissue function in maintaining strength, flexibility, transport and storage. The light path is very short through small microalgae. Only a small fraction of the incoming irradiance (< 10 %) is absorbed during passage, but the efficiency of absorption is large when expressed as the number of absorbed photons per amount of pigment or per unit length of the path. It is therefore advantageous for small microalgae to increase pigment concentration in low light (Agusti et al. 1994). The investment in higher pigment concentration – expressed as the number of photons that it costs to produce the pigments – is much lower than the extra photons absorbed by the additional pigments (Markager and Sand-Jensen 1994). However, if the irradiance is very low, this advantage disappears, and pigment concentration decreases (Sand-Jensen 1988a,b). Thus, pigment concentration peaks at a low to medium irradiance (Sand-Jensen 1988a). An unusually large variation in pigment concentration (up to ten times) is indeed observed in microalgae as an adaptation to various light and nutrient conditions (Agusti et al. 1994). This variation leads to very large differences in light absorption (up to six times) and almost proportional changes in photosynthesis at low light.

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Photon absorptance (%)

100

75

50

25

0 0

200

400

600

Chlorophyll density (mg

800

cm -2)

Fig. 1. Photon absorptance versus areal chlorophyll concentration for phytoplankton, macroalgae and leaves. Data compiled from many sources by Agusti et al. (1994) and redrawn. 3

Photon absorbance (mg -1 DW)

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-3

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Fig. 2. Photon absorbance per unit of tissue DW versus tissue thickness of phytoplankton, macroalgae and leaves. Data compiled from many sources by Agusti et al. (1994 and redrawn).

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Chlorophyll concentration (mg g-1 DW)

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-3

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Fig. 3. Chlorophyll concentration versus tissue thickness of microalge, macroalgae and leaves. Data compiled from may sources by Agusti et al. (1994).

For large unicellular algae, colonial algae, multicellular algal thalli and leaves of vascular plants the situation is different (Frost-Christensen and Sand-Jensen 1992). With an increasing thickness of the photosynthetic tissues and larger total amount of pigments along the light path, the light absorptance increases asymptotically towards 100 % (Fig. 1). In large unicellular algae, colonial algae and thalli with 1 – 4 cell layers, variations in pigment concentration still affect light absorptance. These organisms are located on the curved section in Fig. 1, where significant variations in pigment concentration can change absorptance up to three-fold, and where the organisms can adapt to variations in irradiance by changing their absorptance. An example: In the two-layered and 70 – 120 µm thick thallus of the green alga Ulva lactuca a weakly pigmented thallus typically absorbs 25 % of the incoming irradiance, while a strongly pigmented thallus with eight times as much chlorophyll absorbs up to 75 % of the incoming irradiance (Markager and Sand-Jensen 1994). Thick macroalgae and leaves of vascular plants almost always absorb between 80 and 98 % of the incoming light and the variation in pigmentation has little effect on light absorption (Frost-Christensen and Sand-Jensen 1992). The pigment content therefore varies little in thick tissues for the purpose of optimizing light absorption and photosynthesis in low light. However, increased pigment content has the advantage that most of the incoming irradiance is absorbed by pigments and therefore can be utilized in photosynthesis, while a very low fraction is absorbed by cell walls and structural compounds (Markager 1993). A high pigment content also means that the photons and their excitation energy are distributed among a larger number of chlorophyll molecules and associated reaction centres and electron transport chains in the photosystems. Hereby, a larger fraction of the photons can be utilized, especially in high light, the light saturation point will increase and the risk of photoinhibition

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decreases. A given amount of pigments is therefore more efficient in absorbing light in thin than in thick tissues, and the efficiency decreases with higher pigment concentration for any given shape and thickness (Fig. 2). This relationship is an example of the Law of Diminishing Returns. It is therefore anticipated that small microalgae have the highest pigment concentrations, while thicker multicellular forms have lower pigment concentrations (Fig. 3). The three figures (1-3) together give a basis for understanding the systematic patterns of light adaptation, photosynthesis and growth in plants. The much larger yield of absorbed photons per investment in pigments in thin unicells compared to thick multicellular organisms contributes to the higher photosynthetic rate per unit volume and higher growth rates of the smallest organisms (Nielsen and Sand-Jensen 1990, Nielsen et al. 1996). 5.3 Nutrient uptake as a function of size and shape Nutrient uptake is strongly dependent on size and thallus thickness in micro- and macroalgae (Smith and Kalff 1982, Hein et al. 1995). Median values for maximal uptake rate of ammonium per biomass are five-fold higher for unicellular phytoplankton than for thin and sheet-like macroalgae and fifty-fold higher than for thick blades and coarsely branched macroalgae (Hein et al. 1995). Over this size range the A/V-ratio varies almost thirty-fold. Within the various types of algae, the nutrient uptake rates also vary markedly, but a significant proportion of this variation can be accounted for by variations in the A/V-ratio among species. Table 4. Scaling statistics of Vmax (µg N g-1 DW h-1), Km (µg N l-1) and α-values (µg N g-1 DW h-1 (µg N l-1)-1) describing the double-log relationships of nitrate and ammonium uptake in micro- and macroalgae as a function of the A/V-ratio (cm-1). Data compiled from many sources by Hein et al. (1995).

Over the total size range of micro- and macroalgae there are significant relationships between maximal uptake rate (Vmax), half-saturation constant (Km) and affinity for uptake at low concentrations (α = Vmax/Km) for both ammonium and nitrate (Fig. 4). Using the power-law (Y = aXb), Vmax is scaled to A/V-ratio with a mean exponent of 0.61 for ammonium and 0.66 for nitrate, while α is scaled to A/V-ratio with an exponent of 1.01 for ammonium and 1.16 for nitrate (Table 4).

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Fig. 4. Allometric scaling of the kinetic parameters describing ammonium (left panels) and nitrate (right panels) uptake of phytoplankton (closed symbols) and macroalgae (open symbols) versus their surface area: volume ratio (A/V). Maximum uptake rate (Vmax, upper panel), apparent half-saturation constant (Ks, mid panel) and affinity (Vmax/Km, lower panel) changed linearly with A/V in log-log relationships. Data compiled from many sources by Hein et al. (1995).

The greater stimulation of α-values by small algal size may derive from the thinner diffusive boundary layers of small cells. The smaller size-dependence of Vmax-values may also be accounted for if larger algae have higher areal density of membrane transporters or better ability to supply them with energy and transfer the absorbed nutrients to deeper layers. For isometric forms the scaling of nutrient uptake to the A/V-ratio with an exponent of 1.0 would mean a direct inverse proportionality with the linear dimension. The size dependence is markedly smaller than the inverse relationship to the diameter squared, previously calculated for spherical organisms (eq. 2), whose nutrient supply is limited by diffusion alone and that are capable of lowering the substrate concentration at the cell surface to zero. The affinity for nutrient uptake (α = Vmax/Km) is dependent on both Vmax and Km. With an increased A/V-ratio, Vmax increases and Km decreases. While Vmax increases about fifty-fold from large macroalgae to phytoplankton, Km only decreases about

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ten-fold. The variation in α-values is therefore due more to variations in Vmax than in Km. While changes in A/V-ratios among species can account for 53 – 72 % of the variation in Vmax, they can only account for 33 – 40 % of the variation in Km-values. For all nutrient kinetic parameters, there is a very large unexplained variation. A number of parameters besides the A/V-ratio can affect kinetics. The most obvious ones are nutrient demand, physiological state and flow environment. Species with a high nutrient demand and a high physiological capacity for uptake are expected to display a markedly higher uptake capacity than nutrient saturated species and species in which the physiological capacity for uptake has declined for some reason (Hein et al. 1995). The relative ability to utilize either ammonium or nitrate will also contribute to variations in the kinetic constants. Finally, variations in the flow environment and in the thickness of diffusive boundary layers for the largest species will affect especially Km- and α-values. Even though small algae capture nitrogen more efficiently per unit biomass from the surrounding water than large algae this does not imply that small algae are competitively superior under nutrient limitation (Fig. 4 and Table 4). There are several additional complexities in the nutrient budget – nutrient requirements, storage capacity and ability to recycle nutrients – that need to be evaluated (Table 5). Table 5. Nutrient economy of small, fast-growing (r-selected) macroalgae and large, slowgrowing (K-selected) macroalgae according to nutrient uptake, nutrient content, nutrient requirement and recirculation capacity. Adapted from Pedersen (1993).

Nutrient demand Nutrient uptake rate Storage capacity Recycling ability

r-selected algae High High Low Absent

K-selected algae Low Low High Present

Nitrogen requirement to sustain growth is the product of attainable growth rate and “critical” tissue N- concentration (sensu Hansiak 1979). Maximum growth rate (µmax) varies across the range from microalgae to macroalgae with a mean allometric slope of 0.66 in the log µmax – log A/V relationship (Nielsen and Sand-Jensen 1990). The hypothetical slope of algal nitrogen requirement versus A/V should, however, show a greater size-dependence because average N-concentrations are considerably higher for microalgae (1.0-14.0% of DW) than for macroalgae (0.4-4.4% of DW, Duarte 1992). There are not sufficient data for the “critical” nutrient concentration to establish the scaling of nitrogen requirements to algal size, but the approximately five-fold lower mean nitrogen concentration of macroalgae, together with their lower growth rates, will lead to a closer balance between nutrient uptake at low substrate concentrations and nutrient requirements for growth of large algae. The gradually shorter duration of nutrient limited growth of macroalgae with increasing size and lower maximum growth capacity demonstrated in a Danish estuary during the nutrient-depleted summer period supports that notion (Fig. 5).

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36

Large, slow-growing algae may have the additional ability to store more nutrients relative to the time-specific demand during growth compared with small, fastgrowing algae (Table 5). This storage ability is obvious in large, differentiated macroalgae which have certain tissue allocated to photosynthesis and other nonphotosynthetic tissue allocated to storage of carbohydrates and nutrients, solute transport, anchorage and mechanical strength. Being perennial large macroalgae also have the ability to take up and store inorganic nutrients during seasons (often the winter) in which nutrients are available in high external concentration but the requirements are low due to slow growth (e.g. low light availability, Chapman and Cragie 1977, Zimmerman and Kremer 1986). The stored nutrients can be used later when nutrient requirements are high but external availabilities are low (often the summer).

Incident light (µmol m -2 s -1 )

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15 N

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Increasing growth rate

Phytoplankton

Ulva Cladophora Ceramium Chaetomorpha Fucus

M

A

M

J

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O

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Fig. 5. Duration of N-limitation of the growth rate of macrophyte species and of the biomass accumulation of mixed phytoplankton communities in a Danish estuary. Duration of Nlimitation increases with increasing maximum growth rate of the species. From Pedersen (1995).

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Recirculation of nutrients within the tissue of large macroalgae and seagrasses is a final strategy by which the organisms are capable of reducing losses by withdrawing nutrients from old, decaying tissue and transferring them to new tissue in active growth. The critical aspect regarding the efficiency of nutrient circulation is the capability to reduce the tissue concentration in the old, decaying tissue relative to the mean concentration in the younger, healthy tissue. Therefore, nutrients pools that are allocated to structural tissue cannot be reclaimed and reused in new production to the same extent as nutrients that are allocated to catalytic purposes. Studies of eelgrass have shown that recirculation could reduced the external nitrogen requirement by 2530% on an annual basis, but by more than 50% during critical periods in early summer when growth rates and biomass accumulation were highest and external nutrient concentrations were lowest (Hemminga et al. 1991, Pedersen and Borum 1993). A compilation of data on several seagrass species shows an average resorption of 20-22 % nitrogen and phosphorus (Hemminga et al. 1999) which is much less than the 41-74% resorption by perennial terrestrial plants (Aerts 1996). Although nutrient resorption is apparently not a pronounced conservation strategy in seagrasses (Hemminga and Duarte 2000) compared to terrestrial plants it does help the plants to sustain a higher and more permanent biomass and organic production throughout the year even in nutrient-poor environments. 5.4 Growth and metabolic rates as a function of organism size and shape Rates of photosynthesis, respiration and growth of algae and vascular plants increase as the photosynthetic tissues become thinner. At the same time the cellular concentrations of pigments, nitrogen and phosphorus increase. For photosynthetic tissue it is possible by comparing the metabolic rate with tissue thickness and with concentrations of pigments, nutrients and enzymes to establish relatively close relationships (Nielsen et al. 1996). For unicellular algae the comparison of growth rate to size and tissue thickness is straightforward because they consist entirely of photosynthetic tissue. Such comparisons have shown a highly significant log-log linear relationship between the growth rate (log y) and cellular carbon content (log x) with a mean slope of about -0.25 (Geider et al. 1986) similar to the size relationship of animal growth to body weight (Fenchel 1974). Some systematic differences apparently exist among different algal groups (e.g. diatoms grow faster than dinoflagellates for the same cell weight, Banse 1982) resulting in residual variation along the regression line, but without disturbing the overall picture. The relationship of animal and plant growth to organism size is remarkable as it basically follows the same overall statistical pattern with scaling exponents resembling each other (Hemmingsen 1960, Enquist et al. 1998, Sand-Jensen 2000). For macroalgae and vascular plants the comparison of growth rate to organism size is, however, more difficult because the organisms do not have a well-defined size as they grow larger with age. Moreover, macroalgae and particularly vascular plants are differentiated in photosynthetic and non-photosynthetic tissue and only the photosynthetic tissues are directly involved in the production of new organic substances, while the non-photosynthetic tissue is a sink for photosynthates and serves other important purposes (e.g., nutrient uptake, storage, anchorage, etc.).

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Growth models constructed for terrestrial plants have been able to encompass the variability of tissue differentiation and weight proportions of photosynthetic tissue to total plant weight (Poorter and Remkes 1990, Lambers and Poorter 1992) but they have not so far been tested on the entire suite of aquatic microalgae, macroalgae and vascular plants. For 35 species of non-rhizoid macroalgae from shallow coastal Danish waters the maximum in situ growth rates were shown to vary from 0.02-0.03 day-1 for thick leathery brown algae (e.g. fucoid species) to 0.3-0.4 day-1 for thin sheets and thin filamentous forms of green and brown algae (e.g. species of Ulva, Cladophora and Pilayella, Nielsen and Sand-Jensen, Markager and Sand-Jensen 1994, 1996). The A/V-ratio ranges from about 20 mm-1 in the thick forms to 2000 mm-1 in the thin forms. The relationship of maximum growth rate to A/V-ratio is log-log linear with a mean slope 0.79 (±0.20, 95% C.L.; Table 6). This scaling exponent can be compared with that of growth rate versus cell weight of unicellular algae at about -0.25 (Geider et al. 1986) by assuming, somewhat optimistically, that the macroalgae are isometric and that their mass (M) relative to volume (V) remains constant. Table 6. Scaling of rates of growth, photosynthesis and respiration (log y) to thallus A/V-ratio (log x) of 35 marine macroalgae from Danish coastal waters in May-June. The table shows the scaling exponent (b) and the intercept of the relationship: log y = a + log x. All rates in mmol C (mol cell C)-1 day-1 and A/V-ratios in cm-1. From Markager and Sand-Jensen (1994), and unpublished data.

Scaling relationship Growth rate vs. A/V Photosynthetic rate vs. A/V Respiration rate vs. A/V

a 0.13 1.16 0.40

b ± 95% C.L. 0.79 ± 0.20 0.62 ± 0.08 0.55 ± 0.10

These assumptions would imply that if the growth rate scales with A/V with an exponent 0.79, then it also scales with (L-1)0.79, (V-1/3 )0.79, V-0.26 and M-0.26, a result that is very close to the size relationships observed for unicellular organisms (Fenchel 1974, Geider et al. 1986). Rates of photosynthesis and respiration have also been studied as log-log functions of the A/V-ratios for the same 35 species of marine macroalgae (Markager and SandJensen 1994). Scaling exponents averaged 0.62 (±0.08) for photosynthesis and 0.55 (±0.10) for respiration and were significantly higher than 0 and lower than 1.0 (Table 6). The smaller size dependence of respiration than photosynthesis is perhaps due to the fact that respiration comprises both growth respiration, which is coupled to the formation of new substrates for growth and is proportional to the growth rate, and maintenance respiration, which is coupled to maintenance of the structural and catalytic machinery of the organism. At very high growth rates in small and thin organisms the majority of total respiration will be allocated to growth respiration and less to maintenance, while at very low growth rates in large and thick organisms a large proportion of total respiration will be allocated to maintenance. This would explain why respiration increases less steeply than photosynthesis with increasing A/V-ratio and, accordingly, that the growth rate increases even faster than photosynthesis.

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A broad-scale comparison of growth rates versus mean A/V-ratios of photosynthetic tissue has been made across the entire range of aquatic photosynthetic organisms from microalgae, over macroalgae to vascular macrophytes (Fig. 6, Nielsen and Sand-Jensen 1990). The log-log relationship follows a common line with a slope of 0.66 and it accounts for 86% of the variation in growth rates among the different species, sizes and forms. Although the relationship is highly significant and can account for much of the variability across this enormous size range, it is not very accurate for predicting the actual growth rate for a selected species with a certain A/V-ratio. The maximum inter-specific variability in maximum growth rate among species is substantial for a given A/V-ratio. This inter-specific variability would even be higher if the particularly slow-growing, stress-selected species from chronically shaded or nutrient-poor habitats had been included in the analysis. Such stressselected species, whose functional performance is poorly known, are expected to fall systematically below the common line (i.e. at lower growth rates for a given A/Vratio) which mainly describes the variation from large, C-selected (= K-selected) species to small r-selected species.

Growth rate (doublings per day)

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-1

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Relativ Surface (m2 m-3)

Fig. 6. Maximum growth rate versus surface area: volume ratio (A/V) of the photosynthetic tissue of phytoplankton, macroalgae and vascular macrophytes. Maximum growth rate increases log-log- linearly with A/V with the slope 0.62 (±0.08, 95% CL). Data compiled from many sources by Nielsen and Sand-Jensen (1990) and redrawn.

5.5 How resource acquisition dictates density and biomass in plant communities In dense communities of phytoplankton, macroalgae and vascular plants there is a common upper limit to areal gross photosynthesis (see next section). With such a common upper limit to total metabolism of the community and a distinct scaling of 0.25 (-1/4) for mass-specific metabolism of the individual organism relative to body mass (Geider et al. 1986) there are two important implications. The first implication is that the total biomass in the community should scale with mean individual mass with the exponent 0.25. It is easy to accept that if large organisms per unit weight

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consume less resources than small organisms, and if the same total amount of resources (e.g. light) are available in the communities irrespective of body mass, then communities composed of large organisms can support much larger biomasses. The second implication is that mean individual mass scales with maximum density with the negative exponent of -1.33 (-4/3) providing a correction based on first principles and energetic arguments to the traditionally accepted exponent of -1.5 (-3/2) for the self-thinning law. The mathematical arguments are presented by West et al. (1997) and Enquist et al. (1998) and are straightforward. Particularly the first implication is important and supported by data for various aquatic primary producers. It has been shown that the maximum biomass of microalgae under standard culture conditions is much higher for large species than small species (Agusti and Kalff 1989). The scaling exponent of algal biomass to algal size was 0.21 (±0.06) and it was significantly larger than zero and not significantly different from the exponent of 0.25 (Enquist et al. 1998). Higher algal biomass is also observed for large than for small algal species under natural bloom conditions at high light and high nutrient availabilities (Agustí et al. 1987). Likewise, maximum biomasses of macroalgae and seagrasses are higher when large species rather than small species dominate the communities. For a 104-fold difference in mean organism size a 10-fold larger biomass can be predicted for the larger species relative to the smaller species. This is the approximate difference in biomass observed between communities of small green or brown macroalgae (e.g. Cladophora and Pilayella) and communities of large brown macroalgae (e.g. Fucus and Laminaria).

6. WHAT ARE THE BASIC FUNCTIONAL PROPERTIES AND INTERRELATIONS OF DIFFERENT PHOTOSYNTHETIC COMMUNITIES? In most coastal waters, the phytoplankton community plays an important and often a dominant role in primary production (Borum and Sand-Jensen 1996). Only in tidal areas with a variable water stage does the importance of the phytoplankton community for light attenuation and total primary production vary over the daily cycle. If the water is very shallow it will restrict the ability of the phytoplankton community to utilize the available light. If the water is exchanged very rapidly because of large freshwater inputs the phytoplankton community will not have sufficient time to grow and develop a biomass that can contribute substantially to total primary productivity. In contrast, benthic microalgae, attached macroalgae and rooted plants vary extensively from place to place depending on the presence of suitable substrata, water depth, light intensity reaching the bottom and physical perturbation (Sand-Jensen and Borum 1991). If the substratum is very unstable there are no macroalgae and rooted plants, and if only rocky substrata are present the attached macroalgae can grow but with a few exceptions not the rooted plants. If strongly attenuated irradiances reach the sea bottom because of deep and/or turbid water then benthic primary producers will be absent all together. In contrast, if high irradiances reach the bottom, as is the case in shallow and/or oligotrophic transparent waters, then the primary production of benthic microalgae and rooted plants can be dominant.

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Phytoplankton communities are, however, in many ways superior in the competition for light among the primary producers because they are closest to the light source and can shade all the benthic algae and plants except for those growing in very shallow water (Sand-Jensen and Borum 1991, Valiela et al. 1997, Havens et al. 2001). Enhanced phytoplankton production is also accompanied by higher concentrations of dissolved and particulate organic matter (i.e. bacteria, zooplankton and detritus) that all contribute to greater light attenuation (Kirk 1994, Krause-Jensen and Sand-Jensen 1998). Light attenuation by dissolved organic matter often resembles or even exceeds that of the phytoplankton (Jensen et al. 1987). The classical development of plant communities accompanying anthropogenic eutrophication has therefore been a restriction of the depth distribution of benthic algae and plants, a decline of species diversity particularly among the large, slow-growing macroalgae and an increased proportion of small, nutrient-demanding and fast-growing macroalgae, which can result in mass accumulation of drift algae in shallow, protected lagoons (GeertzHansen et al. 1993, Valiela et al. 1997, Middelboe and Sand-Jensen 1998, 2000). Generally, the increased biomass of phytoplankton together with the greater abundance of drift macroalgae and epiphytic microalgae – all mainly supplied with nutrients from the water – have been responsible for the profound decline or complete extermination of seagrasses and large macroalgae. Various plant communities also display marked differences in the temporal dynamics of environmental parameters and of plant growth and losses. But there are also similarities. Even though pigment concentrations and photosynthesis per unit volume vary many-fold among photosynthetic communities integral photosynthesis summed across the photic zone has the same upper limit in phytoplankton and macrophyte communities probably because it is constrained by the same factor - light availability (Krause-Jensen and Sand-Jensen 1998). 6.1 Phytoplankton communities All phototrophic organisms experience the natural diurnal light cycle. Phytoplankton circulating in a well-mixed water column in addition experience profound (102- to 103-fold) variations in irradiance over short time intervals as they are moved from full irradiance at the surface to shade conditions close to or below the photic zone. Under such conditions phytoplankton have to acclimate to a mean irradiance that makes them susceptible to photoinhibition at the surface and to light limitation close to the lower limit of the photic zone. In stratified water masses photoacclimatization can be more precise because phytoplankton populations experience less variability in irradiance at their respective depths. Deep-water maxima are known to consist of shade-acclimated phytoplankton with minimum light requirements for growth which can approach 1 µmol photon m-2 s-1 (i.e. about 0.1% of maximum sunlight, Geider et al. 1985a,b) by attaining the maximum photon absorptance and efficient conversion of energy into photosynthates for the minimum of costs to investment and maintenance of the cells. The measured light compensation points of photosynthesis in microalgae range from low to intermediate (often 1-10 µmol m-2 s-1; Table 7) and as expected from the smaller size, higher pigment concentrations and lower selfshading of microalgae (Figs. 2-3) compensation points are lower than most values for macroalgae and vascular plants (often 7 to 42 µmol m-2 s-1, Table 7).

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The phytoplankton community has a high growth capacity and high cellular nutrient concentrations. The community is also highly susceptible to losses by grazing, attack by pathogens (bacteria, virus, protozoa), senescence and sedimentation (Table 7). Sedimentation loss is a special and often a very prominent loss process for phytoplankton, which does not exist for benthic microalgae and macrophytes. The numerical response of phytoplankton populations is very fast and extensive and species appear, bloom and vanish again in the community over regular, seasonal cycles or more stochastically in connection with extensive light and nutrient availability (Reynolds 1984, Kiørboe 1993) and periods of minimum or maximum grazing from zooplankton or zoobenthos (Kaas et al. 1996; Cloern 2001). It is noteworthy that shallow coastal waters with extensive mussel beds can experience long-lasting clear-water phases which allow benthic algae and plants to thrive even under nutrient-rich conditions (Kaas et al. 1996). In broad-scale comparisons from open oceanic waters to highly eutrophic coastal lagoons phytoplankton communities vary 105-fold in density from about 0.02 to 2000 mg Chl. m-3 (Fig. 7, Krause-Jensen and Sand-Jensen 1998). The increase in light attenuation accompanying the increase in chlorophyll concentration is relatively small in oceanic waters where the chlorophyll concentration is very low and its contribution to total light attenuation is limited relative to background attenuation. The coupling between light attenuation and chlorophyll concentration is tighter at higher chlorophyll concentrations, where the algae (and their released dissolved organic matter) are responsible for most light attenuation. Thus, across the range from sparse to very dense phytoplankton communities the total light attenuation coefficient (K) increases from about 0.04 m-1 to 20 m-1 and the photic zone (zeu = 4.6/K) declines from about 100 m to 0.2 m. Maximum photosynthesis of phytoplankton per unit volume is closely related to the chlorophyll concentration with a mean slope of 1.06 in Model I and 1.13 in Model II regression, suggesting that photosynthesis normalized to chlorophyll increases slightly from oligotrophic to eutrophic habitats (Krause-Jensen and Sand-Jensen 1998). Volumetric photosynthesis increased from 3 x 10-3 mmol O2 m-3 h-1 in the most sparse phytoplankton communities to 103 mmol m-3 h-1 in the most dense. At the highest rates of photosynthesis the turnover rate of the dissolved oxygen pool is only about 20 minutes so oxygen concentrations are liable to profound diurnal changes and photosynthesis to self-limitation in the afternoon due to oxygen accumulation and carbon dioxide depletion. Integral photosynthesis is the depth-integral of volumetric photosynthesis across the entire photosynthetic layer. Integral photosynthesis increases gradually with increasing density of phytoplankton in different aquatic communities and peaks at about 60 mmol O2 m-2 h-1 above chlorophyll concentrations of 20-30 mg Chl. m-3 (Fig. 8, Krause-Jensen and Sand-Jensen 1998). Integral photosynthesis increases hyperbolically as the proportion of available light absorbed by photosynthetic organisms exceeds 40-50% in dense communities. The low photosynthesis of oceanic communities with few algae and thick photic zones is thus comparable to the inefficiency of light capture and photosynthesis of thick macrophyte tissue low in chlorophyll concentration, where light absorption by non-photosynthetic structures has a major impact (Fig. 2, Markager 1993). The same scale-invariant laws thus

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describe light absorption and photosynthesis over the full scale of photosynthetic systems from the smallest cells to the thickest plant communities. Table 7. Characteristic range (25-75 percentiles) of physiological and ecological key-variables for phytoplankton, macroalgae and seagrasses. Data compiled from many sources by Duarte (1995).

Light compensation point (µmol m-2 s-1) Photosynthetic capacity (mg C g C-1 h-1) Tissue C/N ratio (atomic) Tissue C/P ratio (atomic) Nitrogen uptake capacity (µmol g DW-1 d-1) Phosphorus uptake capacity (µmol g DW-1 d-1) Growth rate (doublings day-1) Grazing rate (% biomass grazed day-1) Decomposition rate (ln units day-1)

Phytoplankton

Macroalgae

Seagrasses

7 – 42

27 – 130

24 – 50

33.8 – 218.8

0.2 – 1.8

0.7 – 1.5

6 – 11

12 – 22

17 - 29

71 – 165

496 – 1220

306 – 669

3376 – 15743

344 – 4370

24 – 206

11208 – 26912

43 – 1061

9 – 161

0.37 – 1.53

0.11 – 0.32

0.02 – 0.03

0.1090 – 0.2600

0.0044 – 0.0071

0.0003 – 0.0030

0.036 – 0.070

0.028 – 0.051

0.009 – 0.056

Applying the light attenuation coefficient per unit of chlorophyll between 0.0081 and 0.0246 m-1 per mg Chl. m-3 (10-90% percentiles) it is possible to estimate the maximum attainable biomass in the photic zone assuming that only the algae contribute to light attenuation (Krause-Jensen and Sand-Jensen 1998). If the photic zone is assumed to extend to depths where only 1% of surface light is available, the total attenuation for the entire photic zone will be 4.6 and estimates of maximum integral chlorophyll concentrations yield 187-568 mg Chl m-2. Phytoplankton biomasses reported for the photic zone typically range from 1 to 580 mg Chl. m -2. The entire phytoplankton biomass in the water column can be higher as the critical depth (i.e. the depth above which net metabolism of the phytoplankton community is zero on a 24-hour basis) surpasses the depth limit of the photic zone (i.e. the depth at which net metabolism of phytoplankton kept at that depth is zero). If losses by respiration, grazing and senescence are small the entire biomass can be substantially higher than the biomass in the photic zone. For phytoplankton the maximum reported total biomass is 1800 mg Chl m-2 (Krause-Jensen and Sand-Jensen 1998). For macrophytes the light attenuation coeficients are lower from 0.0041 to 0.012 m-1 per mg Chl m-3 (10-90 percentiles), which yield higher photic zone biomasses of 3801120 mg Chl m-2. Measured total macrophyte biomass reaches 5000 mg Chl m-2 or three-fold higher than measured values for phytoplankton communities because

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chlorophyll-specific attenuation coefficients and loss processes are smaller for macrophytes than for phytoplankton.

Fig. 7. Maximum volumetric productivity versus chlorophyll concentration within phytoplankton, macrophyte and benthic microalgal communities. The double-log-transformed dataset fits separate model I linear regressions for phytoplankton (line a), y = 1.06x – 0.4, r2 = 0.88, n = 140, P < 0.001; macrophytes (line b), y = 0.35x + 0.81, r2 = 0.22, n = 27, P < 0.05; and benthic macroalgae (line c), y = 0.55x + 0.82, r2 = 0.53, n = 63, P < 0.001. Copyright (1998) by the American Society of Limnology and Oceanography, Inc.

Fig. 8. Maximum depth-integrated (integral) productivity versus chlorophyll concentration within phytoplankton, macrophyte and benthic microalgal communities. The dataset fits an envelope function representing 90% percentiles of 10 consecutive data points. Copyright (1998) by the American Society of Limnology and Oceanography, Inc.

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6.2 Benthic microalgal communities Benthic microalgae resemble planktonic microalgae in many ways. They often come from the same algal groups and cover the same range in cell size and shapes. Therefore, inherent properties regarding photoadaptation, light-use efficiencies, nutrient kinetics, metabolic rates and growth rates should be the same for benthic and planktonic microalgae (e.g. Borchardt 1996, Hill 1996, Stevenson 1996). There are, however, some systematic differences in species composition, life forms and contribution of different algal groups between the two communities. Blue-green algae and diatoms are more prominent in the benthos than in the plankton and though some species are found in both habitats, many species are mainly restricted to one of them. Large, heavy and non-motile forms and firmly attached species that grow flat on the substrata and better resist grazing are likely to be more common in the benthos (Stevenson 1996). Diatoms on mucilage stalks and filamentous algae with holdfasts also confer advantages to benthic algae and permit them to form a three-dimensional structure consisting of an overstory layer of filamentous and stalked forms and an understory layer of firmly attached algae in closer contact with the substratum. This difference of contact with the substratum and the water phase can influence the relative supply of nutrients to overstory and understory algae from the water, from the microbial community through recirculation and from the substratum (Moeller et al. 1988). Species that move by means of flagella are common in both the benthos and the plankton, while species that glide slowly by means of sheaths (blue-green algae) or raphes (diatoms) in close contact with particles are more prominent in the benthos. By movement along the steep gradients in light, nutrients, inorganic carbon, oxygen and sulfide, microalgae have the opportunity to optimize resource acquisition and minimize chemical stress. Being unaffected by sedimentation, benthic microalgae can afford being larger and heavier than planktonic microalgae but whether they do indeed have different size distributions and specific densities remain to be tested. Most of the differences between benthic and planktonic microalgae should therefore stem from differences in the environmental conditions caused by the solid surface and the complex structure and very dynamic processes of microbial communities containing inorganic and organic particles, muco-polysaccharides, mixed flocculates of iron, aluminum, manganese, carbonate and organic matter and a large variety of microorganisms. Compared to the phytoplankton, the light climate is more predictable for the benthic microalgae, because they can take up a fixed position exposed to a regular daily light cycle. It should be easier for benthic algae occupying distinct depth zones and light climates to become fully acclimated by means of regulations of pigment and enzyme concentrations. This hypothesis has been supported by measurements showing: 1) high concentrations of photo-protective pigments (e.g. carotenoids) and lack of photoinhibition of surface algae exposed to permanently high irradiances (Revsbech and Jørgensen 1986, Howard-Williams and Vincent 1989, Hill et al. 1995), 2) prominent photoinhibition of benthic communities only when they derive from highly shaded habitats and experimentally are exposed to full irradiance (Boston and Hill 1991), and 3) pigment and photosynthetic adjustments to alterations of intensity and spectral composition of light with depth through microbial photosynthetic mats showing highly selective absorption and

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scattering (Plough et al. 1993, Kühl and Jørgensen 1994). The motile forms can readily change light exposure by moving in front or behind sediment particles and by moving towards or away from the sediment surface. The surface sediments are usually characterized by high nutrient concentrations due to microbial decomposition or chemical dissolution of fine organic and mixed organic-inorganic particles settling from the water column. Under anoxic conditions a few mm below the sediment surface, iron- and manganese oxy-hydroxides dissolve and release bound phosphate, while in oxic surface sediments they precipitate again providing an efficient internal recirculation and re-binding of limiting phosphate. Photosynthesis in illuminated sediments with high oxygen concentrations in bottom waters will help to trap phosphate within the sediment by oxygenation and thereby restrict release to the overlying water and subsequent phytoplankton growth (Carlton and Wetzel 1988, Hansson 1988, 1990, 1992). In contrast, the short distance between oxic and anoxic zones in the sediment can lead to profound losses of nitrogen by the coupling of nitrification in the oxic surface layers to denitrification in the anoxic, deeper layers. Nonetheless, nutrient availabilities are usually so high that nutrient limitation of benthic microalgae is much less common than of phytoplankton, though the likelihood of benthic nutrient limitation increases in coarse, nutrient-poor sediments, in calcareous sediments where phosphate precipitates as apatite and in communities of high algal biomass and associated high nutrient requirements. Growth restrictions of benthic microalgae are more often due to insufficient light reaching the community, to burial of sediment microalgae in darkness below the thin photic zone and to self-limitation of photosynthesis due to profound build-up of oxygen and pH in the illuminated surface layers of high algal density (Revsbech and Jørgensen 1986). The main loss processes constraining biomass accrual are grazing and sediment pertubation (Admiraal 1984). Phosphorus affinity is usually several-fold higher in phytoplankton than benthic microalgal communities. These differences are influenced both by the differences in cellular concentrations and, thus, in nutrient demands of the two communities and by differences in transport resistance from the surrounding medium to the cells. Algal cells that are in short supply relative to nutrient demand can actively increase their nutrient affinity by adjustment of the type, density and activity of membrane transporters. This adjustment is evident both in planktonic and benthic communities when nutrient-poor and nutrient-rich habitats or seasons are compared. Thus, phosphorus affinity (i.e. Vmax/Km, Healey 1980) can vary about 100-fold in the same habitat between seasons of variable nutrient availability both in the plankton and in the benthos (Hwang et al. 1998). Likewise, phosphorus affinity can vary about 10fold between nutrient-rich and nutrient-poor sites within the same ecosystem and season (Hwang et al. 1998). Increases of phosphorus affinity in response to limitation are due both to increases of Vmax-values and to reduction of Km-values (Gotham and Rhee 1981). Km-values are strongly influenced by transport limitation, which is more likely to take place in benthic than planktonic communities because there are relatively thick diffusive boundary layers in the water overlying sediments and in the surface matrix of the benthic community. Thus, the observed systematic differences in nutrient affinities between planktonic and benthic microalgae are likely to derive from differences in nutrient requirements and transport resistance rather than from differences in inherent kinetic constants.

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Benthic microalgal communities can be exceptionally dense and have the highest chlorophyll concentration (up to 7 x 105 mg Chl m-3) and volumetric photosynthesis (up to 2 x 105 mmol O2 m-3 h-1) of any aquatic community (Fig. 7, Krause-Jensen and Sand-Jensen 1998). Maximum photosynthetic rates will contribute to a turnover of the oxygen pool in the photic layer within a few seconds explaining why oxygen concentrations are highly supersaturated and CO2 concentrations highly depleted in the light. Rates of volumetric photosynthesis increased with chlorophyll concentrations with average slopes of 0.55 and 0.76 (Model I and II regressions) which are significantly lower than those for phytoplankton. In the region where chlorophyll concentrations overlap for benthic and planktonic communities photosynthetic rates are significantly lower for the benthic communities suggesting that their photosynthetic rates are somewhat constrained either by inherent differences from the phytoplankton and/or by self-limitation of benthic photosynthesis because of the extensive build-up of oxygen and pH and the depletion of CO2 in the dense microbial mats. The lower upper limits for integral photosynthesis of communities of benthic microalgal compared with those of phytoplankton and macrophytes (Fig. 8) support the hypothesis of transport limitation of photosynthesis in microbial mats possibly because high ratios of oxygen to inorganic carbon hamper the activity of the primary enzyme (Rubisco) involved in the assimilation of carbon dioxide. This aspect awaits experimental tests. 6.3 Macrophyte communities of macroalgae and rooted plants Macroscopic macroalgae and rooted plants differ from microalgae by living longer and having lower cellular concentrations of nutrients, pigments and enzymes, lower uptake capacity and affinity for nutrients, and lower rates of photosynthesis, respiration and growth normalized to biomass (Table 7). As mentioned above, macrophyte communities can reach three-fold higher maximum areal chlorophyll densities than phytoplankton communities, but the difference in carbon biomass is even higher as the carbon to chlorophyll ratio of most macrophytes markedly exceeds that of phytoplankton (Duarte 1992). The higher carbon biomass of macrophytes can be explained by the much lower resource requirements relative to biomass and by the lower losses by respiration, senescence and grazing compared to the phytoplankton. Although microalgae and macrophytes can be treated as contrasts to the microalgae, the variability among macroalgae is profound both in terms of size, shape, life forms and functional properties, while the few seagrass species show much less variability. Macroalgae vary more than 103-104-fold in linear size and 30-fold in maximum growth rates, whereas seagrasses vary only about 10-fold in size (for rhizome thickness), and less in growth rates (Duarte 1991a, Hemminga and Duarte 2000). Macroalgae grow attached to solid rocks and stones and their community structure and depth distribution are very susceptible to physical perturbation and grazing. Large leathery algae can extend to depths receiving, on average, only about 0.5% of surface irradiance, while thin foliose forms extend to 0.1% and crust-forming species to 0.01% of surface irradiance or less (Markager and Sand-Jensen 1992). These depth limits are shifted upwards in environments with intensive grazing as higher irradiances and faster growth are needed to compensate for the greater losses (Vadas

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and Steneck 1988, Markager and Sand-Jensen 1994), and leathery and foliose macroalgae may disappear entirely under heavy grazing leaving only the grazingresistant crust algae (Foreman 1977). Seagrasses grow in soft-bottom sediments and, in contrast to macroalgae that receive all inorganic nutrients from the water, seagrasses can take up a large proportion of their nutrients from the sediment. The relative role of water versus sediment as nutrient sources to seagrasses depends on the concentrations of available nutrients, the thickness of diffusive boundary layers sourrounding the leaves, the surface area of leaves relative to roots and their ability to extract nutrients from the surrounding medium. Phosphorus can be mobilized from carbonates in tropical sediments (Jensen et al. 1998) perhaps by lowering pH through proton extrusion or release of organic acids. Associations with fungi and bacteria in the rhizosphere can also improve access to firmly-bound sediment nutrients. Even in chronically nutrient-deficient environments (e.g. tropical seas with carbonate sands) seagrasses can develop very extensive and dense meadows due to their access to sediment nutrients and their ability through slow growth, small losses and high tissue longevity to accumulate nutrients in their biomass. Colonization depths of seagrasses show highly significant relationships to water transparency and light attenuation coefficients (Duarte 1991b, Nielsen et al. 2002). On average, seagrasses grow to depths receiving about 11% of surface irradiance, but the variability of minimum light requirements is substantial among species (about 30% for some of them, Kenworthy and Fonseca 1996) and within species from different habitats. For example, eelgrass depth limits can vary from 2 to 8 m among sites having the same mean transparency of 5 m (Nielsen et al. 2002). Analytical errors in the measurements of mean transparency and colonization play a role in this variability. Different light requirements of species, different seasonal light availabilities, and temporal delays in the coupling between light availability and realized depth distribution, and sediment conditions may also play a role (Greve 2004). The systematically higher minimum light requirements of seagrasses (and other aquatic rooted plants) compared to macroalgae deserves an explicit explanation. The most likely candidates are the greater metabolic costs associated with the presence of roots and rhizomes of seagrasses and the suggested need to transport sufficient oxygen, through photosynthesis, to supply the below-ground tissue with oxygen for aerobic respiration and detoxification of plant metabolites and sulfide threatening to invade the root tissue (Borum et al. in press). Chlorophyll concentrations in macrophyte communities typically vary from 102 to 104 mg m-3 and volumetric photosynthesis from 20 to 440 mmol m-3 h-1 (Fig. 8, KrauseJensen and Sand-Jensen 1998). The mean chlorophyll-specific photosynthesis was 0.37 mmol O2 (mg Chl)-1 h-1 for macrophytes, 0.15 for benthic microalgae and 0.52 for phytoplankton. The lower chlorophyll-specific photosynthesis found for macrophytes relative to phytoplankton is a result of a lower chlorophyll-specific attenuation coefficient and thus involvement of a higher total chlorophyll mass in the photosynthetic process as compared to phytoplankton. Thereby, the same integral photosynthesis can be attained in macrophyte communities as in phytoplankton communities.

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7. SUMMARY 1.

2.

3.

4.

5.

6.

7.

Size and shape of photosynthetic organisms display strong statistical relationships to resource utilization, metabolism, growth and loss processes. Much of the functional variability between individual species and communities dominated by different species can, therefore, be related to different sizes and shapes. In addition, there is an evolutionary component involving the adaptation of species to different resource availabilities as reflected by r- and C-selected species in resource-rich environments and S-selected species in chronically resource-limited environments. Suspended or attached life form of photosynthetic organisms influence light availability and transport resistance of dissolved molecules and, thereby, the limitations on photosynthesis and growth. Substrata with variable physical stability and nutrient release rates can support photosynthetic communities of different species composition, biomass and photosynthetic rates. Many of the differences in functional properties between photosynthetic communities can be explained by environmental differences. The importance of different evolutionary adaptations of the species forming the communities is less studied and understood, i.e. the autecological knowledge is sparse and difficult to put into community perspective. Maximum chlorophyll concentrations and volumetric photosynthesis are highest in benthic microalgal communities and lowest in phytoplankton communities. Because of high metabolic rates and high transport resistance, benthic microalgal communities display profound spatial and temporal gradients in oxygen, carbon dioxide, pH, nutrients, sulfide, etc. which require effective physiological adjustments. Integral photosynthesis across the photic zone reaches the same upper limit in phytoplankton and macrophyte communities, while the upper limit is lower in benthic microalgal communities perhaps because of photosynthetic constrains accompanying extensive build up of oxygen and depletion of carbon dioxide in the illuminated dense mats. The same upper limit to integral photosynthesis in dense phytoplankton and macrophyte communities implies that the same carbon-based primary productivity can be attained in shallow coastal waters constrained by light but dominated either by dense macrophyte or phytoplankton communities. In a competitive situation phytoplankton has the advantage by being closer to the light source and benthic organisms by being closer to the nutrient pools in sediments. The common critical factor reducing total primary productivity is background light attenuation by dissolved and particulate organic matter, mineral particles and substrata reducing the amount of light absorbed by photosynthetic pigments of whatever origin. Methodological advances have made it possible to measure the environmental conditions (e.g. light, temperature, water movement, oxygen, pH, nutrient, etc) frequently or continuously together with the photosynthetic properties in microand macrohabitats of aquatic primary producers. The potential of these

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K. Sand-Jensen and S.L. Nielsen advancements awaits to be fully exploited in broad-scale comparative analyses among communities and ecosystems to reach more general understandings and predictions. The greatest challenge is perhaps to integrate food-web interactions in the future models of estuarine behavior.

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Hansson, L.A. (1992). Factors regulating periphytic algal biomass. Limnology and Oceanography 37: 322328. Havens,K.E., Hauxwell, J., Tyler, A.C., Thomas, S., McGlathery, K.J., Cebrian, J., Valiela, I., Steinman, A.D. and Hwang, Soon-Jin. (2001). Complex interactions between autotrophs in shallow marine and freshwater ecosystems: implications for community responses to nutrient stress. Environmental Pollution 113: 95-107. Healey, F.P. (1980). Slope of the Monod equation as an indicator of advantage in nutrient competition. Microbial Ecology 5: 281-286. Hein, M., Pedersen, M.F. and Sand-Jensen, K. (1995). Size-dependent nitrogen uptake in micro-and macroalgae. Marine Ecology Progress Series 118: 247-253. Hemminga; M.A. and Duarte, C.M. (2000). Seagrass ecology. Cambridge University Press, Cambridge.. Hemminga, M.A., Harrison, P.G. and van Lent, F. (1991). The balance of nutrient losses and gains in seagrass meadows. Marine Ecology Progress Series 71: 85-96. Hemminga, M.A., Marba, N. and Stapel, J. (1999). Leaf nutrient resorption, leaf lifespan and the retention of nutrients in seagrass systems. Aquatic Botany 59: 185-194. Hemmingsen, A.M. (1960). Energy metabolism as related to body size and respiratory surfaces. Rep. Steno Mem. Hosp. Nord. Insulin Lab. (Copenhagen) 9: 1-110. Hill, W.R., Ryon, M.G. and Schilling, E.M. (1995). Light limitation in a stream ecosystem:responses to primary producers and consumers. Ecology 76: 1297-1309. Hill, W.R. (1996). Effect of light. In: Stevenson, R.J., Bothwell, M.L. and Lowe, R.L. (eds.), Algal ecology, Freshwater benthic ecosystems, 121-148. Academic Press, San Diego. Holland, A.F., Zingmark, R.G. and Dean, J.M. (1974). Quantitative evidence concerning the stabilization of sediments by marine benthic diatoms. Marine Biology 27: 191-196. Howard-Williams, C. and Vincent, W.F. (1989). Microbial communities in southern Victoria Land streams (Antarctica). i. Photosynthesis. Hydrobiologia 172: 27-38. Hurd, C.L., Harrison, P.J. and Druehl, L.D. (1996). Effect of seawater velocity on inorganic nitrogen uptake by morphologically distinct forms of Macrocystis integrifolia from wave-sheltered and exposed sites. Marine Biology 126: 205-214. Hurd, C.L., Stevens, C.L., Laval, B., Lawrence, G.A. and Harrison, P.J. (1997). Visualization of seawater flow around morphologically distinct forms of the giant kelp Macrocystis integrifolia from wavesheltered and exposed sites. Limnology and Oceanography 42: 156-163. Hwang, Soon-Jin, Havens, K.E. and Steinman, A.D. (1998). Phosphorus kinetics of planktonic and benthic assemblages in a shallow subtropical lake. Freshwater Biology 40: 729-745. Jensen, H., McGlathery, K.J., Marino, R. and Howarth, R.W. (1998). Forms and availability of sediment phosphorus in carbonate sand of Bermuda seagrass beds. Limnology and Oceanography 43: 799-810. Jensen, L.M., Marcher, S. and Hansen, M. (1987). Produktion og omsætning af organisk stof I de frie vandmasser i Roskilde Fjord. MS-Thesis. Freshwater Biological Laboratory, University of Copenhagen. Jørgensen, B.B. (2001). Life in the diffusive boundary layer. In: Boudreau, B.P. and Jørgensen, B.B. (eds.), The benthic boundary layer, 348-373. Oxford University Press, Oxford.

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Jørgensen, B.B. and Des Marais, D.J. (1988). Optical properties of benthic photosynthetic communities. Limnology and Oceanography 33: 99-113. Jørgensen, B.B. and Revsbech, N.P. (1990). The diffusive boundary layer of sediments: Oxygen microgradients over a microbial mat. Limnology and Oceanography 35: 1343-1355. Jumars, P.A., Eckman, J.E. and Koch, E. (2001). Macroscopic animals and plants in bentic flow. In: Boudreau, B.P. and Jørgensen, B.B. (eds.), The benthic boundary layer, 320-347. Oxford University Press, Oxford. Kaas, H., Møhlenberg, F., Josefson, A., Rasmussen, B., Krause-Jensen, D., Jensen, H.S., Svendsen, L.M., Windolf, J., Middelboe, A.L., Sand-Jensen, K. and Pedersen. M.F. (1996). Danske fjorde – status over miljøtilstand, årsagssammenhænge og udvikling. Faglig rapport fra DMU, nr. 179. Miljø- og Energiministeriet, Danmarks Miljøundersøgelser, Roskilde. Karp-Boss, L., Boss, E. and Jumars, P.A. (1996). Nutrient fluxes to planktonic osmotrophs in the presence of fluid motion. Oceanogr. Marine Biology Annual Review 34: 71-107. Kautsky, N, Kautsky, H., Kautsky I. And Waern, M. (1986). Decreased depth penetration of Fucus vesiculosus (L.) since the 1940’ies indicates eutrophication of th Baltic Sea. Marine Ecology Progress Series 28: 1-8. Kenworthy, W.J. and Fonseca, M.S. (1996). Light requirements of seagrasses Halodule wrightii and Syringodium filiforme derived from the relationship between diffuse light attenuation and maximum light penetration. Estuaries 19: 740-750. Kirk, J.T.O. (1994). Light composition and photosynthesis in aquatic communities. 2. Ed. Cambridge University Press, Cambridge. Kiørboe, T. (1993). Turbulence, phytoplankton cell size ad the structure of pelagic food webs. Advances in Marine Biology 29: 1-72. Krause-Jensen, D. and Sand-Jensen, K. (1998). Light attenuation and photosynthesis of aquatic plant communities. Limnology and Oceanography 43: 396-407. Kühl. M. and Jørgensen, B.B. (1994). The light field of microbenthic communities: radiance distribution and microscale optics of sandy coastal sediments. Limnology and Oceanography 39: 1368-1398. Lambers, H. and Poorter, H. (1992). Inherent variation in growth rate between higher plants: a search for physiological causes and ecological consequences. Advances in Ecological Research 23: 187-261. Lazier, J.R.N. and Mann, K.H. (1989). Tubulence and the diffusive layers around small organisms. DeepSea Research 36: 1721-1733. Les, D.H., Cleland, M.A. and Waycott, M.A. (1997). Phylogenetic studies in Alismatidae. II. Evaluation of marine angiosperms (seagrasses) and hydrophily. Systematic Botany 22: 443-465. Littler, M.M (1980). Morphological form and photosynthetic performance of marine macroalgae: A test of the functional/form hypothesis. Botanica Marina 22: 161-165. Littler, M.M. and Littler, S. (1985). Deepest known plant life discovered on an uncharted seamount. Science 227: 57-59. Lucas, L.V., Koseff, J.R., Cloern, J.E., Monismith, S.G. and Thompson, J. K. (1999a). Processes governing phytoplankton blooms in estuaries. I. The local production-loss balance. Marine Ecology Progress Series 187: 1-15. Lucas, L.V., Koseff, J.R., Cloern, J.E., Monismith, S.G. and Thompson, J.K. (1999b). Processes governing phytoplankton blooms in estuaries. II. The local production-loss balance. Marine Ecology Progress Series 187: 17-30.

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Markager, S. (1993). Light absorption and quantum yield for growth in five species of marine macroalgae. Journal of Phycology 29: 54-63. Markager, S. and Sand-Jensen, K. (1992). Light requirements and depth zonation of marine macroalgae. Marine Ecology Progress Series 88: 83-92. Markager, S. and Sand-Jensen, K. (1994). The physiology and ecology of light-growth relationships in macroalgae. In Round, F.E. and Chapman, D.J. (eds.), Progress in phycological research, Vol. 10, 210-298. Biopress, Bristol. Markager, S. and Sand-Jensen, K. (1996). Implications of thallus thickness for growth-irradiance relationships of marine macroalgae. European Journal of Phycology 31: 79-87. Maxwell K. & Johnson G.N. (2000) Chlorophyll fluorescence - a practical guide. Journal of Experimental Botany, 51, 659-668. Meijer M.-L., Jeppesen E., van Donk E., Moss B., Scheffer M., Lammens E., van Nes E., van Berkum J.A., de Long G.J., Faafeng B.A. & Jensen J.P. (1994) Long-term responses to fish-stock reduction in small shallow lakes: interpretation of five-year results of four biomanipulation cases in The Netherlands and Denmark. Hydrobiologia, 275/276, 457-466. Meuwig, J.J., Rasmussen, J. and Peters, R.H. (1998). Turbid waters and clarifying mussels: their moderation of Chl: nutrient relations in estuaries. Marine Ecology Progress Series 171: 139-150. Middelboe, A.L. and Sand-Jensen, K. (1998). Patterns of macroalgal species diversity in Danish estuaries. Journal of Phycology 34: 457-466. Middelboe, A.L. and Sand-Jensen, K. (2000). Long-term changes in macroalgal communities in a Danish estuary. Phycologia 39, 245-257. Moeller, R.E., Burkholder, J.M. and Wetzel, R.G. (1988). Significance of sedimentary phosphorus to a rooted submersed macrophyte (Najas flexilis) and its algal epiphytes. Aquatic Botany 32: 261-281. Monbet, Y. (1992). Control of phytoplankton biomass in estuaries: a comparative analysis of microtidal and macrotidal estuaries. Estuaries 15: 563-571. Nielsen. S.L., Enriquez, S., Duarte, C.M. and Sand-Jensen, K. (1996). Scaling maximum growth rate across photosynthetic organisms. Functional Ecology 10: 167-175. Nielsen, S.L. and Sand-Jensen, K. (1990). Allometric scaling of maximal photosynthetic growth rate to surface/volume ratio. Limnology and Oceanography 35: 177-181. Nielsen, S.L., Sand-Jensen, K., Borum, J., Geertz-Hansen, O. (2002). Depth colonization of eelgrass (Zostera marina) and macroalgae as determined by water transparency in Danish coastal waters. Estuaries 25: 1025-1032. Niklas, K. (1994). Plant allometry. The scaling of form and process. The University of Chicago Press, Chicago. Niklas, K. (1997). The evolutionary biology of plants. The University of Chicago Press, Chicago. Nixon, S., Buckley, B., Granger, S. and Bintz, J. (2001). Responses of very shallow marine ecosystems to nutrient enrichment. Human and Ecological Risk Assessment 7: 1457-1481. Pedersen, M.F. (1993). Growth and nutrient dynamics in marine plants. Ph.D.-Thesis, Freshwater Biological Laboratory, University of Copenhagen. Pedersen, M.F. (1995). Nitrogen limitation of photosynthesis and growth: comparison across aquatic plant communities in a Danish estuary (Roskilde Fjord). Ophelia 41: 261-272.

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Pedersen, M.F. and Borum, J. (1993). An annual budget for a seagrass Zostera marina population. Marine Ecology Progress Series 101: 169-177. Plough, H., Lassen, C. and Jørgensen, B.B. (1993). Action spectra of microalgal photosynthesis and depth distribution of spectral scalar irradiance in a coastal marine sediment in Limfjorden, Denmark. FEMS Microbial Ecology 102: 261-270. Poorter, H. and Remkes, C. (1990). Leaf area ratio and net assimilation rate of 24 wild species differing in relative growth rate. Oecologia 83: 553-559. Rapoport, E.H. (1994). Remarks on marine and continental biogeography: an aerographical viewpoint. Philosophical Transactions of the Royal Society London 343: 71-78. Raven J.A. (1999) Picophytoplankton. In: Progress in phycological research, vol. 13 (eds F.E. Round & D.J. Chapman), pp. 82-106. Biopress. Reuter, J.E., Loeb, S.L. and Goldman, C.R. (1986). Inorganic nitrogen uptake by epilithic periphyton in an N-deficient lake. Limnology and Oceanography 31: 149-160. Reynolds, C.R. (1984). The ecology of freshwater phytoplankton. Cambridge University Press, Cambridge. Reynolds, C.R. (1987). The response of phytoplankton to changing lake lake environments. Schweizerische Zeitschrift von Hydrologie 49: 220-236. Sand-Jensen, K. (1988a). Minimum light requirements for growth in Ulva lactuca. Marine Ecology Progress Series 50: 187-193. Sand-Jensen, K. (1988b). Photosynthetic responses of Ulva lactuca at very low light. Marine Ecology Progress Series 50: 195-201. Sand-Jensen, K. (1989). Environmental variables and their effect on photosynthesis of aquatic plant communities. Aquatic Botany 34: 5-25. Sand-Jensen, K. (2000). Økologi og biodiversitet. Overordnede mønstre for individer, bestande og økosystemer. Gad Publishers, Copenhagen. Sand-Jensen, K. and Borum, J. (1991). Interactions among phytoplankton, periphyton and macrophytes in temperate freshwaters and estuaries. Aquatic Botany 41: 137-176. Sand-Jensen, K. and Mebus, J.R. (1996). Fine-scale patterns of water velocity within macrophyte patches in Danish streams. Oikos 76: 169-180. Sand-Jensen K. and Pedersen O. (1999) Velocity gradients and turbulence around macrophyte stands in streams. Freshwater Biology, 42, 315-328. Sand-Jensen, K., Revsbech, N.P. and Jørgensen, B.B. (1985). Microprofiles of oxygen in epiphyte communities on submerged macrophytes. Marine Biology 89: 55-62. Schindler, D.W. (1987). Detecting ecosystem responses to anthropogenic stress. Canadian Journal of Fisheries and Aquatic Sciences 44 (suppl. 1): 6-25. Schmidt-Nielsen, K. (1984). Scaling. Why is animal size so important? Cambridge University Press, Cambridge. Smith, R.E.H. and Kalff, J. (1982). Size dependent phosphorus uptake kinetics and cell quota in phytoplankton. Journal of Phycology 18: 275-284.

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Stevenson, R.J. (1996). An introduction to algal ecology in freshwater benthic habitats. In: Stevenson, R.J., Bothwell, M.L. and Lowe, R.L. Algal ecology: Freshwater benthic ecosystems, 3-30. Academic Press, San Diego. Vadas, R.L. and Steneck, R.S. (1988). Zonation of deep water benthic algae in the Gulf of Maine. Journal of Phycology 24: 338-346. Valiela, I. McClelland, J, Hauxwell, J, Behr, P.J., Hersh, D., Foreman, K. (1997). Macroalgal blooms in shallow coastal estuaries: controls and ecophysiological and ecosystem consequences. Limnology and Oceanography 42: 1105-1118. Vogel S. (1994) Life in moving fluids. (2. ed.). Princeton University Press, Princeton, N.J. Vollenweider, R.A. (1976). Advances in defining critical loading levels of phosphorus in lake eutrophication. Memorie dell’Istituto Italiano di Idrobiologia. 33: 53-83. Wallentinus, I. (1991). The Baltic Sea gradient. In: Mathiesen, D.C. and Nienhuis, P.H. (eds.), Ecosystems of the world: Intertidal and littoral ecosystems 83-108. Elsevier, Amsterdam. West, G.B., Brown, J.H. and Enquist, B.J. (1997). A general model for the origin of allometric scaling laws in biology. Science 276: 122-126. Zimmerman, R.C. and Kremer, J.N. (1986). In situ growth and chemical composition of of the giant kelp Macrocystic pyrifera: response to temporal changes in ambient nutrient availability. Marine Ecology Progres Series 27: 277-285.

10. AFFILIATIONS K. Sand-Jensen: Freshwater Biological Laboratory, University of Copenhagen, Helsingørsgade 51, DK-3400 Hillerød, Denmark. S.L. Nielsen: Department of Life Sciences and Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark.

J. HAUXWELL AND I. VALIELA

EFFECTS OF NUTRIENT LOADING ON SHALLOW SEAGRASS-DOMINATED COASTAL SYSTEMS: PATTERNS AND PROCESSES 1. INTRODUCTION Coastal waters, including estuaries and nearshore oceanic environments, are among the most highly productive areas in the world. Despite comprising only 1-2% of total ocean area, coastal waters support approximately 20% of total oceanic primary production (Smith 1981, Charpy-Robaud & Sournia 1990) - which in turn fuels approximately 50% of marine fish production (Ryther 1969). Coastal waters are productive because of delivery of land-derived and upwelled nutrients, but in recent decades, terrestrial human sources have prompted cultural eutrophication across the coasts of the world. The modern increases of nutrient inputs, particularly of nitrogen, have changed the composition and abundance of the producers present in coastal systems. Phytoplankton, benthic microalgae, macroalgae, corals, mangroves, salt marshes, and seagrasses are currently being altered by added nutrients. Here we use the case history of the interactions of components in seagrass meadows as an example of the magnitude and complexity of effects of nitrogen additions to coastal ecosystems. Seagrass meadows have historically been a predominant feature of many nearshore coastal environments. However, over the past several decades, shifts in community structure of coastal primary producers, including loss of seagrass habitat, has been a reoccurring phenomenon worldwide (Bayley et al. 1978, Rybicki and Carter 1986, Short and Wyllie-Echeverria 1996, Valiela et al. 1997b, Duffy and Baltz 1998, Hauxwell et al. 2001, Hauxwell et al. 2003). Human-induced disturbances as a result of anthropogenic alterations of landscapes have increasingly degraded water quality of adjacent aquatic systems, and resulted in loss of seagrass habitat (GESAMP 1990, National Research Council 1994, Short and Wyllie-Echeverria 1996, U.S. Geological Survey 1999). In this Chapter, we discuss effects of cultural eutrophication on seagrass meadows of coastal shallow water systems, with an emphasis on describing shifts in the assemblage of primary producers, and processes driving those patterns.

59 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 59-92. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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2. NUTRIENTS: BACKGROUND, NUTRIENT LIMITATION, AND INCREASED LOADING TO COASTAL ZONES 2.1. Background There are several essential nutrients required by plants and algae. Elements essential for survival include: nitrogen, phosphorus, potassium, calcium, magnesium, sulphur, iron, manganese, copper, zinc, molybdenum, sodium, cobalt, chlorine, bromine, silicon, boron, and iodine. The two nutrients that are most generally in short supply and that limit growth of primary producers are nitrogen and phosphorus. Nitrogen, a key component in amino acids, DNA, and RNA, is used in proteins, genes, and chlorophyll. Phosphorus is also a key component in DNA, as well as ATP, and is particularly important in energy transfer and storage in primary producers. While phosphorus is often limiting in freshwater environments, nitrogen is usually limiting in marine environments (Howarth 1988; Caraco et al. 1990). There are exceptions to this pattern, however, usually resulting from unique geological characteristics of certain coastal zones. Phosphate, the form of phosphorus often taken up by primary producers, has a strong binding affinity for sediments rich in calcium carbonate and may be less available to plants or algae when these types of sediments are present, commonly in tropical to semi-tropical waters (Short 1988, Lapointe et al. 1992). Because nitrogen most often limits production in coastal waters, we continue our discussion of nutrient loading with an emphasis on nitrogen. 2.2. Nitrogen limitation of primary production and human alterations to the global nitrogen cycle Many essential nutrients have a mineral source and are made available to primary producers through the weathering of the earth’s crust. Nitrogen, however, is primarily abundant in gaseous form; dinitrogen gas (N2) comprises 78% of the air we breathe. Ironically, such an abundant element may, in fact, limit the growth of primary producers, because nitrogen in the form of N2 cannot be taken up directly by plants or algae, and the atoms of nitrogen in N2 are connected via very strong and stable covalent triple bonds. Lightning can produce the localized energy source necessary to break these bonds, allowing spontaneous formation of NO2 (nitrite), NO3 (nitrate), and NH3 (ammonia). Nitrogen-fixers (in terrestrial environments = symbiotic bacteria in legumes and free-living microbes, in marine environment = cyanobacteria) can also convert N2 to NH4 (ammonium), the form of nitrogen preferentially utilized by many primary producers. Nitrate may be utilized as well, however, its uptake requires energy (ATP), and it must be reduced to ammonium before protein synthesis can occur. The only other natural source of nitrogen is made available through the decomposition of organic material (decay of organic matter, wastewater, burning of fossil fuels). On a global scale, humans have, over the past century, more than doubled the natural rate of transfer of organically-bound or atmospheric nitrogen to biologically available forms of nitrogen (Table 1, Vitousek et al. 1997). Transfer of organically-bound nitrogen has increased due to 1) burning of fossil fuels and 2) burning or clearing

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land. Atmospheric transfer of nitrogen has increased due to 1) increased cultivation of nitrogen-fixing crops (e.g., peas, alfalfa, soybeans) and 2) production of inorganic fertilizer.a Increased production of inorganic fertilizer represents the single largest anthropogenic alteration to the global nitrogen cycle, accounting for over half of the increase in bioavailable N (Vitousek et al. 1997; Howarth et al. 2002). Table 1. Global production of new bioavailable nitrogen in terrestrial environments. Data summarized from Vitousek et al. (1997).

Process Pre-1900s Bacterial fixation Lightning Total Post-1900s Bacterial fixation Lightning Cultivation of N-fixing crops Fertilizer production Burning fossil fuels Burning/clearing existing land Total

New bioavailable nitrogen (million metric tons y-1) 90-140 <10 ∼100-150 90-140 <10 40 80 >20 70 ∼310-360

2.3. Addition of nutrients to coastal zones Increased nutrient loads into estuarine waters have resulted from increases in human populations along coastlines and associated transformation of natural land into urban development, agricultural land, and recreational facilities (i.e. golf courses, Nixon 1995, Cloern 2001). In addition to the sources of new bioavailable nitrogen resulting from human activities (Table 1), it has been estimated that 37% of the world’s population lives within 100 km of the shoreline (Cohen et al. 1997) and that by the year 2010, 75% of the U.S. population will live within 75 km of the coastline (Williams et al. 1991). Not only are concentrated populations of people contributing large quantities of nitrogen to coastal waters via wastewater and fertilizers - but these nutrients are often being imported to coastal watersheds from inland regions. In the United States, for example, a vast majority of crop production occurring inland is imported to support human populations concentrated along coastlines. Therefore, coastal zones receive a disproportionately high load of anthropogenic nutrients. a

In 1913, the process by which artificial fertilizer could be manufactured was discovered in Germany by Fritz Haber, who synthesized ammonia by combining N2 and hydrogen gas at high temperature and pressure. This discovery earned Haber the Nobel prize - ecologists at the time had predicted that the earth had reached its carrying capacity for food production and that mass starvation was imminent unless nitrogen fertilizer could be manufactured.

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Atmospheric deposition, wastewater disposal, and fertilizer use are the primary sources of anthropogenic nitrogen delivered to coastal watersheds (Fig. 1). The nitrogen from these sources is in turn, after substantial biogeochemical transformations—conveyed to receiving estuaries via surface or groundwater flow. Atmospheric deposition of nitrogen and sewage outfalls may also deliver nitrogen directly into estuaries (Fig. 1). Nutrient loading and alterations in nutrient cycling vary across global landscapes as a result of population density (Cole et al. 1993, Seitzinger et al. 2002) and agricultural production (Howarth et al. 2002, Seitzinger et al. 2002). Overall, it has been estimated that the global flux of phosphorus has increased almost 3-fold since the increase in human agricultural and industrial activity over the past century (Howarth et al. 2002). Increase in the pattern of nitrogen use is even greater than that of phosphorus, and, at the high end, nitrogen inputs have been estimated to have increased 6-fold to 8-fold (e.g., the north-eastern U.S. and Chesapeake Bay) (Howarth et al. 2002).

Figure. 1. Sources of anthropogenically-derived nitrogen (atmospheric, fertilizers, and wastewater) and mechanisms by which nitrogen may be delivered to coastal waters (precipitation, run-off, groundwater) (prepared by S. Mazzilli).

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3. COMPONENTS OF SEAGRASS ECOSYSTEMS Seagrasses, algae, and certain bacteria make up the assemblage of primary producers in coastal areas. Under pristine conditions, meadow biomass [but not necessarily primary production (Montcreiff et al. 1992, Kaldy et al. 2002, Hauxwell et al. 2003; Table 2)], is dominated by seagrasses. There are approximately 49 species of seagrasses in the world represented within 12 genera (den Hartog 1970) (with some debate, estimates of number of species range from 47 to 57), and seagrass meadows occupy the coastlines of every continent except Antarctica. Seagrasses exhibit both sexual (monoecious) and vegetative reproductive strategies. In most cases, vegetative reproduction, via elongation of rhizomes and growth of new clonal shoots, contributes the majority to overall recruitment within established meadows. Seagrasses are primarily inter- and subtidal species, with the upper limit of distribution controlled by physical factors, including desiccation, wave action, and ice scour (Robertson and Mann 1984), and the lower limit primarily controlled by light availability (Backman and Barilotti 1976, Dennison 1987, Duarte 1991, Koch and Beer 1996). Nutrient uptake can occur via roots or leaves. Because seagrass roots are typically buried in nutrient-rich sediments, growth is often light-limited and not nutrient-limited. Seagrass meadows are highly productive, fixing, on average across the globe and for all species, 1012 g DW m-2 y-1 (Duarte and Chiscano 1999). Seagrass meadows serve additional ecological roles, including: (1) providing food and shelter for many marine organisms (including commercially important fish and shellfish species), and (2) preventing coastal erosion (binding sediments with root/rhizomes, and by reducing current velocities with aboveground leaves). Seagrass meadows may also be an important carbon sink for the oceans; despite contributing only 1% to total oceanic production, seagrasses may account for 12% of total ocean carbon storage due to their refractory composition (Duarte and Cebrian 1996). Within seagrass meadows, algae may be visually less conspicuous, but ecologically important components. From Table 2, it is apparent that even in relatively healthy seagrass meadows, total primary production is dominated by the algal components. Algae differ from plants in that they lack vascular structures such as roots and stems. Algae may be microscopic, such as single-celled phytoplankton, or macroscopic, and may be free-living (like many fast-growing estuarine species of filamentous macroalgae) or attached to a substrate (epiphytes on seagrass leaves, benthic microalgae, and certain species of macroalgae). Because physiological characteristics of algae result in lower light requirements than seagrasses (Valiela et al. 1997b), algal growth is typically nutrient-limited.

30 (119)

31 (122)

ND

84

13 (256)

Montcreiff et al. 1992 (within-meadow values, not extrapolated to areal coverage) Kinney and Roman 1998 Kaldy et al. 2002 (Budget 3)

69 (518)

53 (180)

Macroalgae

13

0 (0)

Hauxwell et al. 2003

Bass Harbor Marsh, ME, USA Lower Laguna Madre, TX, USA

20 (67)

Hauxwell et al. 2003

Low N estuary, Waquoit Bay, MA, USA High N estuary, Waquoit Bay, MA, USA Mississippi Sound, USA

Seagrasses

Reference

Location

ND

46 (905)

ND

ND

Epiphytes

14 (55)

3

24 (468)

31 (233)

27 (94)

Phytoplankton

25 (96)

17 (339)

ND

Benthic microalgae ND

Table 2. Relative contribution by different producers in shallow water and seagrass meadows, from published reports for various locations, expressed as percentage of production by the different producer types calculated after summing available data. In parentheses, magnitude of production is provided in g C m-2 y-1. ND = not determined.

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4. ESTABLISHING PATTERNS: EFFECT OF NUTRIENT LOADING ON ASSEMBLAGES OF COASTAL PRIMARY PRODUCERS As nutrients are increasingly delivered to coastal waters, response by assemblages of primary producers in shallow waters follows a general pattern: nutrient-limited algal producers respond by growing at faster rates, algal standing stocks increase, and seagrasses decline (Sand-Jensen 1977, Borum and Wium-Andersen 1980, Borum 1985, Silberstein et al. 1986, Valiela et al. 1992, Neckles et al. 1994, Hauxwell et al. 2001). The associations between increased nitrogen loading, algal production, and seagrass decline have been established through field observations (Cowper 1978, Harlin and Thorne-Miller 1981, Neckles et al. 1994, Duarte 1995, Kinney and Roman 1998) and within experimental mesocosms (Twilley et al. 1985, Burkholder et al. 1994; Neckles et al. 1994, Short et al. 1995, Taylor et al. 1995a and 1995b, Moore and Wetzel 2000). Recent and future research aimed at understanding the numerical relationships between nitrogen loading and shifts in community structure of estuarine producers at the watershed-estuary scale is invaluable in answering several questions. For example, do the patterns observed in small-scale experimental settings hold for natural systems? At what level of urban development do ecologically significant changes occur? More specifically, at what loading rate might we expect seagrasses to disappear? How do patterns vary among estuaries that differ in key physical features (e.g., water residence time, depth, latitude, etc.)? In the following sections we will describe the observed changes that occur among different taxa of primary producers as nutrient loads increase and, where data are available, quantify the patterns at the watershed-estuary scale. 4.1. Loss of seagrasses and blooms of algae - observations from around the world and experimental links Over the past 3 decades, seagrass habitat has disappeared from coastal ecosystems worldwide. Habitat can be lost due to natural disturbances, including biological (herbivory, disease, bioturbation), geological (earthquakes), or meteorological (hurricanes, ice scour) events (Short and Wyllie-Echeverria 1996). Human activities may also cause seagrass decline, as a result of practices occurring either (1) directly within the estuary (e.g., dredging, addition of docks, mooring of boats, harvesting of shellfish using rakes or trawls, use of motorboats) (Orth et al. 2002 and references therein) or, (2) as a result of urban development and agricultural activities within adjoining land parcels (e.g., sediment transport, herbicide runoff, increased nutrient loading (Kemp et al. 1983)). Short and Wyllie-Escheverria (1996) published a comprehensive list of reports describing large-scale seagrass loss. In Table 3, we summarize even more recent reports of seagrass habitat loss and the various causes. We present the combined dataset in aggregate (by decade) in Fig. 2. The majority of loss occurring at relatively large geographic scales, for various seagrass species and from sites around the world has been due to anthropogenic influences within adjoining land parcels (Fig. 2, top). In recent decades, the number of documented cases of seagrass decline has increased and ∼90% of loss was

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specifically attributed to human-induced disturbances. In particular, loss resulting from increased nutrient loads or loss that was likely mediated by nutrient loading (e.g., increased turbidity, degraded water quality, eutrophication, pollution) accounted for over 60% of the decline attributed specifically to human activities. Documented cases of seagrass expansion are few (Fig. 2, bottom; Table 4). These few cases, however, have occurred relatively recently, with the majority resulting from successful restoration projects. In a handful of situations (e.g., Tampa Bay, FL, USA; Johansson and Greening 2000), reductions in nutrient loads have been achieved so that seagrasses have returned to systems in which they had been lost. Along with loss of seagrasses, there have often been concomitant reports of increased standing stocks of algae worldwide. Phytoplankton (Valiela et al. 1992, Duarte 1995) and periphyton (Sand-Jensen 1977, Borum and Wium-Andersen 1980, Borum 1985, Silberstein et al. 1986, Neckles et al. 1994, Frankovich and Fourqurean 1997; but see also Lin et al. 1996) often increase with nutrient supply, incurring detrimental effects on seagrasses. Fletcher (1996) listed 50 sites around the world where macroalgal accumulations (typically comprised of Chlorophytes such as Chaetomorpha, Cladophora, Enteromorpha, and Ulva spp.) washed up on beaches were impressive enough to be described as "green tides." In Table 5 we list several sites around the world in which accumulations of drift algae have been reported to actually displace seagrasses. Experimental work has repeatedly demonstrated that macroalgal canopies competitively exclude seagrasses (Holmquist 1997, Hauxwell et al. 2001, Nelson and Lee 2001; but see also Bell and Hall 1997 and Maciá 2000). Though our focus has been documenting the loss of seagrass habitat, increased standing stocks of fast-growing nutrient limited algae may also displace slower growing algal taxa (e.g., fucoids in the North Atlantic, Baltic, Norway, Mediterranean) (Fletcher 1996 and references therein) and corals (LaPointe and Clark 1992, McCook et al. 2001; but see also Miller et al. 1999 and references therein). 4.2. The numerical relationship between nitrogen loading rates and shifts in assemblages of primary producers - ecosystems Given the observational and experimental evidence linking seagrass decline with algal blooms and increased nutrient loads, there is an obvious need for ecosystemlevel studies that can be applied to guide watershed management. There is some difficulty in quantifying the response of estuarine producers to nitrogen loading at relatively large spatial scales (e.g., global or watershed-estuary scales) (Nixon et al. 2001), due largely to 1) acquiring long-term data (significant drain on time and financial resources), 2) the effort and difficulty involved with measuring loads of nitrogen delivered from land to sea (again, significant drain on time and financial resources; Valiela et al. 1992, Tomasko et al. 2001, Howarth et al. 2002, Sigua and Tweedale 2003), and 3) the ability to observe biological patterns, given potential wide variation in background physical and other attributes of different coastal areas (e.g., water residence time, temperature, depth, geology, etc.).

Zostera marina

Syringodium filiforme Syringodium filiforme Thalassia testudinum Thalassia testudinum Zostera capricorni Zostera marina Zostera marina

Posidonia oceanica Posidonia oceanica

Halodule wrightii Posidonia oceanica

Species Amphibolis antarctica Zostera spp. Halodule wrightii

Airport construction Increased turbidity Nitrogen-loading pollution

90% in past decade 77% lost over last

Botany Bay, NSW Australia Barnegat Bay, NJ, USA Waquoit Bay, MA, USA (Hamblin and Jehu Ponds) Niedersachsen, Wadden

18.75 ha

Chronic light reductions

Florida Bay, FL, USA

Florida Bay, FL, USA Outer FL Bay, FL, USA Florida Bay, FL, USA

Chronic light reductions Increased turbidity from fish farming Pollution (sewage) Construction of marinas, dredging, hydrological alterations Chronic light reductions Sea urchins Salinity and other reasons

Persistent brown tide

2.6 km2 (4%) in 3.5 years

Laguna Madre, TX, USA

Florida Bay, FL, USA Harnillo Bay, Murcia Spain (SE coast) Gulf of Marseille, France Corsican coast, Mediterranean

Disturbance Natural, notably warm El Niño summer

Area lost

Location S. Australia

Kastler and Michaelis

Gibbs 1997 Lathrop et al. 2001 Hauxwell et al. 2003

Hall et al. 1999

Hall et al. 1999 Rose et al. 1999 Zieman et al. 1999

Jupp 1977 Pasqualini et al. 1999

Onuf 1996, Street et al. 1997 Hall et al. 1999 Ruiz et al. 2001

Reference Seddon et al. 2000

Table 3. Published reports in which loss of seagrass (on a relatively large (km2) geographic scale) was described including a description of the disturbance attributed to the decline. These reports were used in addition to those reviewed by Short and Wyllie-Echeverria (1996) to generate Figure 2.

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Zostera noltii Zostera spp.

Zostera noltii

Zostera marina 77% lost over last century

Niedersachsen, Wadden Sea German Wadden Sea S. Australia New South Wales, Australia Torres Strait, b/t Papua New Guinea and Australia 85% 60% loss between 1986 and 1993

century 70% since 1977

Sea Seto Inland Sea, Japan

Pesticides Natural, hot El Nino summer

Nutrient loading, drag nets, reclamation and port construction pollution

Kastler and Michaelis 1999 Bester 2000 Seddon et al. 2000 Smith 1997 Long et al. 1997

1999 Komatsu 1997

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Figure. 2. The number of published reports in which significant loss (top panel) or expansion (bottom panel) of seagrass areal coverage was documented, shown in aggregate by recent decades.

Zostera capricorni

Zostera capricorni

Zostera marina Ruppia maritima

Zostera marina

Botany Bay, New South Wales, Australia

Prince William Sound, AK, USA coastal bays of the Delmarva Peninsula, USA Moreton Bay, Australia Recovery less than 1 year after extensive dugong grazing Successful transplantation following airport construction

32 km increase (~145%) in areal coverage since opening of the Gulf Intracoastal Waterway which resulted in reductions in hypersalinity Recovery two years after oil spill

2

56% increase between 1972 and 1993 in holes created by seismic survey

Jervis Bay, New South Wales, Australia Success Bank, W. Australia Laguna Madre, TX, USA

Posidonia australis

Posidonia coriacea Amphibolis griffithii Thalassia testudinum

Recovery after specific disturbance? 87% increase in areal coverage between 1982 and 1992, following reductions in nutrient loading

Location Tampa Bay, FL, USA

Species Thalassia testudinum Halodule wrightii

120% increase in areal coverage between ~1986 and 2000

50% between 1972 and 1993

General expansion

Peterken and Conacher 1997 Gibbs 1997

Orth et al. 2002

Dean et al. 1998

Kaldy and Dunton 1999

Kendrick et al. 1999

Meehan and West 2000

Reference Robbins 1997, Johansson and Greening 2000

Table 4. Documented cases of seagrass expansion including cases in which expansion was attributed to recovery after an initial disturbance and cases in which general expansion was noted.

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Phytoplankton, epiphytes and macroalgae

Enteromorpha radiata

Enteromorpha linza Cladophora laetevirens Ulva rigida Cladophora vagabunda Gracilaria tikvahiae Ulvaria obscura Ulva lactuca

Taxa Caulerpa taxifolia

Valiela et al. 1992, Peckol et al. 1994, Hauxwell et al. 2001 Nelson and Lee 2001 Sfriso 1987, Sfriso et al. 1992

Waquoit Bay, MA, USA NW coast, USA Venice Lagoon, Italy

Zostera marina Zostera marina Cymodocea nodosa Zostera noltii Zostera marina Zostera noltii Thalassia testudinum Halodule wrightii Tampa Bay, FL

Johansson and Greening 2000

den Hartog 1994

Zaitsev 1992, Vasiliu 1996

NW Black Sea

Hampshire, UK

Reference Meinesz et al. 1993

Location N. Mediterranean coasts

Seagrass Cymodocea nodosa Posidonia oceanica Zostera marina Zostera noltii

Table 5. Reports in which drift algae replaced seagrasses.

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Global-scale regional assessments of loadings have been accomplished (Howarth et al. 2002, Seitzinger et al. 2002) that reflect broad-scale patterns in population density, agricultural practices, and industrial and meteorological factors affecting atmospheric deposition, and may be particularly useful from a management perspective in terms of realistic goal-setting within a broad geographic region particularly if reductions in nutrient loads cannot be achieved due to regional patterns in atmospheric deposition (difficult to manage at the smaller watershed-estuary geographic scale). Assessment of nutrient loading and its ecological effects at the watershed-estuary scale seems most appropriate for advancing scientific knowledge related to patterns and processes associated with seagrass loss, as well as serving as the scale most appropriate for management and restoration. Development of numerical relationships among nutrient loading (or concentration), seagrass decline, and algal blooms at the watershed-estuary scale has recently been accomplished in a few systems, including Tampa Bay, FL, USA (Johansson and Greening 2000), Danish estuaries (Nielsen et al. 1989, Sand-Jensen and Borum 1991), and Waquoit Bay, MA, USA (Valiela et al. 1992, Hauxwell et al. 2003). Based on relationships developed in nearby Tampa Bay, projections of future nitrogen loads and accompanying changes in chlorophyll a concentrations and seagrass depth limits have been made for Lemon Bay, FL, USA (Tomasko et al. 2001). Throughout the remaining sections, we highlight our efforts in Waquoit Bay as a case study to illustrate both patterns in changes to community structure of primary producers and mechanisms that drive those patterns. 4.3. Case study - Waquoit Bay Over the past decade, we have conducted research in the Waquoit Bay estuarine complex located on Cape Cod, MA, USA, where different land use patterns within the watersheds of seven estuaries, similar in physical characteristics (e.g., depth, water residence times, temperature, latitude, etc.); have resulted in different annual loads of nitrogen delivered to those estuaries. Because increased urbanization within certain watersheds is accompanied by increases in delivery of land-derived nitrogen, we were able to use the estuaries of Waquoit Bay in a space-for-time substitution (Pickett 1989) to infer the time course of increased eutrophication created by increasing urbanization of watersheds. Because the range of nitrogen loads delivered to Waquoit Bay estuaries encompasses ~75% of the range of reported loads to different estuaries around the world, the response by estuarine producers to increased nitrogen loading in Waquoit Bay may be representative of all but the most eutrophic conditions encountered worldwide. In Figure 3, we plotted seagrass loss, macroalgae, and phytoplankton versus nitrogen loading. Seagrass loss occurs rapidly as nitrogen loads increase (Fig. 3, top), and the pattern observed in the Waquoit Bay estuaries holds at the larger scale; the same logarithmic relationship was observed using a compilation of published reports around the world (Valiela and Cole 2002). Seagrasses remain only in estuaries of the lowest nitrogen loading rates (lowest 12% of the range of loads delivered to estuaries worldwide). In the Waquoit Bay estuaries, macroalgal canopy heights (Fig. 3, middle) and phytoplankton (Fig. 3, bottom) increased linearly with nitrogen loading rate. These sorts of patterns have been observed in other systems; Nielsen et al. (1989) compared 20 Danish estuaries that spanned a nitrogen loading gradient and

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demonstrated a positive linear relationship between phytoplankton and total nitrogen concentration, as well as depth penetration of seagrasses and total nitrogen.

Figure. 3. Relationship between seagrass loss (Zostera marina for the Waquoit estuaries between 1987 and 1997 adapted from Hauxwell et al. (2003); other reports from around the world adapted from Valiela and Cole (2002)), maximum macroalgal canopy height (annual maximum from monthly sampling, adapted from Hauxwell et al. (2003)), and phytoplankton standing stocks (4-year annual average between 1990 and 1994 monthly sampling, Foreman et al. (in preparation); lower panel figure prepared by G. Tomasky) and nitrogen loading rates (modelled, Valiela et al. 1997a) to the Waquoit Bay estuaries.

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4.4. The effect of physical variation among systems on ecosystem-scale patterns It is important to note that variations in the physical backdrop will effect (and potentially mask) the patterns we described above. In essence, the relationships described above (specific to small, shallow, temperate estuaries with a Cape Cod geology, with short water residence times….) can be viewed as one subset within a suite of patterns, with each subset representing systems that differ in key physical features important enough to disrupt ecosystem-scale patterns. In Figure 4, we summarized the effect of land-derived nitrogen loads on the relative contribution to estuarine primary production by 3 major components (seagrasses, macroalgae, and phytoplankton) in estuaries with short or longer water residence times. In the upper panel, we present data from systems with water residence times of duration similar to that of the Waquoit estuaries (≤3 d). Seagrass production decreases sharply with nitrogen load and is replaced by phytoplankton and macroalgal production. In the lower panel, we present data from systems with significantly longer water residence times (≥45 d). In these settings, seagrasses are also rapidly lost, but the percentage of production carried out by phytoplankton and macroalgae follow a different pattern from that described for systems of shorter water residence times. Presumably, because phytoplankton has more time to grow, it is able to outcompete benthic producers (via shading or nutrient competition) and dominate production across much of the nitrogen loading range. In summary, we emphasize the importance of physical factors in mediating the patterns observed with nitrogen loading, and suggest that we may be most successful in observing patterns when we aggregate data within physically similar systems.

5. UNDERSTANDING PROCESSES: HOW ARE SEAGRASSES LOST? Increasing evidence demonstrates at least 3 mechanisms (both direct and indirect) by which increased nitrogen supply leads to seagrass decline: (1) direct nitrate toxicity (Burkholder et al. 1992, 1994; Touchette et al. 2003), (2) light limitation (Short et al. 1993, Valiela et al. 1997b) via phytoplankton, epiphytes, and/or macroalgae (Backman and Barilotti 1976, Harlin and Thorne-Miller 1981, Twilley et al. 1985, Howard and Short 1986, Short et al. 1993, Duarte 1995, Short et al. 1995, Taylor et al. 1995a and 1995b, Jernakoff et al. 1996, Moore et al. 1996, Kinney and Roman 1998), and (3) unfavourable alterations in the biogeochemical environment surrounding seagrasses (van Katwijk et al. 1997, Greve et al. 2003) that may be imposed by macroalgal canopies (Valiela et al. 1992, Bierzychudek et al. 1993, D’Avanzo and Kremer 1994, Krause-Jensen et al. 1996, McGlathery et al. 1997). 5.1. Nitrate toxicity Burkholder et al. (1992, 1994) measured decreased production by Zostera marina (but increased production by Halodule wrightii and Ruppia maritima) in mesocosm plots receiving 5-10 µM daily pulsed additions of nitrate. Touchette et al. (2003) measured a 40% decrease in shoot density, and decreased leaf and root production by Zostera marina in mesocosms receiving daily pulsed additions of nitrate (8 µM) relative to controls.

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Figure. 4. Partition of total primary production in shallow estuaries into contributions by phytoplankton, macroalgae, and seagrasses, in relation to measured annual land-derived nitrogen loading rates (Valiela et al. 2000a) in estuaries of short (top panel) or longer (bottom panel) water residence times. The dark shaded, light shaded and white areas represent the percentage of total primary production contributed by seagrasses, macroalgae, and phytoplankton, respectively. Adapted from Valiela et al. 2000b.

Because photosynthetically active radiation reaching eelgrass was not reduced in nitrate-enriched mesocosms and epiphytes were routinely cleaned from leaves, the discrepancy was attributed directly to nitrate variation. We are not aware of published studies in which a link between nitrate toxicity and seagrass loss at the estuary scale has been established, and further research in this area is warranted.

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Nitrate concentrations in highly loaded systems can exceed critical levels demonstrated in these experiments. 5.2. Light limitation Seagrasses are generally light-limited, with intensity and photoperiod as important factors. In numerous studies, when light has been experimentally reduced, seagrass production decreased (Backman and Barilotti 1976, Dennison and Alberte 1985, Longstaff et al. 1999, and many others). When light has been experimentally augmented, seagrass production increased (Dennison and Alberte 1985). In numerous cases, declines in seagrass coverage have been attributed to concomitant increases in standing stocks of phytoplankton, epiphytes, and/or macroalgae. Because these primary producers are also dynamic biological and chemical components of systems, other processes must be considered as well, including their ability to compete with seagrasses for nutrients. Below, we outline the evidence for light limitation of seagrasses by the various types of algal producers, and attempt to distinguish cases in which light limitation can be ruled as the probable mechanism. 5.2.1. Effect of phytoplankton Calculating the contribution of phytoplankton to shading of seagrasses is relatively straightforward, but it is important to recognize that phytoplankton is one component of the entire water column affecting light penetration1. In a comparative study, Krause-Jensen and Sand-Jensen (1998) found phytoplankton alone (minus associated water column components) to be much less effective at reducing light than other taxa of producers such as macrophytes and benthic microalgae. Since the relationship between light intensity and growth has been experimentally established for several species of seagrasses, it is possible to estimate reductions in growth rates, given depth and background attenuation due to nonphotosynthetic particles, and water column chlorophyll concentrations. For example, at a high extreme, standing stocks of phytoplankton in a eutrophic estuary such as the Childs River, Waquoit Bay, MA, may reduce up to 89% of light reaching the water surface during summer peaks (120 mg chlorophyll a m-3; Tomasky et al. 1999), which would decrease eelgrass (Zostera marina L.) production by 80% (Short et al. 1995). Proliferations of phytoplankton have been directly linked to seagrass loss. In the Laguna Madre, Texas, a persistent brown tide resulted in a rapid loss of 2.6 km2 of Halodule wrightii after 3.5 years (Onuf, 1996). In this case, light limitation as a process by which phytoplankton affect seagrasses is corroborated by a similar example in which sediment loading resulted in similar losses of seagrass habitat. 1

Beer’s Law can be used to determine the fraction of surface irradiance (Iz/Io) reaching depth (z): Iz/Io = e-(KB+KP)(z), and the contribution to attenuation made by the various components of the water column. The total water column light attenuation coefficient (KT , in units /m), is equal to the sum of 1) the light extinction coefficient due to absorption by water and its nonphytoplankton components (KB) and 2) the light extinction coefficient due to phytoplankton (KP), which equals chlorophyll concentration (C, in units mg m-3) multiplied by the chlorophyll-specific light attenuation coefficient (kc = 0.016 m2/mg chlorophyll; Bannister 1974, Kirk 1994).

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Light attenuation due to sediment resuspension (resulting from maintenance dredging on the Gulf Intracostal Waterway) had same effect on Halodule in the lower Laguna Madre, resulting in a 60% decrease between 1965 and 1988 (Onuf 1994, Pulich et al. 1997). The relationship between depth limits for seagrasses and light availability, as modified by phytoplankton or sediment, also supports light availability as a primary factor by which phytoplankton may influence distributions (Duarte 1991). In a study that included 27 Danish fjords, Nielsen et al. (2002a) demonstrated that (1) there was a positive linear relationship between eelgrass depth limit and Secchi depth (which is strongly dependent on phytoplankton biomass and decreased with nitrogen concentration; Nielsen et al. 2002b), and (2) there was a negative relationship between total water column nitrogen concentration and eelgrass depth limit (power function after log-transformation). Similar relationships were observed between seagrass depth limits in Moreton Bay, Australia, and water column light attenuation coefficients (also related to chlorophyll a concentration) (Abal and Dennison 1996). 5.2.2. Effect of epiphytes There are a number of ways increased epiphytes may negatively affect seagrasses (Jernakoff et al. 1996). In high energy systems, increased epiphyte loads may increase the drag on seagrasses, resulting in increased loss of leaves or whole shoots. Epiphytes also create a boundary layer around the leaf, which may affect carbon and nutrient uptake and oxygen diffusion. Architecturally, epiphytes may reduce the amount of light reaching seagrass leaves. The most convincing evidence linking epiphyte loads to decreased growth by seagrass plants comes from a handful of experiments. Sand-Jensen (1977) demonstrated that epiphytes (at ambient loads in Vellerup Vig, Denmark), reduced eelgrass photosynthesis by 31% at optimum light conditions relative to leaves that had been cleaned. Howard and Short (1986) documented a 22% decrease in Halodule wrightii leaf elongation in plots that excluded epiphyte-consuming grazers. The numerical relationship between epiphyte biomass and light attenuation is best represented in Brush and Nixon (2002), where they directly measured light transmission through unscraped and scraped seagrass blades across a wide range of epiphyte densities, using natural plants and undisturbed epiphyte communities. They found that attenuation of photosynthetically active radiation through epiphytes was well-described by a negative hyperbolic equation, with variation in terms depending on the type of epiphytic communities.2 Maximum epiphyte densities for marine systems (as reviewed in Brush and Nixon (2002)), ranged between 2.1 and 99 mg cm2

Equations from Brush and Nixon (2002) are listed below for various types of epiphytic communities (epiphyte biomass units in mg DW cm-2) (see Brush and Nixon 2002 for review of all other published equations): (1) for an epiphytic community dominated by Cladophora sp.: % transmission = 100 (89.1(biomass/(biomass + 17.3)), (2) for an epiphytic community dominated by Polysiphonia sp.: % transmission = 100 (92.4(biomass/(biomass + 2.2)), (3) for an epiphytic community dominated by an unidentified green filamentous (probably Ulothrix sp.): % transmission = 100 - (95.7(biomass/(biomass + 3.3)).

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2

of leaf, resulting in % attenuations approximately 10-45% and 76-90%, respectively (depending on epiphyte architecture). These calculations demonstrate relatively substantial losses of light at even relatively low epiphyte loads. Though robust measurements of light attenuation are possible, the exercise of applying them and understanding the effect of light limitation imposed by epiphytes is more complex than that of phytoplankton or macroalgae, and inconsistent effects of epiphytes on seagrass photosynthesis are represented in the literature (see review by Jernakoff et al. 1996). The growth pattern of most seagrasses, in which old leaves are sloughed as new leaves appear, is undoubtedly advantageous, in terms of minimizing build-up of epiphytic material on photosynthetically-active surfaces (Borum 1987). 5.2.3. Effect of macroalgae Accumulations of macroalgae may shade seagrasses. Peckol and Rivers (1996) measured rapid attenuation of light through macroalgal canopies comprised of either a fine green filamentous alga (Cladophora vagabunda) or a thicker red branched alga (Gracilaria tikvahiae) (Figure 5, top). Regardless of algal taxon, the percentage of light reaching the surface of the macroalgal canopy was reduced by > 95% after penetrating only 6-8 cm into the canopy, and for both species, light attenuation was best described by a negative logarithmic equation. A rapid decay of light was also measured by McComb et al. (1981), in which 99% of light was attenuated within a 1cm thick algal mat. Algal standing stocks are usually reported as biomass, so in Figure 5 (bottom), we illustrate the relationship between biomass and canopy height for green filamentous (top, Peckol and Rivers (1996)), red branched (middle, Peckol and Rivers (1996)), and a green foliose (bottom, Coffaro and Bocci (1997)) algae. In all cases, a linear fit best described the data, with variation in slopes resulting from different morphologies. In Waquoit Bay, macroalgal canopy heights of 6-8 cm may be sparsely distributed in estuaries with low nitrogen loads, and, under high nitrogen loading conditions, evenly distributed macroalgal canopies of 75 cm may be attained (Hauxwell et al. 2001). Examples of standing stocks of this order and greater can be found around the world. In Hog Island Bay, VA, USA, canopies comprised of Ulva sp., Gracilaria tikvahiae, and Cladophora sp., have reached 650 g DW m-2 (Havens et al. 2001). In Scotland, Coleman and Stewart (1979) reported standing stocks of Enteromorpha prolifera that reached 1000 g DW m-2. In the Upper Newport Bay estuary, CA, USA, Kamer et al. (2001) reported algal peaks >150 g DW m-2 (comprised of Enteromorpha intestinalis, Ulva expansa, and Ceramium spp.). In Coos Bay, OR, USA, biomass of Enteromorpha spp. and Ulva spp. have reached 750 g DW m-2 (Pregnall and Rudy (1985)), and in the Palmones River estuary, southern Spain, standing stocks of Ulva spp. have reached 375 g DW m-2 (Hernández et al. 1997).

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Figure. 5. Top: Light transmission through macroalgal canopies comprised mainly of a filamentous green alga (Cladophora vagabunda) and a branched red alga (Gracilaria tikvahiae) as a function of canopy height, redrawn from Peckol and Rivers (1996), with permission from Elsevier. Best-fit equations were: y = -24.8 log (x) + 21.1, r2 = 0.92 and y = -30.7 log (x) + 32.6, r2 = 0.89, respectively. Bottom: Relationship between macroalgal biomass and canopy height for canopies comprised mainly of a filamentous green alga (Cladophora vagabunda), a branched red alga (Gracilaria tikvahiae) (redrawn from Peckol and Rivers (1996), or a foliose green alga (Ulva rigida) (redrawn from Coffaro and Bocci (1997)). Best fit equations were: y = 39.7x + 19.9, r2 = 0.68 (Peckol and Rivers (1996)); y = 36.2x 12.7, r2 = 0.55 (Peckol and Rivers (1996)); and y = 0.17x + 15.5, r2 = 0.91 (estimated from Coffaro and Bocci (1997)), respectively.

Incredibly, at a peak in 1987, biomass of macroalgae in Venice Lagoon (primarily comprised of Ulva rigida) reached over 1800 g DW m-2 (Sfriso et al. 1992), corresponding to canopy heights estimated to have exceeded 2 m in height. This

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collection of observations is by no means an exhaustive list, and these few are mentioned to (1) reiterate the cosmopolitan nature of macroalgal accumulations, and (2) illustrate that standing stocks easily attain heights that may attenuate >99% of light available for newly recruiting seagrasses. 5.2.4. Case study - Waquoit Bay For 2 Waquoit Bay estuaries, we conducted a modelling exercise to partition the relative importance of summer standing stocks of phytoplankton, epiphytes, and macroalgae to potential shading of eelgrass (Fig. 6, Hauxwell et al. (2001)). One estuary featured a low nitrogen loading rate (5 kg N ha-1 y-1) and sustained a relatively pristine eelgrass population with patchy distributions of macroalgae, on average 2-cm high. The other estuary featured a six-fold higher nitrogen loading rate (30 kg N ha-1 y-1) and a declining eelgrass population, with a relatively uniform 9-cm high macroalgal canopy. Because irradiance reaching photosynthetic surfaces is dependent on plant height (for water column attenuation and macroalgal interaction) and age (for epiphytes), we considered scenarios for tall established shoots and also for small newly recruiting shoots (details in Hauxwell et al. (2001, 2003)). Though background attenuation through the water column was numerically important (22-60%), we estimated that light attenuation due to phytoplankton , at most, to be only 6% of incoming light (Fig. 6). This is quite low compared to our maximum calculations for epiphytes and macroalgae, which may have attenuated 63% and 99% of incoming light, respectively. While water column and epiphyte shading were estimated to be quite important for older shoots, light reduction values resulting from shading via macroalgae were numerically more important than the other categories of producers for newly recruiting shoots. Based on eelgrass light requirements (Dennison and Alberte 1982; light saturation at 100 µmol photons m-2 s1 and compensation at 10 µmol photons m-2 s-1), we estimated newly recruiting shoots in the higher N estuary to be light limited (4 µmol photons m-2 s-1). To conclude, light limitation was determined to be a possible mechanism by which macroalgal canopies exclude eelgrass, both experimentally and within certain estuaries of Waquoit Bay (Hauxwell et al. 2001, 2003), and newly recruiting shoots are particularly susceptible. 5.3. Biogeochemical effects 5.3.1. Observational and experimental work There is little doubt that land-derived nutrient loads alter all aspects of nutrient cycling in the receiving coastal waters and here we list just a few possible effects relevant to the response of seagrass meadows.

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Figure. 6. Mean summer light intensity (µmol photons m-2 s-1) at the water surface (corrected for surface reflectance; Peckol and Rivers 1996; 17-year average from R. Payne, Woods Hole Oceanographic Institution), and estimated light intensity reaching eelgrass (Zostera marina) leaves of established and newly recruiting shoots after interception of light due to background attenuation and summer standing stocks of phytoplankton, epiphytes, and macroalgae in 2 estuaries of Waquoit Bay subject to different nitrogen loading rates. Epiphyte and macroalgal shading were assumed to occur simultaneously and were based on intensity of incoming light after total water column attenuation. Adapted from Hauxwell et al. 2001.

Both architectural and biological features of macroalgal canopies may dramatically affect the biogeochemical environment surrounding seagrasses. Architecturally, canopies not only reduce light penetration (see discussion above), but also inhibit advective water exchange at the base of rooted plants, leading to altered sedimentwater redox conditions. Because algal mats are also biologically active components of the ecosystem, they may compete with seagrasses for nutrients (Bierzychudek et al. 1993, Krause-Jensen et al. 1996, McGlathery et al. 1997), and, through high rates of respiration, exacerbate physiological difficulties incurred by the plants in an already-reduced environment. In effect, we continue our discussion of the effect of macroalgae on seagrasses, now with a focus on the biogeochemical alterations they impose in the microenvironment around roots, rhizomes, and the base of leaves. Seagrass roots and rhizomes are adapted to exist within often anaerobic sediments. To support aerobic respiration in roots, seagrasses transport oxygen from shoots to roots (e.g., within 10-15 minutes in Zostera marina; Smith et al. 1984; Greve et al. 2003), and via diffusion from the water column at night (Pedersen et al. 1998). Prolonged anoxic conditions within sediments increased energy requirements for translocating oxygen from photosynthetically active shoots to roots and inhibited ammonium uptake by roots (Pregnall et al. 1984), and resulted in decreased rates of

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photosynthesis, leaf elongation, and number of leaves per shoot (Holmer and Bondgaard 2001). High concentrations of sulphide often occur in anoxic marine sediments, and sulphide has been identified as a potent phytotoxin for many wetland plants (Koch et al. 1990). High sulphide concentrations have been shown to decrease maximum photosynthetic rate, increase light intensity requirements for compensating growth rates, and decrease the initial slope of the photosynthesis versus irradiance curve for eelgrass (Goodman et al. 1995), causing a 55% decrease in shoot to root ratios and mortality of eelgrass shoots within 6 days of exposure (Holmer and Bondgaard 2001). Koch and Erskine (2001) demonstrated sulphide-induced mortality to Thalassia testudinum (at 6mM), but only when coupled with high salinity (55-60 PSU) and/or high temperature (35 ˚C). Though plants seemed adapted to survive periodic swings in sulphide concentrations (up to 28 d exposure to high sulphide concentration), interactions with other variables (temperature and salinity conditions commonly occurring during summer in subtropical or tropical estuaries) have been implicated in large-scale loss of seagrasses. Macroalgal canopies may drastically reduce redox conditions within sediments and the overlying water column (Valiela et al. 1992, D’Avanzo and Kremer 1994, Krause-Jensen et al. 1996), which, in addition to making a rather unhospitable environment worse for belowground seagrass material, may also affect aboveground material. Greve et al. (2003) showed that meristematic tissue was particularly sensitive to anoxia, since aboveground tissues (unlike belowground tissues) apparently lack physiological adaptations to cope with anoxia. Increased frequencies of hypoxic and anoxic events have been documented in estuaries receiving high loads of nitrogen (e.g., Waquoit Bay, D'Avanzo and Kremer 1994; Venice Lagoon, Sfriso et al. 1992), and have been attributed to high rates of respiration within algal mats. Depending on the thickness of these canopies and the height of seagrass shoots, a significant fraction of aboveground photosynthetic material may be surrounded by anoxic water. Newly recruiting shoots, in many cases, will be completely within the anoxic layer. Ammonium toxicity is another biogeochemical effect incurred within macroalgal canopies that may cause seagrass mortality (van Katwijk et al. 1997; mechanisms of toxicity to plants reviewed in Britto and Kronzucker 2002). Exceedingly high concentrations of ammonium have been measured within macroalgal canopies as a result of 1) nitrogen regeneration by the canopies and underlying sediments, and 2) reduced advective losses due to the physical structure of the canopies. In a field study, Bierzychudek et al. (1993) measured uniformly low concentrations within the water column (< 2 µM) and high concentrations within the underlying macroalgal canopy (up to 127 µM), and these data were corroborated by Hauxwell et al. (2001). Similar patterns were experimentally demonstrated by Krause-Jensen et al. (1996) and McGlathery et al. (1997).

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5.3.2. Case study - Waquoit Bay We assessed how macroalgal canopies may have altered the biogeochemical environment for eelgrass by measuring the range of concentrations of O2, NH4+, and NO3- in the water column above macroalgal canopies and above bare sediments at dawn and dusk after a sunny day in late August (described in Hauxwell et al. 2001). We found that macroalgal canopies affected concentrations of O2 and NH4+ (Fig. 7), but not NO3-. While oxygen concentrations within the top 7-cm layer of macroalgae were similar to overlying water, respiratory demands of the photosynthetically inactive bottom layers of macroalgal canopies (Peckol and Rivers 1996) resulted in a sharp decline in O2 concentrations (see also Krause-Jensen et al. 1996), both day and night within the algal canopy, with anoxia in the higher N estuary at dawn. NH4+ concentrations ranged 5-30 µM in the water column, and were much higher within the macroalgal canopy, where they ranged 6-260 µM. Though only a small percentage of leaf material of each established shoot was within the macroalgal canopy and surrounded by high concentrations of ammonium, newly recruiting shoots were entirely exposed to toxic concentrations (> 100 µM NH4+). In summary, the total mortality of eelgrass shoots within plots containing algal canopies may be attributed to physiological impairment (low [O2], high [sulphide], and/or toxic [NH4+]) and/or light limitation (Hauxwell et al. 2001). 5.4. Dynamic interactions among physical, biological, and chemical alterations imposed by nutrient loading 5.4.1. Within systems Though we separated our discussions of light and biogeochemistry, we must briefly emphasize their dynamic interactions, in particular a negative feedback for seagrasses. As nutrients are increasingly delivered to coastal waters, increases in algae result in decreased light for seagrasses, which increases energy requirements necessary to support aerobic respiration by roots in anaerobic sediments. As living and dying organic material increases, sediment oxygen demand increases, further increasing the amount of energy required to support basic seagrass physiology (Holmer and Bondgaard 2001).

5.4.2. Among systems In an effort to simplify our presentation of the various factors affecting processes that drive patterns associated with nutrient loading, we have not emphasized their dynamic nature within different types of systems.

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Figure. 7. Dawn and dusk concentrations of oxygen, ammonium, and nitrate in the water column of enclosures containing no macroalgae or thick canopies of macroalgae in 2 estuaries of Waquoit Bay subject to different nitrogen loading rates (mean ± SE). Measurements made within the 25-cm canopy in the low nitrogen estuary and the 18-cm canopy in the higher nitrogen estuary are plotted in areas under the dashed lines. Adapted from Hauxwell et al. 2001.

For example, the pattern that emerged between nutrient loading and primary producers in systems with short water residence like Waquoit Bay, did not hold for systems with longer water residence times (Fig. 4), because of modifications of the effect of phytoplankton. Throughout this chapter, we have emphasized the importance of macroalgal canopies in competitively excluding seagrasses. While this process is important in certain shallow, low energy systems, it may be less important in deeper systems or higher energy shallow systems. Hydrodynamically more-active environments are less likely to accumulate drift algae, and retain it for long periods of time. Since seagrasses able to withstand short periods of anoxia, rapid turnover of macroalgal biomass probably has less of an effect on seagrass survival in these types of systems. Hence, energy regime may affect macroalgal-seagrass interaction, with stronger interactions in lower energy environments (Bell and Hall 1997, Maciá 2000). Inherent differences in or changes to the biological setting may also affect patterns and underlying processes related to nutrient loading. Valiela and Cole (2002) showed

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that the larger the area of fringing salt marsh or mangrove swamp, the greater the production by seagrasses, and the smaller the loss of seagrasses as nitrogen load increased. Because fringing wetlands sequester nutrients, increased wetland coverage in effect decreases nutrient loads to seagrass meadows. Displacement of native plants by non-native species is a common occurrence in freshwater environments (e.g., Myriophyllum spicatum, Hydrilla verticillata, and Trapa natans displacing beds of Vallisneria americana in rivers, springs, and lakes; Caraco and Cole 2002, Hauxwell et al. 2004). Though less common in areas occupied by seagrasses, invasives have similar potential to displace native seagrasses and to alter abundance, biodiversity, and biogeochemical cycles. Examples of invasives to coastal areas include Zostera japonica (seagrass) occupying regions along the northwest coast of the United States (Larned 2003) and Caulerpa taxifolia (alga) in the Mediterranean (Ceccherelli and Cinelli 1997, Meinesz 1999; but see also Jaubert et al. 1999). Caulerpa has recently invaded the California coast of the USA, and preliminary evidence suggests it may outcompete native seagrasses Ruppia maritima and Zostera marina (Willliams and Grosholz 2002). Native and non-native animals may also reduce seagrass areal coverage. For example, extensive sea urchin grazing due to unusually high numbers of urchins (densities up to 364 m-2) caused loss of almost 1 km2 of meadows comprised mainly of Syringodium filiforme in outer Florida Bay between August 1997 and May 1998 (Rose et al. 1999). The non-native Asian mussel (Musculista senhousia) contributed to loss of eelgrass in San Diego via competition for space (Reusch and Williams 1999) and illustrates a negative feedback between nutrients and invasives on seagrasses: while in healthy eelgrass meadows, the mussel is food-limited (due to the hydrodynamics of seagrass structure, reduced water velocities result in reduced food availability), increased nutrients stimulate phytoplankton production resulting in both shading of seagrasses and stimulation of mussel growth. 5.4.3. Summary In summary, the relative importance of the various factors that have contributed to seagrass decline around the world vary with the physical, biological, and chemical setting. Our ability to observe and quantify patterns in changes of estuarine primary producers related to increased nutrients relies on our ability to classify systems based on important features (physical, or other) that affect ecological processes underlying the patterns. While in certain settings, the relationship between nutrient loading and seagrass loss seems relatively straightforward, we must emphasize that additional ecological complexities may alter processes and patterns. The changes in abundance and assemblages of primary producers that accompany increased nutrient loads (shift from more refractory seagrass material to more labile algal biomass), have higher order ecological effects as well. These include ecosystem-scale changes to overall total primary production and the various fates of that production (herbivory, export, burial, decomposition), alterations in nutrient and carbon cycling, changes to biogeochemical cycles, and effects on consumers (see following Chapters).

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6. NOTES Portions of this manuscript closely follow Florida Sea Grant publication (SGEB-55). "Nutrients and Florida's coastal waters: the links between people, increased nutrients, and changes to coastal aquatic systems," and the Hauxwell et al. 2001 publication referenced, and we thank Thomas Frazer, Charles Jacoby and Just Cebrian for their contributions. JH was supported by the Wisconsin Department of Natural Resources during preparation of the manuscript. We thank Gabrielle Tomasky, Stefanno Mazzilli, and Marci Cole for contributing data and figures. 7. REFERENCES Abal, E. G., & Dennison, W. C. (1996). Seagrass depth range and water quality in southern Moreton Bay, Queensland, Australia. Marine and Freshwater Research, 47, 763-771. Backman, T. W., & Barilotti, D. C. (1976). Irradiance reduction: effects on standing crops of the eelgrass, Zostera marina, in a coastal lagoon. Marine Biology, 34, 33-40. Bannister, T. T. (1974). Production equations in terms of chlorophyll concentration, quantum yield, and upper limit to production. Limnology and Oceanography, 19, 1-12. Bayley, S., Stotts, V. D., Springer, P. F., & Steenis, J. (1978). Changes in submerged aquatic macrophyte populations at the head of Chesapeake Bay, 1958-1975. Estuaries, 1, 73-84. Bell, S. S., & Hall, M. O. (1997). Drift macroalgal abundance in seagrass beds: investigating large-scale associations with physical and biotic attributes. Marine Ecology Progress Series, 147, 277-283. Bester, K. (2000). Effects of pesticides on seagrass beds. Helgoland Marine Research, 54, 95-98. Bierzychudek, A., D’Avanzo, C., & Valiela, I. (1993). Effects of macroalgae, night and day, on ammonium profiles in Waquoit Bay. Biological Bulletin, 185, 330-331. Borum, J. (1985). Development of epiphytic communities on eelgrass (Zostera marina) along a nutrient gradient in a Danish estuary. Marine Biology, 87, 211-218. Borum, J. (1987). Dynamics of epiphyton on eelgrass (Zostera marina L.) leaves: relative roles of algal growth, herbivory, and substratum turnover. Limnology and Oceanography, 32, 986-992. Borum, J., & Wium-Andersen, S. (1980). Biomass and production of epiphytes on eelgrass (Zostera marina L.) in the Øresund, Denmark. Ophelia, Supplement, 1, 57-64. Britto, D. T., & Kronzucker, H. J. (2002). NH4+ toxicity in higher plants: a critical review. Journal of Plant Physiology, 159, 567-584. Brush, M. J., & Nixon, S. W. (2002). Direct measurements of light attenuation by epiphytes on eelgrass Zostera marina. Marine Ecology Progress Series, 238, 73-79. Burkholder, J. M., Glasgow, Jr., H. B., & Cooke, J. E. (1994). Comparative effects of water-column nitrate enrichment on eelgrass Zostera marina, shoalgrass Halodule wrightii, and widgeongrass Ruppia maritima. Marine Ecology Progress Series, 105, 121-138. Burkholder, J. M., Mason, K. M., & Glasgow, Jr., H. B. (1992). Water-column nitrate enrichment promotes decline of eelgrass Zostera marina: evidence from seasonal mesocosm experiments. Marine Ecology Progress Series, 81, 163-178. Caraco, N., & Cole, J. (2002). Contrasting impacts of a native and alien macrophyte on dissolved oxygen in a large river. Ecological Applications, 12, 1496-1509. Caraco, N., Cole, J., & Likens, G. E. (1990). A comparison of phosphorus immobilization in sediments of freshwater and coastal marine sediments. Biogeochemistry, 9, 277-290. Ceccherelli, G., & Cinelli, F. (1997). Short-term effects of nutrient enrichment of the sediment and interactions between the seagrass Cymodocea nodosa and the introduced green alga Caulerpa taxifolia in a Mediterranean bay. Journal of Experimental Marine Biology and Ecology, 217, 165-177. Charpy-Robaud, C., & Sournia, A. (1990). The comparative estimation of phytoplanktonic, microphytobenthic, and macrophytobenthic primary production in the oceans. Marine Microbial Food Webs, 4, 31-57. Cloern, J. E. (2001). Our evolving conceptual model of the coastal eutrophication problem. Marine Ecology Progress Series, 210, 223-253. Coffaro, G., & Bocci, M. (1997). Resources competition between Ulva rigida and Zostera marina: a quantitative approach applied to the Lagoon of Venice. Ecological Modeling, 102, 81-95.

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Onuf, C. P. (1994). Seagrasses, dredging and light in Laguna Madre, Texas, U.S.A. Estuarine, Coastal and Shelf Science, 39, 75-91 Onuf, C. P. (1996). Seagrass responses to long-term light reduction by brown tide in upper Laguna Madre, Texas: distribution and biomass patterns. Marine Ecology Progress Series, 138, 219-231. Orth, R. J., Fishman, J. R., Wilcox, D. J., & Moore, K. A. (2002). Identification and management of fishing gear impacts in a recovering seagrass system in the coastal bays of the Delmarva Peninsula, USA. Journal of Coastal Research, 37, 111-129. Pasquilini, V., Pergent-Martini, C., & Pergent, G. (1999). Environmental impact identification along the Corsican coast (Mediterranean Sea) using image processing. Aquatic Botany, 65, 311-320. Peckol, P., DeMeo-Anderson, B., Rivers, J., Valiela, I., Maldonado, M., & Yates, J. (1994). Growth, nutrient uptake capacities and tissue constituents of the macroalgae Cladophora vagabunda and Gracilaria tikvahiae related to site-specific nitrogen loading rates. Marine Biology, 121, 175-185. Peckol, P., & Rivers, J. S. (1996). Contribution by macroalgal mats to primary production of a shallow embayment under high and low nitrogen-loading rates. Estuarine, Coastal and Shelf Science, 43, 311325. Pedersen, O., Borum, J., Duarte, C. M. & Fortes, M. D. (1998). Oxygen dynamics in the rhizosphere of Cymodocea rotundata. Marine Ecology Progress Series, 169, 283-288. Peterken, C. J., & Conacher, C. A. (1997). Seed germination and recolonization of Zostera capricorni after grazing by dugongs. Aquatic Botany, 59, 333-340. Pickett, S. T. A. (1989). Space-for-time substitution as an alternative to long-term studies. In G. E. Likens (Ed.), Long-term studies in ecology: approaches and alternatives (pp. 110-135). New York: Springer-Verlag. Pregnall, A. M., & Rudy, P. P. (1985). Contribution of green macroalgal mats (Enteromorpha spp.) to seasonal production in an estuary. Marine Ecology Progress Series, 24, 167-176. Pregnall, A. M., Smith, R. D., Kursar, T. A., & Alberte, R. S. (1984). Metabolic adaptations of Zostera marina (eelgrass) to diurnal periods of root anoxia. Marine Biology, 83, 141-147. Pulich, Jr., W., Blair, C., & White, W. A. (1997). Current status and historical trends of seagrasses in the Corpus Christi Bay National Estuary Program Study Areas. Publication number CCBNEP - 20. Texas Natural Resource Conservation Commission, TX, USA. pp. 1-57. Reusch, T. H. B., & Williams, S. L. (1999). Macrophyte canopy structure and the success of an invasive marine bivalve. Oikos, 84. 398-416. Robbins, B. D. (1997). Quantifying temporal change in seagrass areal coverage: the use of GIS and low resolution aerial photography. Aquatic Botany, 58, 259-267. Robertson, A. I., & Mann, K. H. (1984). Disturbance by ice and life-history adaptations of the seagrass Zostera marina. Marine Biology, 80, 131-141. Rose, C. D., Sharp, W. C., Kenworthy, W. J., Hunt, J. H., Lyons, W. G., Prager, E. J., Valentine, J. F., Hall, M. O., Whitfield, P. E., & Fourqurean, J. W. (1999). Overgrazing of a large seagrass bed by the sea urchin Lytechinus variegatus in Outer Florida Bay. Marine Ecology Progress Series, 190, 211222. Ruiz, J. M., Perez, M., & Romero, J. (2001). Effects of fish farm loadings on seagrass (Posidonia oceanica) distribution, growth and photosynthesis. Marine Pollution Bulletin, 42, 749-760. Rybicki, N. B., & Carter, V. (1986). Effect of sediment depth and sediment type on the survival of Vallisneria americana Michx grown from tubers. Aquatic Botany, 24, 233-240. Ryther, J. H. (1969). Photosynthesis and fish production in the sea. Science, 166, 72-76. Sand-Jensen, K. (1977). Effects of epiphytes on eelgrass photosynthesis. Aquatic Botany, 3, 55-63. Sand-Jensen, K., & Borum, J. (1991). Interactions among phytoplankton, periphyton, and macrophytes in temperate freshwaters and estuaries. Aquatic Botany, 41, 137-175. Seddon, S., Connolly, R. M., & Edyvane, K. S. (2000). Large-scale seagrass dieback in northern Spencer Gulf, South Australia. Aquatic Botany, 66, 297-310. Seitzinger, S. P., Kroeze, C., Bouwman, A. F., Caraco, N., Dentener, F., & Styles, R. V. (2002). Global patterns of dissolved inorganic and particulate nitrogen inputs to coastal systems: recent conditions and future projections. Estuaries, 25, 640-655. Sfriso, A. (1987). Flora and vertical distribution of macroalgae in the lagoon of Venice: a comparison with previous studies. Giornale Botanico Italiano, 121, 69-85. Sfriso, A., Pavoni, B., Marcomini, A., & Orio, A. A. (1992). Macroalgae, nutrient cycles, and pollutants in the lagoon of Venice. Estuaries, 15, 517-528. Short, F. T. (1988). Effects of sediment nutrients on seagrasses: literature review and mesocosm experiment. Aquatic Botany, 27, 41-57. Short, F. T., Burdick, D. M., & Kaldy III, J. E. (1995). Mesocosm experiments quantify the effects of eutrophication on eelgrass, Zostera marina. Limnology and Oceanography, 40, 740-749.

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Short, F. T., Burdick, D. M., Wolf, J. S., & Jones, G. E. (1993). Eelgrass in estuarine research reserves along the East Coast, U.S.A., Part I: Declines from pollution and disease and Part II: Management of eelgrass meadows. National Oceanic and Atmospheric Administration, Coastal Ocean Program Publication, Rockville, Maryland. Pp. 1-83, M1-M24. Short, F. T, & Wyllie-Echeverria, S. (1996). Natural and human-induced disturbance of seagrasses. Environmental Conservation, 23, 17-27. Sigua, G. C., & Tweedale, W. A. (2003). Watershed scale assessment of nitrogen and phosphorus loadings in the Indian River Lagoon basin, Florida. Journal of Environmental Management, 67, 363372. Silberstein, K., Chiffings, A. W., & McComb, A. J. (1986). The loss of seagrass in Cockburn Sound, Western Australia. III. The effect of epiphytes on productivity of Posidonia australis Hook. F. Aquatic Botany, 24, 355-371. Smith, A. (1997). Seagrass protected in NSW estuaries. Fisheries-NSW, 1, 19-20. Smith, R. D., Dennison, W. C., & Alberte, R. S. (1984). Role of seagrass photosynthesis in root aerobic processes. Plant Physiology, 74, 1055-1058. Smith, S. V. (1981). Marine macrophytes as a global carbon sink. Science, 211, 838-840. Street, G. T., Montagna, P. A., & Parker, P. L. (1997). Incorporation of brown tide into an estuarine food web. Marine Ecology Progress Series, 152, 67-78. Taylor, D., Nixon, S., Granger, S., & Buckley, B. (1995a). Nutrient limitation and the eutrophication of coastal lagoons. Marine Ecology Progress Series, 127, 235-344. Taylor, D. I., Nixon, S. W., Granger, S. L., Buckley, B. A., McMahon, J. P., & Lin, H.-J. (1995b). Responses of coastal lagoon plant communities to different forms of nutrient enrichment - a mesocosm experiment. Aquatic Botany, 52, 19-34. Tomasko, D. A., Bristol, D. L., & Ott, J. A. (2001). Assessment of present and future nitrogen loads, water quality, and seagrass (Thalassia testudinum) depth distribution in Lemon Bay, Florida. Estuaries, 24, 926-938. Tomasky, G., Barak, J., Valiela, I., Behr, P., Soucy, L., & Foreman, K. (1999). Nutrient limitation of phytoplankton growth in Waquoit Bay, Massachussetts, USA: a nutrient enrichment study. Aquatic Ecology, 33, 147-155. Touchette, B. W., Burkholder, J. M., & Glasgow, Jr, H. B. (2003). Variations in eelgrass (Zostera marina) morphology and internal nutrient composition as influenced by increased temperature and water column nitrate. Estuaries, 26, 142-155. Twilley, R. R., Kemp, W. M., Stave, K. W., Stevenson, J. C., & Boynton, W. R. (1985). Nutrient enrichment of estuarine communities. 1. Algal growth and effects on production of plants and associated communities. Marine Ecology Progress Series, 23, 179-191. U.S. Geological Survey. (1999). The quality of our nation’s waters - nutrients and pesticides. U.S. Geological Survey Circular 1225. Pp. 1-82. Valiela, I., & Cole, M. L. (2002). Comparative evidence that salt marshes and mangroves may protect seagrass meadows from land-derived nitrogen loads. Ecosystems, 5, 92-102. Valiela, I., Collins, G., Kremer, J., Lathja, K., Geist, M., Seely, B., Brawley, J., & Sham, C.H. (1997a). Nitrogen loading from coastal watersheds to receiving estuaries: new methods and application. Ecological Applications, 7, 358-380. Valiela, I., Foreman, K., LaMontagne, M., Hersh, D., Costa, J., Peckol, P., DeMeo-Anderson, B., D'Avanzo, C., Babione, M., Sham, C., Brawley, J., & Lathja, K. (1992). Couplings of watersheds and coastal waters: sources and consequences of nutrient enrichment in Waquoit Bay, Massachusetts. Estuaries, 15, 443-457. Valiela, I., Geist, M., McClelland, J., & Tomasky, G. (2000a). Nitrogen loading from watersheds to estuaries: verification of Waquoit Bay Nitrogen Loading Model. Biogeochemistry, 49, 277-293. Valiela, I., McClelland, J., Hauxwell, J., Behr, P. J., Hersh, D., & Foreman, K. (1997b). Macroalgal blooms in shallow estuaries: Controls and ecophysiological and ecosystem consequences. Limnology and Oceanography, 42, 1105-1118. Valiela, I., Tomasky, G., Hauxwell, J., Cole, M. L., Cebrián, J., & Kroeger, K. D. (2000b). Operationalizing sustainability: management and risk assessment of land-derived nitrogen loads to estuaries. Ecological Applications, 10, 1006-1023. van Katwijk, M. M., Vergeer, L. H. T., Schmitz, G. H. W., & Roelofs, J. G. M. (1997). Ammonium toxicity in eelgrass Zostera marina. Marine Ecology Progress Series, 157, 159-173. Vasiliu. F. (1996). The Black Sea. In W. Schramm & P. H. Nienhuis (Eds.), Marine benthic vegetation (pp. 435-447). Berlin: Springer-Verlag.

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Vitousek, P. M., Aber, J., Howarth, R. W., Likens, G. E., Matson, P. A., Schindler, D. W., Schlesinger, W. H., & Tilman, G. D. (1997). Human alteration of the global nitrogen cycle: causes and consequences. Ecological Applications, 7, 737-750 Williams, S. J., Dodd, D., & Gohm, K. K. (1991). Coasts in crisis. United States Geological Survey Circulation 1075. Williams, S. L., & Grosholz, E. D. (2002). Preliminary reports from the Caulerpa taxifolia invasion in southern California. Marine Ecology Progress Series, 233, 307-310. Zaitsev, Y. P. (1992). Recent changes in the trophic structure of the Black Sea. Fisheries Oceanography, 1, 180-189. Zieman, J. C., Fourqurean, J. W., & Frankovich, T. A. (1999). Seagrass die-off in Florida Bay: long-term trends in abundance and growth of turtle grass, Thalassia testudinum. Estuaries, 22, 460-470.

8. AFFILIATIONS J. Hauxwell: Wisconsin Department of Natural Resources, Department of Natural Resources Research Center, 1350 Femrite Drive, Monona, WI 53761, USA. I. Valiela: Boston University, Marine Program, Marine Biology Laboratory, Woods Hole, MA 02543, USA.

MOGENS R. FLINDT, JOÁO NETO, CARL L. AMOS, MIGUEL A. PARDAL, ALESSANDRO BERGAMASCO, CATHRINE B. PEDERSEN & FREDE Ø. ANDERSEN

PLANT BOUND NUTRIENT TRANSPORT. MASS TRANSPORT IN ESTUARIES AND LAGOONS. 1. INTRODUCTION In many regions of the world external nutrient loading to estuaries has recently been decreasing. These estuaries are only beginning to adjust to this change. The change is visible as better water quality. The period where the phytoplankton is growth limited has become longer, resulting in a better light climate at the bottom with the potential for recolonization of benthic macrophytes. This phase of recovery often shows a nonlinear behaviour and can be longer than expected (Borum 1996, 1997). National monitoring programs have been developed with the purpose of following the expected recovery phase. Nutrient loading, water quality and the export of nutrients from these systems are the main focus of this monitoring. For this reason, increasing attention has been paid to nutrient mass balances at the outer boundary of lagoons and estuaries. The measurements of estuarine mass balances have traditionally been limited to dissolved inorganic nutrients (ammonia, nitrite, nitrate, phosphate) and fine particulate matter fraction trapped on filters (e.g. glass fiber filters). Often no distinction is made between living and detrital particulate matter. This chapter will show that most of these nutrient mass balances are incomplete. The reason is that in shallow productive micro- and meso-tidal estuaries plant bound nutrient transport is essential for the nutrient mass balance (Flindt et al. 1997a, 1999, Salomonsen et al. 1997, 1999). Why is plant bound nutrient transport not included in the mass balances? A part of the explanation is that the plant matter accumulates at the bottom – and the transport afterward takes place as bedload transport. Although Odum et al. (1979) pointed out that import/export studies have failed to account for transport of particulate loads on or near the bed of estuaries, this transport is still often neglected. Another explanation is that when a piece of a macroalgae is infrequently trapped in a water sample of a few litres, this piece will usually be removed, because the sample is considered non-representative and would introduce a high variability among the samples. The inclusion of a 50 g wwt piece of Ulva in a 5 l water sample may increase the total nitrogen and total phosphorus concentration by a factor of 50. The problem with monitoring nutrient import/export from estuaries is primarily related to the sampling strategy. Instead of only filtrating 1-5 l water samples through a glass fiber filter, we also need to filter the water column horizontally using nets with a mesh size of 0.5-1.0 cm in diameter, and the filtered amount has to be between 100 to 10,000 m3. The sampling strategy should also reflect the different depths in which the plants are transported in the water column. 93 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 93-128. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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During field work in the lagoon of Venice, we were watching drifting ephemeral macroalgae at our stations. This generated the question about how much of the nutrient export that was transported as plant bound nutrients, and hereby questioned earlier mass balances. Therefore studies were made where all fractions of nutrients, both dissolved and particulate, were measured. Three estuaries with different tidal amplitude and loading were compared. The estuaries were: Venice Lagoon (Italy), Mondego Estuary (Portugal), and Roskilde Fjord (Denmark). There is substantial evidence in the literature for transport of macrophytic material from estuaries and coastal areas to the open ocean. For example Menzies et al. (1967) and Menzies & Rowe (1969) observed export of seagrass material to depths greater than 3,000 m in the deep sea, and Greenway (1976) estimated that 9.5% of the weekly production of Thalassia was exported from Kingston Harbour, Jamaica. Unbalanced energy budgets for salt marsh ecosystems led Odum (1968) to propose the “Outwelling Hypothesis”. This hypothesis has inspired many investigations and reviews (e.g. Nixon 1980, Dame 1989, 1994, Dame & Allen 1996). Most of these studies focus on salt marsh dynamics, and to our knowledge not much attention has been given to shallow estuaries. There is therefore a need for describing the transport patterns for macrophytic material and to quantify the importance of this plant transport to the nutrient budgets of estuaries. Here we (1) describe the modes of transport of different forms of macrophytes subject to unidirectional flows, (2) determine the threshold velocity for initiation of macroalgal transport, (3) examine the relationship between plant transport velocity and mean current speed, (4) describe settling rates of macrophytic material, (5) compare growth, loss, and plant transport rates, and (6) exemplify mass balances at outer boundaries of estuaries including plant bound nutrient transport. The overall goal is to establish an ecological model able to describe plant bound nutrient transport as an integrated part of a full scale ecological model for estuaries.

2. PLANT TRANSPORT PATTERNS MEASURED IN THE FIELD A major fraction of submerged estuarine plants is found either living unattached to the bed as for example Ulva sp. and Chaetomorpha sp., loosely attached as Enteromorpha sp., Ceramium sp. and Polysiphonia sp., or anchored to the site by holdfasts as for example Fucus sp. or by roots as seagrasses. All types are observed drifting in the water column or moving as bedload transport within estuaries and lagoons. However, almost nothing is known about the erosion threshold, settling rates, or transport pattern of these submersed aquatic plants. To quantify the horizontal transport of macroalgae and rooted macrophytes in Venice Lagoon, nets with a height of 1.5 m, width 1.0 m, and vertically divided into 5 sections: 0-30 cm, 30-60 cm, 60-90 cm, 90-120 cm, 120-150 cm were placed with the opening perpendicular to the main current direction to trap plant material drifting in different depths of the water column. The results are shown in Figure 1.

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The resulting horizontal plant transport during a falling tide showed that Chaetomorpha sp., Ulva sp. and Zostera sp. had different drift patterns in the water column. Most of the Zostera material (89 %) was drifting close to the water surface, while most of the Chaetomorpha material (65 %) was transported along the bottom. Ulva also showed a tendency to be transported near the bottom (Fig. 1). The different transport patterns between the plants may be explained by the air filled aerenchymatic tissue in the Zostera leaves which create buoyancy relative to the water, while Chaetomorpha most often has a higher specific mass, because these mats are often inhabited by various types of crustaceans (crabs, shrimps etc.) that use this micro environment as a shelter. Zostera material caught in the lowest sections were either dark senescent leaves where the aerenchym was filled with water, or whole living plants with roots, rhizomes and leaves. Zieman et al. (1979) observed that two species of seagrasses were transported in different parts of the water column. Syringodium filiforme, which have large air filled lacunae, were nearly always floating at the surface, while Thalassia testudinum usually sank to the bottom. Leaves of the latter species only float when young and green, while they become denser on senescence and more coated with calcareous epiphytes. Other nets with a hydraulic cross section of 1 or 2 m2 were developed to trap plants over the whole water column. To prevent the current inside the net from being restricted by the trapped plant matter, the nets were up to 15 m in length with a mesh size of 0.9 cm. The current velocity and direction were measured concurrently permitting a correlation between current speed and plant catch. The results from all field measurements in 1995 (Flindt et al. 1997a) are presented in Figure 2. Positive correlations between plant transport and current velocity were found for both Ulva rigida (r2= 0.87, P<0.01) and Chaetomorpha

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aerea. (r2= 0.82, P<0.01), while the transport of Zostera spp. (r2=0.082) was not significantly related to the water current.

Figure 2. Correlation between current velocity and catchment of dominating plant material Venice Lagoon, Malamocco site near the inlet, May 1995. (Flindt et al. 1997a, reprinted with permission from Elsevier)

The correlation between drift of macroalgae and the current velocity indicated that advective transport was a major mechanism for plant loss from the studied area. The advective transport of rooted macrophytes like Zostera sp. did not show any clear correlation to current velocity. This is most likely because the sloughed leaves from Zostera, in contrast to the macroalgae, are floating at the water surface. Here the drift direction is influenced by the wind. It may also be expected that the sloughing of leaves from rooted macrophytes are not only controlled by current velocities but are also dependent on the age of the leaves, the production, the physiological condition of the plant, and various environmental conditions like anoxia and wind or wave created resuspension. Many field campaigns in different estuaries have been carried out over full tidal cycles, so complete import-export transport patterns and net balances could be established. Such a campaign is shown on Figure 3, where measurements were made close to the Lido entrance in the Lagoon of Venice. The station was 1.5 meter deep. The measurements started on a spring tide with an outgoing current where the tidal amplitude was about 80 cm. As the current velocity increased, the plant transport

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started, and the earlier established correlation between current speed and plant catch was evident, because the highest current speed occurred between 2-4 h.

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This example also underlines the general tendency in most of the campaigns, that the loss of plant material across the boundary is very high. No more than 10-15% of the exported plant matter returned with the subsequent flooding tide where the water returned to the lagoon. Only 10% of the transported Ulva sp. returned, about 4% of Chaetomorpha sp. and mixed red algae (Gracilaria sp, Ceramium sp., Polysiphonia sp.) returned, while close to 25% of the leaves from rooted plants (Zostera marina, Zostera noltii, Cymodocia sp.) returned to the lagoon. These differences can be explained by the differences in settling rates among these macrophytes. In conclusion, about 90 % of the plant mass was permanently lost to the coastal area – in this example the Adriatic Sea. Although there may be variations in the import-export balances of plant transport, the above discussed example may be considered as representative. The variations in the import-export balances are primarily dependent on: 1.

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The bathymetric conditions around the outer boundary. If the transport is from a shallow area to a deeper coastal area almost all transported plant matter is lost. This is the case in Venice Lagoon, which has an average depth of about 1 m, while the depth of the Adriatic Sea is about 20-50 m. The current patterns outside the boundary, where the water may be more or less stagnant – depending on local wind and coastal current dynamics.

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3. MACROALGAE EROSION THRESHOLDS AND SETTLING RATES. To produce correct mass balances and to develop predictive models of macrophyte advection and transport in estuaries, it is important to take the plant bound export and dispersion of plant material into account. To do this, we need to know the threshold for macroalgae mobilization, the mode of transport as a function of current velocity, and the settling rates of the different dominating plant material. The Southampton Oceanographic Lab Carousel (Amos et al 2000, Thomson & Amos 2002) was used in an experimental study on erosion and transport of macroalgal material. The carousel was filled with prefiltered estuarine water and a 4 cm thick sediment bed was created. Four species of macroalgae (Ulva lactuca, Cladophora sp., Chaetomorpha linum and Ceramium rubrum) were collected from the same site as the sediment. Epiphytes and inorganic particulate matter were removed from the surface of the algae to eliminate possible differences in densities and surface to volume ratio between the specimens. Afterwards, the macroalgae were placed unattached on the sediment surface. Air bubbles were removed. Mean current speed was increased in steps of 0.5 cm s-1 at 10 minutes intervals to 11.0 cm s-1. The current speed at the point where the plants started to move was termed the traction threshold. In all experiments the plants started to move as bedload after which they lifted into suspension at higher current velocities. The current speed at which the plants lifted into suspension was the suspension threshold. Two pieces of Ulva sp., differing in weight by a factor of 10, were monitored at increasing current speeds (to 9 cm s-1). The smallest piece had a traction threshold at a current speed of 2.7 cm s-1 and a suspension threshold of 7.0 cm s-1. The largest piece of Ulva sp. had a slightly lower traction threshold of 2.0 cm s-1, the speed of this algae increased with the current speed in a similar fashion (Fig. 4A). However, the differences in threshold velocity and transport rate between the two sizes were not significant. The traction threshold for Chaetomorpha sp. (Fig. 4B) varied between 1.7 and 3.0 cm s-1 and was thus similar to Ulva. None of the three Chaetomorpha pieces passed into suspension but continued to tumble along the bed at all current speeds. All three data sets fitted a logarithmic curve (r2= 0.963 to 0.986). The filamentous red algae Ceramium (Fig. 4C) showed a transport pattern and traction thresholds from 1.7 to 3.0 cm s-1. Two pieces of Ceramium showed a suspension threshold of 9 cm s-1. The three data sets also fitted well to a logarithmic curve (r2= 0.934 to 0.986). The three pieces of similar size of Cladophora (Fig. 4D) started to move along the bottom at a current speed between 1.7 and 3.0 cm s-1 similar to the red algae Ceramium. The first piece of Cladophora came into suspension at 5.0 cm s-1, the other two at 9 cm s-1. All three data sets fitted well a logarithmic curve (r2= 0.947 to 0.992). When these experiments were initiated we expected a broad range of traction thresholds. The hypothesis was that different morphological types of macroalgae: sheet-formed (eg. Ulva sp.), unbranched filamentous (eg. Chaetomorpha sp. and branched filamentous (eg. Ceramium sp.) would need hydraulically different traction due to their differences in surface area. The results did not support these expectations.

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A summary of the erosion experiments is presented in Figure 5. All species of macroalgae began to move at very low current speeds (2 to 4 cm s-1). Once in motion, the algae moved slowly along the sediment bed as traction. With increasing current speed, the algae began to tumble along the bed and at a current speed from 6 to 10 cm s-1 three of the algae species were in suspension, only Chaetomorpha sp. did not go into suspension during the experiment. When in suspension the velocity of the algae was the same as the current speed. A major fraction of submerged estuarine plants is found either unattached at the bed as for example Ulva sp. and Chaetomorpha sp. or loosely attached as Enteromorpha sp., Ceramium sp. and Polysiphonia sp. All these types of algae are observed drifting in the water column or moving as bedload transport within Venice Lagoon and at The Solent estuary near Southampton. When these drifting macroalgae were observed at new areas they were earlier mistakenly considered as algae produced at the site. The plants release oxygen from their tissue, which sometimes becomes trapped as attached bubbles thereby changing the density of the macroalgae, causing them to float. Buoyancy and lift is thus responsible for part of the erosion of macroalgae.

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Figure 5. Transport pattern (bed load or suspended) for four algae species: Ulva, Chaetomorpha, Ceramium, and Cladophora. Results from all carousel experiments (see text). Error lines indicate standard deviation of mean.

The present experiments were undertaken without associated oxygen bubbles. Nevertheless, the macroalgae started moving at very low current speeds and at 9 cm s1 all, regardless size and weight, moved into suspension. It thus appears that the unattached macroalgae are highly mobile and respond to even low steady flows. This fact can have a great impact on how we interpret the biology of unattached macroalgae in shallow estuaries, especially in estuaries influenced by tide or meteorological forcing. Furthermore, the plant transport can play an important role in the nutrient balance of these estuaries through the export and/or import of large amounts of plant material by tides. In Venice Lagoon a significant correlation was observed between current velocity and the catch of Ulva rigida and Chaetomorpha aeraea. By extrapolating the regression lines to zero catch, it was possible to find the current velocity for the erosion threshold. The field erosion thresholds were 16 cm s-1 for Ulva rigida and 6 cm s-1 for Chaetomorpha aeraea were a little above the measured traction thresholds of the laboratory study, where both macroalgae started to move at 2-4 cm s-1. The difference in threshold velocity could be caused by differences in bed roughness in the field and laboratory. No doubt that bed roughness creates variations in threshold. The sediment in Venice was silty sand, while the laboratory bed was muddy silt. In nature, macrophytes tumbling along the bottom initially may be trapped by pools, stones or in the rooted vegetation when the current velocities are low. The differences may also be due to more imprecise measurement in the field.

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4. ATTACHED MACROALGAE SLOUGHING RATES It is obvious that unattached macroalgae may be transported by the currents, but a major part of macroalgae grows fixed to shells or stones. In this section we describe results for attached macroalgae like Enteromorpha intestinalis, Porphyra sp., Ulva sp. and Fucus sp. on sloughing rates and later these are compared with other loss rates like sporulation and grazing. Sloughing rates were measured at controlled conditions in a laboratory carousel, where current velocities could be varied. Ten specimens of each macroalgal species attached to shells or stones were placed evenly distributed in the lab-carousel. The current speed was increased in steps from still water until the macroalgae including the attachment started to move. Between each speed increment the sloughed macroalgae material was collected manually. The biomasses of the sloughed macroalgal material as well as the remaining attached parts left at the end of each experiment were measured. Figure 6 shows the results from the sloughing experiment for four species of macroalgae. The first attached macroalgal species to loose biomass were Porphyra and Ulva which at a current speed of 12 cm s-1 had lost 3-4% of their biomass, increasing to 4-7% at 15 cm s-1. At a current speed of 40 cm s-1 the loss was 40 - 50%. The sloughing of Enteromorpha was slightly less affected by the current speed. For this macroalgae the sloughing was 1% of the biomass at 20 cm s-1, increasing to 20% at 35 cm s-1. Above this speed Enteromorpha and shells started to move. The most resistant species to sloughing was Fucus that started to lose biomass at a speed of 30 cm s-1, increasing to 1.5% at a speed of 35 cm s-1. The maximum loss measured on Fucus was 4% at a relatively high current speed of 50 cm s-1. Furthermore, Fucus was attached to large stones that did not move with the highest current speeds. Linear regression between current speed and accumulated sloughing rates were made for each species and presented in Table 1. The statistics show that macroalgae sloughing rates have a high dependency on current speed and that the differences in thresholds are much bigger than the measured variation in erosion thresholds for unattached macroalgae. Table 1. Sloughing threshold and linear regression between sloughing and current speed for four species of macroalgae

Species Porphyra sp.

Threshold (cm s-1) 9.5

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11.5

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14.0

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29.2

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r2

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Figure 6. Relationship between accumulated sloughed macroalgae biomass (% of total biomass) and current speed. The study was made in the laboratory at Southampton Oceanography Centre, May 2001

Assuming that the force created by the current on the macroalgae increases linearly with current speed, it becomes possible to predict when the attached species will start to lose mass, and at which current speed all macroalgal matter becomes lost (Table 2). Compared to the erosion threshold for unattached macroalgae, the loss by sloughing requires a much higher current speed making the attached macroalgae less sensitive to transport losses. However, the ephemeral macroalgae started to loose biomass at current speeds of 10-15 cm s-1 and they had a considerable loss at 30-50 cm s-1. For Fucus sp. and comparable macroalgae with more firm structural mass and a stronger holdfast, the sloughing rate will never become a major loss rate, because an unrealistic high current speed of 655 cm s-1 would be required to make loss rates by sloughing important.

Table 2. Current speeds for beginning and maximum sloughing rates for four macroalgae species

Species Porphyra sp. Ulva sp. Enteromorpha sp. Fucus sp.

Current speed (cm s-1) where sloughing rate begin is maximal 9.5 88 11.5 79 14.0 127 29.2 655

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These results are based on conservative assumptions because all measurements were carried out on healthy plants without ghost tissue, which may not be representative for an average macroalgal community. The experiments were also carried out so no thalli contained air bubbles. There may be many explanations for the presence or absence of macroalgae in different areas at certain times of the growth season, but the above results suggest that the presence of e.g. Enteromorpha sp. in the intertidal becomes limited to period where they can realize nearly maximum growth rates. In sub optimal periods, the biomass loss due to sloughing will result in decreases of biomass and extinction. This seems the case in the Mondego Estuary where Enteromorpha is absent in the channel with the highest current velocities of 1 m s-1, but once in a while is able to manage to grow in the shallow hydrodynamically protected intertidal areas. The experimental studies on sloughing were verified by qualitative field studies at different tidal currents. Here we observed that Ulva, Enteromorpha and Porphyra were transported in the estuary attached to shells and small stones, while we never encountered Fucus sp. transported. Enteromorpha sp. was the algae that had the highest transportation frequency as attached and alive. This may be explained as a combined morphological and physiological phenomenon where photosynthetic produced oxygen easily becomes trapped in the tube-formed thallus. When this happens, the buoyancy of the algae increases resulting in the ability to compensate for the extra mass the shells or small stones exert. Thus, much lower current velocities are needed to bring the plant into the water column. Fucus, attached to stones, may only be transported at very high current speeds. This occurs as bed load transport which is difficult to observe in turbid estuaries. 5. MACROALGAE SETTLING RATES To model transportation of macrophytes, both a parameterization of the erosion process as well as the settling process are necessary. This section describes the diversity among different morphological types of macrophytes with respect to settling rates. The dominant macrophytes living in Venice Lagoon were collected for settling studies in the laboratory. Epiphytes and sediment particles were gently removed by washing in estuarine water. All settling rates were measured on living plant material except the leaves from the rooted vegetation, which only settle when the aerenchym is filled with water. This occurs at senescence, and therefore dead leaves were used for settling measurements on rooted plants. The settling rate was measured for various sizes (1 – 10 cm2) of circular and square discs of Ulva lactuca. The settling experiments were carried out in a plexiglass tube with a height of 2.2 m and 20 cm in diameter. The tube was filled with prefiltered estuarine water and kept at a temperature of 24 oC. Plant material touching the walls of the cylinder was excluded. The settling mode of circular and square sheets of algae was either on the edge or in a horizontal position. The experiments were repeated with other species of macroalgae and leaves from rooted vegetation in order to determine the variation in settling rates. The results are presented in Fig. 8.

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Figure 8. Settling rates for different sizes and forms of Ulva lactuca. The size is indicated by the area. Circles and squares indicate circular and square discs of the algae, respectively. No significant differences between forms or size classes were observed.

It was clear that the settling rates of Ulva were independent of size even over one order of magnitude. Square pieces of Ulva sp. settled with rates of 17-70 m h-1 and circular pieces with rates of 17-75 m h-1, thus there seemed to be no significant differences in the settling rate between the two shapes. Ulva began settling by random motion and continued to settle in two modes. Settling with the thallus in a horizontal position was slower (35 m h-1) than settling with the thallus on its edge (60 m h-1).

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Figure 9. Settling rates for four macrophyte species with different morphology. Size classes are indicated by weight (g wwt). There was no significant change in settling rates within the range of one decade of wet weight for any of the macrophytes. Studies were performed in Venice, August, 1998.

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Our results showed that settling rates of different sized pieces of macroalgae and a seagrass species was relatively independent of shape or size (Fig. 8, 9). The comparison of settling rates for different species of macroalgae and rooted plants showed more variation, although all plants settled relatively fast (Fig. 10). The red algae, Chondrus sp. and Gracilaria sp. had the highest settling rates (80-85 m h-1). The red algae Polysiphonia sp. settled much slower (30 m h-1), and the rooted macrophyte Cymodocea sp. was intermediate. The experiment also showed that there was no difference in settling rates among Ulva pieces in different conditions (light and dark thallus); the major difference was in the chaotic settling pattern. We attempted to apply our experimental findings to observations of current velocities in Venice Lagoon to get an idea about the erosion and deposition areas (Bergamasco et al. 2003). The results were that almost all areas in the Lagoon showed current velocities higher than the measured traction thresholds. Thus, 90% of Venice Lagoon was subject to macrophyte mobilization. In some areas this resulted in biomass losses through erosion which were higher than the plant growth rate. The plants therefore disappeared from these areas.

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Figure 10. Average settling rates for all studied plants. The Ulva class is subdivided into light and dark plants which are settling either with the thallus horizontally (h) or vertically (v) during the settling.

Despite some variation in the measured settling rates of macrophytes, the experiments showed that resuspended plant material settle fast. This mechanism is important in explaining the general picture of the export/import balance at the open boundary of Venice Lagoon (Fig. 3). It appears that about 90% of the exported plant biomass did not return with the following flooding tide. The reason is that, that once in the Adriatic Sea, the plant material within 15-40 minutes will settle to 20 m depth, long before the tide turns and the strong in-going current eventually could carry the plant material back into the Lagoon.

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It is complicated to evaluate the relative importance of growth, grazing, sporulation and transport processes for the standing stock of non-attached or loosely attached macroalgae in a shallow system like the Lagoon of Venice because the dynamics of the physical, chemical and biological parameters are acting on these processes with both positive and negative feed-back mechanisms. The plant material caught at the outer boundary may have different origins which may be difficult to identify. Most often the caught material is a mixture of plants transported from distant inner areas and more locally grown. How can the local contribution to the transport loss be estimated and what is the contribution of this process to the dynamic biomass budget? Complex and time consuming campaigns are necessary. We tried to make these estimation based on direct field measurements in Roskilde Fjord and Venice Lagoon. The growth and grazing rates of Ulva sp. were measured by an enclosure/exclosure technique described by Geertz-Hansen et al. (1993). Cages were built of 20 cm long transparent Perspex tubes (Ө = 20 cm) cut longitudinally in halves. The resulting Perspex arches constituted the roof of the cages and the sides and bottom were covered with nylon net. The cages were fixed to the sediment surface by drilling coiled steel pegs into the sediment. In the growth measurements, a mesh size of 0.5 mm was used, preventing grazers from entering the cages. In the grazing measurement a mesh size of 4 mm was used, allowing the invertebrate grazers to enter the cage. In each chamber the specific growth rate µ was calculated (eq. 1) from the increase in surface area of 5 Ulva discs, cut from free-floating thalli with a sharpened Perspex tube (Ө = 2.5 cm). µ = (ln (biomasst) – ln (biomasst0)) d-1

(1)

where µ is specific growth rate (d-1), biomasst0 and biomasst are the tissue areas at the start and at the end of the incubation, respectively, and t is the number of days. The specific grazing rate g (d-1) was calculated as g = (ln (biomasst,0.5) – ln (biomasst,4.0)) t-1

(2)

where biomasst,0.5 and biomasst,4.0 are the final surface areas of Ulva discs in 0.5 mm and 4.0 mm cages, respectively. The specific sporulating rate s (d-1) was calculated as s = (ln (ghost_tissuet) – ln (ghost_tisuet0)) t-1

(3)

where ghost_tissuet and ghost_tisuet0 are the surface area of transparent ghost tissue of Ulva discs area at time t and t0. At the same time plant loss due to transportation was measured as net catches with the procedure described above.

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Figure 11. Growth, grazing and erosion/sedimentation for three transects in Møllekrogen, southern part of Roskilde Fjord, Denmark, 1996. Biomass losses are negative while gains are positive. From Salomonsen et al. (2000). The transects East and West represent stations with an average depth of 1m, while the Deep transect has an average depth of 3m.

The first example (Fig. 11) is from the shallow protected Møllekrogen in the inner part of Roskilde Fjord, where Ulva lactuca was the dominating macroalgae with a biomass of 2-35 g dwt m-2. During this study, 10 representative stations were placed inside an area of 0.42 km2. The stations were placed on three transects, an eastern and a western with average depths of 1 meter, and a middle deep transect at 3 meters depth. The tidal amplitude in this area is only 20 cm, resulting in the lowest hydrodynamic forces among the studied sites (Mondego Estuary, Portugal, Venice Lagoon, Italy and Roskilde Fjord, Denmark). Measurements of area specific biomass, growth and loss rates were based on a sampling frequency of twice a week. It is evident that the changes in rates were more dynamic at the shallow stations where gain and loss processes constituted a higher fraction of the biomass than at the deeper stations. This was also reflected by the fact that the deeper area had a higher biomass than the shallow area.

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Although the growth rates were much lower (Salomonsen et al. 1999). In some periods, the biomass was transported from the shallow area to the deeper areas. During the study there were no observations of sporulation, while the loss rate by isopod grazing at the western transect varied between 0.0 and 0.4 d-1 of the biomass. The maximal loss rate by transportation was about 1.0 d-1, while the biomass import rate was about 2.0 d-1. However, the transportation loss rate cannot be higher than 1.0 d-1 plus the new biomass produced by growth, altogether about 1.2 d-1, while gains as import may be several times the existing biomass.

Figure 12. Growth, grazing and erosion/sedimentation for the whole study area in Møllekrogen, southern part of Roskilde Fjord, Denmark, 1996. Positive and negative bars indicate growth and loss of biomass, respectively. Rising hatching indicates growth, falling hatching indicate grazing, and cross-hatching indicate export. The solid line is biomass abundance. From Salomonsen et al. (1999).

The plant dynamics for the study area is shown in Fig. 12. It is evident that the major loss process is transportation of plant matter across the open boundary. This process accounted for 75% of the biomass loss, while grazing only accounted for about 25%, and sporulation was not observed. The current velocity in this study area was simulated by a well-calibrated 2D-hydrodynamic model. This model showed that the current velocity peaked at about 10 cm s-1, which happened between days 15-20, where the highest plant export was observed (Fig. 12). For the rest of the study period, the current velocity was below 5 cm s-1. The second example is from Venice Lagoon, where we measured plant losses and gains at stations representative for huge subtidal areas and in intertidal

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St.50

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flats. The field techniques and methods applied for measurements of plant loss and growth rates were the same as described for the previous study. The results from the summer 1998 are shown in Fig. 13, where the average rates for the whole area were: Growth 0.075 d-1, sporulation 0.014 d-1, grazing 0.027 d-1, transport 0.064 d-1. It was verified that transport was the dominating loss process resulting in a slightly negative biomass balance.

0.10 0.00 -0.10

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transport

Figure 13. Field measurements of growth, grazing, sporulation and transport of benthic macrophytes at different stations representing important habitats in Venice Lagoon, August, 1998. The dominating species at the stations is indicated. Positive values indicate gain, and negative values indicate loss of biomass for the stations.

Looking for patterns in the measured rates at the different locations, it seems that the grazing rates are independent of whether the locations were subtidal or intertidal. The grazing rates varied between 0.01 d-1 and 0.04 d-1. The range of growth rates varied from 0.02 d-1 to 0.16 d-1, with the lowest rates at the high elevated intertidal flats (Zostera noltii stations). However, these growth rates were not high compared to maximal rates of about 0.67 d-1 found in Roskilde Fjord (Geertz-Hansen et al. 1993). Sporulation was only present at the high elevated stations of the intertidal zone, where it varied from 0.03 d-1 to 0.06 d-1. The sporulation at these stations was explained by high water temperatures of about 30°C in the stagnant shallow intertidal pools, which together with a very high and long lasting insolation of about 1800-2000 µmol m-2 s-1, resulted in unfavourable growth conditions for Ulva sp. The balance of the measured rates was estimated according to the equation: Balance = growth – (sporulation + grazing + transport)

(4)

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The negative result was in agreement with the very low biomass of macroalgae at these stations. The net transport was negative at all stations in this campaign, with the highest exports in the subtidal areas, where the current velocities were highest and where the plants were exposed to the current for the longest periods. At the subtidal Station 10 close to the Lido entrance, where a mixed vegetation was found (Cymodocea sp. (the dominating species), Chaetomorpha sp., Ulva sp., and Zostera sp.) the plant population managed to sustain a loss by transport although the loss rate were 0.13 d-1. The biomass did not decrease significantly during the study. In May-June 1995 we made a similar study on loss and growth processes at the station Malamocco (subtidal, mixed vegetation). Here the growth rate was only 0.043 d-1, the grazing rate was 0.01 d-1, the current velocity dependent transport loss ranged from 0.03 and 0.17 d-1, and no sporulation was observed. Further, we mapped biomasses of the ephemeral macroalgae and did nutrient mass balances across the large subtidal flat. The initial biomass of Ulva sp. around the station decreased from 0.82 to 0.10 kg wwt m-2 during the one month study period, while the biomass of Chaetomorpha sp. decreased from 0.45 to 0.03 kg wwt m-2. Assuming that these biomasses, growth and loss rates are representative for the complete 4 km2 subtidal flat, the initial biomass of macroalgae was 600 ton dwt, but decreasing to 61 ton dwt during one month. By construction of a small dynamic model for the area (Fig. 14), we reach an understanding of how big a transport loss a community of ephemeral macroalgae is able to compensate. Furthermore, we have the possibility to study the importance of the variability of the initial biomass, growth and loss rates. The simulation shown in Fig 15 A+B was based on the measured initial biomass and rates from the Malamocco station in 1995. However, it appears that the simulated biomass did not decrease to the measured final biomass of 61 tons dwt; it only decreased to 170s ton dwt. Therefore, the loss rates have to be higher in the model, although the accumulated transported plant mass is 810 and the total loss about 910 tons dwt. This scenario is very conservative, because the area also receives a major import from the inner part of the lagoon. The simulation thus verifies that our measurements of plant bound nutrient transport were realistic. The plant bound nutrient export made up about 90-95% of the total nutrient export from Venice Lagoon in this part of the growth season (Flindt et al. 1997a). We also simulated the influence on plant transport of a macroalgal growth rate which was increased sufficiently to keep the biomass in steady state (Fig. 15 C+D). To obtain a steady state for the macroalgae biomass, the growth rate constant has to be increased to 0.1 d-1, which still is a very realistic rate and far from maximal growth. The simulation shows that the growth rate and the plant transport rate alternate. During neap tides the growth rate compensates for the losses during spring tide.

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biomass

growth_rate

transport

accum_transport

graz_rate

growt_rate_const graz_rate_const transport_rate_const

accum_graz Figure 14. Powersim conceptual diagram of a macroalgae model with three state variables (boxes): macroalgae biomass accumulated grazing and accumulated transport. Rates (circles): growth rate, grazing rate and transport rate. The rate constants were the measured rates, while the transport_rate_constant was made dependent on current speed using the tidal cycle and introducing neap and spring tide cycles.

The major consequence is that the plant transport increases from 810 ton dwt to 1857 ton dwt, 2.3 times the measured transport. This shows that high growth rates support high losses, which is the case in Venice Lagoon. The nutrient loading is very high in Venice Lagoon, and not limiting the primary producers. Because the average depth is only 1 meter, light never becomes a limiting factor, apart from when self shading occur. Deeper channels exist but they only cover a few percentage of the total lagoon area. A simulation with a higher growth rate of 0.16 d-1 combined with a grazing rate of 0.03 d-1 is shown in Figure 15 (E+F). This simulation results in a large increase of macroalgal biomass, increasing from 600 tons dwt to a final amount of about 1800 tons dwt. A tripling within a month, which is not unrealistic. The accumulated simulated grazing becomes almost twice the initial macroalgae biomass, while the accumulated plant transport becomes 5 times that measured in 1995.

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Figure 15. The left panels (A+C+E) show the simulated development in the standing biomass of macroalgae, accumulated plant transport and accumulated grazing. The right panels (B+D+F) show the realized rates per day for growth, transport and grazing. Scenario (A+B): Rate constants as measured in the summer 1995: grazing 0.01 d-1, growth 0.043 d-1, and average plant transport 0.08 d-1. Scenario (C+D): Rate constants as measured during the summer 1995: Grazing 0.01 d-1 and average plant transport 0.08 d-1, while the growth constant was increased to 0.1 d-1. Scenario (E+F): Rate constants were measured during the summer 1998: grazing 0.03 d-1, growth 0.16 d-1, and average plant transport 0.08 d-1.

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7. FIELD STUDIES ON PLANT BOUND NUTRIENT TRANSPORT. During several field campaigns in Venice Lagoon (May-June 1995, August 1998, March 1999, June 1999), Roskilde Fjord (May 1996, August 1996, May-July 1999, September 1999) and Mondego Estuary (monthly sampling in 2000-2001), growth, grazing, and transport of macroalgae and rooted macrophytes were measured during different tidal excursions, while water samples were analysed for dissolved inorganic nutrients and total suspended solids. Water samples were often taken from platforms by an ISCO Manning water sampler. Samples were taken with a high frequency ensuring a fine temporal resolution over the tidal cycles. The samples were filtrated and analysed for suspended solids, loss on ignition of the particulate organic matter, total phosphorus and total carbon and nitrogen. The filtrated water was analysed for dissolved reactive phosphate, ammonia, and nitrite + nitrate, and sometimes also for dissolved organic nitrogen. Nets were fixed to wooden poles at the sides and anchored to the bottom to ensure that plant material did not escape at the bottom. The measured horizontal transport was normalized to effective exposure of the net, by using the hydrodynamic simulation to calculate the areal means of the current direction from the six nearest grid points. This mean value was then used to calculate the effective width of the net. Other times the current velocity and direction was measured manually. The net was harvested every 0.5-1 hour, and after sorting, the wwt. of each species was recorded. Each macroalgal species were analysed for dry/wet weight ratio, total carbon, total nitrogen and total phosphorus. The current velocity and direction was either logged or it was calculated by a local well-calibrated and validated hydrodynamic model (Venice Lagoon: 2D models and measurements (Flindt et al. (1997)), Umgisser & Zampato (2001) and loggers, Mondego Estuary: 2D modelling (Duarte et al. (2001)) and measurements, Roskilde Fjord: 1D model (Flindt & Kamp-Nielsen (1997)), 2D Lagrangian model (Salomonsen et al. (1999)), 3D finite element model (Umgisser & Flindt (2001)) and loggers. For Venice Lagoon and Roskilde Fjord, the measurements were repeated so many times that it was possible to calculate yearly mass balances at the outer boundary. Such mass balances has never before included all nutrient fractions, so here we are able to verify to what extent plant bound nutrient export is essential at different tidal regimes, where Venice represents a meso-tidal system and Roskilde Fjord a micro-tidal system. Venice Lagoon Venice Lagoon is 50 km long and the average width is about 10 km covering an area of 540 km2. It communicates with the Adriatic Sea by three outer boundaries: Lido, Malamocco and Pellestrina. The width of these boundaries varies from 400 m to 900 m. The average depth is only 1 m, while the maximum tidal excursion is about 1.1 m. About 20% of the area is inter-tidal flats, which are only periodically submerged, while the major part of the lagoon is very shallow subtidal areas.

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The combination of the shallowness and the relatively large tidal range ensure a short residence time of 1-10 days depending on the location. Because of the bathymetry light never becomes growth limiting for benthic plants, which are responsible for 9599% of the primary production. The nutrient loading is about 6000 tons nitrogen and 1000 tons phosphorus, which together with the good light conditions makes the lagoon one of the most productive aquatic systems (Sfriso, 1992, 1995).

Fig. 16. Venice Lagoon with the active stations during the field campaigns in the summer 1998 and winter 1999.

Nutrient transport was measured at the three outer boundaries (station 10, 60 and 90) during the field campaigns in summer 1998 and winter 1999. The northern Station 10 covered the drainage from a mixture of Spartina sp. salt marsh areas and intertidal flats dominated by ephemeral macroalgae and Zostera noltii. The middle and southern part of the lagoon were covered by measurements at Station 60 and Station 90. These areas were dominated by large subtidal areas where seagrasses and ephemeral macroalgae dominated. The measured export balances during summer and winter are presented in Figure 17.

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Figure 17. Nitrogen and phosphorus mass balance for Venice Lagoon based on transport measurements during summer 1998 (A+B) and winter 1999 (C+D). DIN-N is dissolved inorganic nitrogen, Veg-X is the summed plant bound nutrient transport (nitrogen or phosphorus), SPM-X is the suspended particulate nutrient, Fyto-X is based on chlorophyll a.measurements converted by the Redfield-ratio to phytoplankton bound nutrient, and PO4-P is phosphate.

The mass balances at the three outer boundaries are based on transport measurements at Stations 10, 60 and 90 (Fig. 16). The results are extrapolated to cover the full hydraulic cross section area for each of the boundaries. This may introduces an underestimation of the mass balance because we were only able to carry out the campaigns at shallow stations. It was impossible to place the nets at the bottom in the deeper channels where most of the transport occurs. During the growth season plant bound nutrient transport dominated. About 75% of the nitrogen export appears plant bound, while about 55% of the phosphorus is exported through plant movements. Both concentrations of dissolved inorganic nitrogen and phosphate are low, although higher than during the measurements in 1995. Phytoplankton did not contribute to the mass balances. Its contribution was less than 0.5% of the mass transport for both phosphorus and nitrogen. Due to the shallowness and extended boat activities in the area resuspension occurs frequently. This results in relatively high concentrations of suspended particulate matter, which for nitrogen accounted for about 20% of the nutrient export during the summer and

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about 25% in the winter. The suspended particulate phosphorus transport was about 40% of the mass balance in the summer and 33% in the winter. During the field campaigns in 1995, 1998 and 1999 the N/P ratio of the suspended particulate matter was relatively low (1.5-5), indicating that suspended particles act as a phosphate sink for the freshwater discharge in the winter and spring, but also as a sink for the internal phosphate loading from the sediment. During the growth season the N/P-ratio of the transported plants was high (about 15) which may indicate phosphorus limited system. During winter export was dominated by dissolved nutrients, where DIN accounted for about 60%, while dissolved phosphate contributed with 40% of the export. Plant bound nutrient transport accounted for 23% of the nitrogen export and about 17% of the phosphorus export.

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Figure 18. Yearly phosphorus and nitrogen mass balance for Venice Lagoon.

The export measurements were integrated for each season to get the annual mass balance (Fig. 18). Here it is assumed that the growth season ranges from March until November (270 days) and the winter covers December-February (90 days). For the whole year, plant bound nitrogen export accounted for about 63% of the total calculated export, while plant bound phosphorus export accounted for 42%. The official nutrient loading is 6000 tons N and 1000 tons P (Flindt et al. 1997b). Our calculated export was 5100 tons nitrogen and 500 tons phosphorus. This means that about 50% of the phosphorus is either missing or accumulated inside the lagoon. For nitrogen, the calculated export accounts for 85% of the loading. Venice Lagoon is net exporting suspended particulate matter. This is due to erosion of tidal flats. The reason for this is that the major tributaries are cut off from the lagoon, so they discharge directly to the Adriatic Sea. The missing sand and silt contribution has

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resulted in degradation of salt marsh habitats in extended areas of the northern part of the lagoon. For this reason we assume that a major part of the missing 50% phosphorus was transported undetected as bedload transport. A partly reintroduction of major tributaries will be necessary to keep the inorganic particle budget balanced in the lagoon; otherwise the salt marshes will degrade. A reintroduction of the major tributaries will also include iron import that will act as a sink so the internal phosphate loading from the sediment not becomes realized as new macroalgae biomass. This will be even more effective if the tributary particle mass contribution happens in spring where the melting water from the mountains carry huge amounts of inorganic particles. Here the import of iron and particle mass will be relatively higher, while the nutrient contribution will be exported before the vegetation starts growing. Roskilde Fjord Roskilde Fjord is a shallow estuary connected to Kattegat on the north coast of Zealand, Denmark (Fig. 19). The length from the inner southern part to the outer northern part is about 40 km and the maximum width is about 10 km. The surface area is 122 km2. The average depth is 3 m and the drainage area is 1127 km2. The tidal excursion is about 0.2 m, so Roskilde Fjord represents a micro-tidal system. The hydraulics are more dynamic than indicated by the tidal range, because wind forcing is able to cause changes in water levels of up to 1.6 m above mean sea water level. Together with the freshwater discharge of about 350 x 106 m3 y-1 from tributaries and diffuse runoff, these mechanisms create gradients of salinity from about 10 ‰ in the southern part to about 20 ‰ at the outer boundary. The residence time in the southern part of Roskilde Fjord is above 1 year and about 2-4 weeks in the northern part close to the open boundary. This long residence time has huge ecological impacts on light climate and on the benthic pelagic nutrient coupling. About half of the primary production is created by benthic primary producers, while phytoplankton is responsible for the other fraction. The nutrient loading is about 1500 tons nitrogen and about 60 tons phosphorus (Hedal & Hansen 2001). For Roskilde Fjord we have grouped the measured integrated nutrient transport across the outer boundary into three seasons. Most of the nutrient transport took place during the winter to early spring (Jan-Apr). This period was dominated by high concentration of nitrate due to freshwater runoff. About 50% of the export was dissolved inorganic nutrient, while suspended particulate nitrogen contributed with about 40%. The phytoplankton nitrogen was about 5%, primarily created during the spring bloom. The nutrient transport was lowest during the growth season from May to September. Here the export was dominated by plant material transport that accounted for about 70% of the nitrogen export. During the autumn and early winter (Oct-Dec), the suspended matter was dominating the export, but dissolved nutrient and plant transport accounted for about 25% and 20%, respectively. Dissolved organic nitrogen was included in these studies, but this fraction accounted at maximum during the growth season for 0.5% of the nitrogen transport.

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Figure 19. Roskilde Fjord. Left: Bathymetric map showing the depth conditions. Right: Current velocity map presenting the range of current speeds in the fjord during a drop in water level of 30 cm.

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Figure 20. Nitrogen and phosphorus mass balance for Roskilde Fjord 2000. The upper panel shows the export of the different N-fractions: NHx-N is ammonia, NOx-N is nitrate + nitrite, DON-N is dissolved organic nitrogen, SPM-N is suspended particulate nitrogen, Veg-N is nitrogen bound in vegetation and Fyto-N is phytoplankton nitrogen. The lower panel shows the export of the phosphorus fractions: POP is phosphate, SPM-P is suspended particulate phosphorus, Veg-P phosphorus bound in vegetation and Fyto-P is phytoplankton phosphorus.

For the phosphorus transport the tendencies were the same with high dissolved inorganic phosphate export in January to April, but also with a huge contribution from suspended particulate phosphorus of about 40%. Phytoplankton transport contributed about 8% of the phosphorus transport in this period increasing to about 11% in the growth season. In the growth season, the phosphorus export was dominated by plant

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material (55%) and suspended particulate phosphorus. In this period there was a small net import of phosphate. In October to December, phosphate was regenerated, which together with an increasing freshwater discharge influence the transport pattern.

Mass transport (%)

100 Phosphorus

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Nitrogen

60 40 20 0 Dis. Inorg.

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Figure 21. Annual contribution of nitrogen and phosphorus fractions to the mass transport in Roskilde Fjord for 2000. Dis.Inorg is dissolved inorganic nutrients (phosphorus or nitrogen). Dis. Org. is dissolved organic nutrients. SPM is suspended particulate nutrients. Veg is plant bound nutrients, and Fyto is phytoplankton nutrients.

An annual nitrogen and phosphorus balance for the different nutrient fractions contributing to the mass balance at the outer boundary of Roskilde Fjord is shown in Figure 21. Here it becomes evident that nitrogen and phosphorus mainly are exported from Roskilde Fjord as dissolved inorganic nutrients and suspended particles (about 70-80% of the total export). But nutrients bound in plant material was also important, the macrophytes accounted for 24% of the phosphorus export and 18% of the nitrogen export, while phytoplankton bound transport accounted for about 8% and 3% of the phosphorus and the nitrogen export, respectively. Mondego Estuary, Portugal. All study sites were within the South arm of the Mondego Estuary. This location is 10 km long and the average width is about 0.3 km covering an area of 3.4 km2. It communicates with the North Atlantic Ocean. The average depth is about 2 m, while the maximal tidal excursion is about 3.3 m. Huge areas are intertidal salt marsh covered downstream by Spartina maritima and upstream by Scirpus sp. The Mondego Estuary represents a macro-tidal salt marsh system. The residence time is about 1-4 days depending on the location. The system has a high turbidity, causing light to be growth limiting for the submersed and lower intertidal plant communities.

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The nutrient loading is about 126 tons nitrogen and about 10 tons phosphorus (Lillebø et al. 2002). Figueira da Foz Harbour

North arm

Atlantic Ocean

Gala bridge

Murraceira Island

PORTUGAL

South arm 1 Km

Armazéns channel

Intertidal areas

Pranto river

Figure 22. Map showing the Mondego Estuary with the South arm where the transport measurements were made.

In Mondego, the nutrient loading and transport is complicated by several processes. First of all, during rainy periods water drains from agricultural areas in the lower part of the Mondego Valley through a sluice at the end of the Prante River. These events are not electronically registered, so the exact discharge of water and nutrient is not known. But when the freshwater discharge is high, the salinity drops to just above zero in the inner part of the south arm. This results in collapse of Enteromorpha sp., because the macroalgae can not achieve a positive net growth rate at low salinities (Martins et al. 1999, 2001). If the freshwater discharge is lower, causing salinities about 5 ‰ in the inner part, huge Enteromorpha blooms are realized. These blooms happens only in moderately wet years, while very rainy or dry years are free of blooms, because the losses through sloughing and transportation and/or sporulation is higher than the net growth rate. During the studied year 2000, there was no spring bloom of Enteromorpha in the estuary. In Figure 23 representative measurements of nitrogen and phosphorus transport at the outer boundary are shown for a summer (Fig.23 A+B) and a winter campaign (Fig.23 C+D). In the summer period dissolved inorganic nitrogen (DIN) dominated the transport in the system with 60% of the net export. About 50% of the exported DIN returned during the following import situation, so about 50% was permanently lost to the coast. With respect to suspended

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particulate nitrogen (SPM-N) the net nitrogen export was about 35%, while the macrophyte bound nitrogen only was about 5% of the total net N-export.

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Figure 23. Nitrogen and phosphorus import, export and mass balance for the Mondego South arm based on transport measurements during summer 2000 (A+B) and winter 2000 (C+D). DIN-N is dissolved inorganic nitrogen, Veg-X is the summed plant bound nutrient transport (nitrogen or phosphorus), SPM-X is the suspended particulate nutrient, Fyto-X is based on chlorophyll a. measurements converted by the Redfield-ratio to phytoplankton nutrient and PO4-P is phosphate.

A higher proportion of SPM-N was permanently lost by the transport to the sea due to a low re-import. The reason for this is that the particles were transported as bedload so the heavy fraction sedimentated before the water returned to the estuary again. The phytoplankton bound nutrients only accounted for less than 1% of the mass transport. The residence time was very short in the estuary, and therefore, if the phytoplankton is not able to obtain growth rates close to the maximal they are washed out of the system. The transport pattern for phosphorus during the summer was similar to that for nitrogen. Here the net phosphorus export was 52%, 40% and 8%, respectively, for phosphate, SMP-P and macrophyte bound phosphorus. During the early winter DIN transport dominated the nitrogen export. But the net loss of DIN measured as the difference between import and export was less, because about 2/3 of the exported DIN was re-imported. During this winter there was a net import of both nitrogen and phosphorus in the form of suspended particulate matter. The reason for this net particle mass import was most probably that some of the huge discharge from the

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north arms was forced into the south arm during rising water levels. Proportionally, a higher amount of the phosphorus transport was particle bound as compared to the nitrogen transport. Macrophyte and phytoplankton bound nutrient transport was negligible. As earlier stated, unbalanced energy budgets for salt marsh ecosystems led Odum (1968) to propose the “Outwelling Hypothesis”. This hypothesis has been inspiration for many investigations and reviews (e.g. Nixon 1980, Dame 1989, 1994, Dame & Allen 1996) of plant transportation from salt marshes to coastal waters. It seems evident that North American and European salt marshes respond different as exporters of the primary production. American salt marshes export a larger fraction of the production than the European ones. A good explanation was given by Beeftinf (1977) and Lefeuvre (1994), who stated that the American marshes topographically is located lower and hereby more exposed to hydrodynamic forcing. Thus, the potential period for hydrodynamic forced sloughing of plant leafs is much longer. Another essential difference is that the vegetation coverage is much less in the lower part of European salt marshes compared with American marshes at the same elevation (Adam 1990, Packham & Willis 1997), and there is therefore less plant material to export. Our study from Mondego is in agreement with other investigations of European salt marshes (Hemminga et al. 1993, 1996, Lefeuvre et al 1994,), where the general tendency is that most of the production is kept inside the systems (Bouchard & Lefeuvre 2000).

8. CONCLUSION. Through laboratory and field measurements it has been verified that unattached living macroalgae start to move as bedload transport at very low current velocities (2-3 cm s-1), and at about 6-9 cm s-1 most of these are lifted from the bed and come into suspension where they move with the same speed as the water. Even attached macroalgae are often transported, and at a current velocity threshold of 12 cm s-1 some ephemeral species start to loose mass due to sloughing. This process increases with current speed and at about 40 cm s-1, they loose a considerable amount of mass ranging between 25% and 50% depending on the species. Other macroalgae like Fucus sp. are relatively insensitive to sloughing, the tissue is simple too strong to be torn apart. The effect of these low loss thresholds is a potentially high export of macroalgae from estuaries to coastal waters. Measurements verify that most submersed plants settle within a range of 35 m h-1 to 85 m h-1. Compared to phytoplankton, macroalgae settling rates are about 1000-5000 times faster. This mechanism is important in explaining the general picture of the export/import balance at the open boundaries between estuaries and coastal waters, where only a small fraction of the exported plant mass returns with the following flooding tide. The plant matter simply settles before the tide changes direction. The only complication is the newly sloughed material from seagrasses, which do not settle due to the aerenchymatic tissue. Seagrass movements are therefore both affected by current speed, wind speed and wind direction. Hereby their transport pattern becomes more complex to predict.

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The field measurements presented from Mondego Estuary, Venice Lagoon and Roskilde Fjord, cover large variations in tidal amplitudes. Mondego Estuary represents a macro-tidal salt marsh system, Venice Lagoon represents a meso-tidal estuary, and Roskilde Fjord represents a micro-tidal system. Table 4.Relative contribution of plant bound nutrient transport to the total mass transport

System Mondego Estuary Venice Lagoon Roskilde Fjord

. Contribution of plant transport to mass balances Nitrogen (%) Phosphorus (%) Summer winter Summer winter 5 1 8 0 72 21 51 18 71 12 55 16

There are large differences in the relative importance of plant bound nutrient transport from turbid salt marshes like Mondego (or other European salt marsh systems) and from estuaries with extensive shallow subtidal areas such as Venice Lagoon or Roskilde Fjord (table 4). This difference is first of all linked to differences in hydrodynamic forcing as well as differences in plant society. In European salt marshes hydrodynamic forcing is so high that turbidity restricts plant growth in the lower elevated salt marsh areas and in subtidal areas. Logically, a low production will lead to lower losses. The production is higher at the high elevated marsh areas but the hydrodynamic forcing is lower and less frequent because the water only reaches the Spartina beds during the days around spring tide. There is no literature about sloughing threshold for Spartina, but the threshold is most probably high. Plant communities in macro-tidal salt marshes and meso- to micro-tidal estuaries differ. Zostera marina is the dominant subtidal rooted plant in meso- and micro-tidal estuaries. This species slough by nature because the oldest leaf is rejected when a new leaf is formed,. This sloughing is non-dependent on current dynamics. Furthermore, the leaves float in the first period after rejection, so no current energy is needed to lift it from the bottom before transportation. Free living or attached ephemeral macroalgae with high transportation potential are also much more abundant in mesoand micro-tidal estuaries, where the residence time is longer and the hydrodynamic forcing is less. These mechanisms explain why estuaries like Venice Lagoon and Roskilde Fjord export large quantities of plant matter, and thus also a large nutrient mass associated with this export. Monitoring programs around the world are set up to understand and evaluate the effects of agricultural dominated nutrient runoff. Managers and scientists working with applied problems like the reasons for oxygen depletion in coastal waters, need to have reliable nutrient mass fluxes at the outer boundary of estuaries. Today, plant bound nutrient transport is not included in these mass fluxes, which leads to nutrient balances for the estuaries where the estimated nutrient retention is too high. The consequence is underestimation of the nutrient loading to the open coastal areas. For Venice Lagoon and Roskilde Fjord the yearly underestimation is 60% and 18% on the nitrogen budget and 44% and 23% on the phosphorus budget, respectively.

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9. ACKNOWLEDGEMENTS

The studies were financed by the European MAST III Program through activities linked to the F-ECTS project (MAS3-CT97-CT97-0145), the European TMR Program through activities linked to the WET project (FMRX-CT96-0051), and by the Portuguese Praxis program through a PhD grant. 10. REFERENCES

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Flindt, M.R., Marques, J.C., Partial, M.R., Bocci, M., Bendoriccho, G., Kamp-Nielsen, L., Nielsen, S.N., & Jørgensen, S.E. (1997b). Description of the three shallow estuaries: Mondego River (Portugal), Roskilde Fjord (Denmark) and the Lagoon of Venice (Italy). Ecological Modelling, 102, 17-31. Flindt, M.R. & Kamp-Nielsen, L. (1997). Modelling an estuarine eutrophication gradient. Ecological Modelling. 102: 143-154 Flindt, M.R., & Kamp-Nielsen, L. (1998). The influence of sediment resuspension on nutrient metabolism in the eutrophic Roskilde Fjord, Denmark. Verhandlungen Internationale Vereinigung Limnologie, 26, 1457-1461. Flindt, M.R., Pardal, M.A., Lillebø, A.I., Martins, I., & Marques, J.C. (1999). Nutrient cycling and plant dynamics in estuaries: A brief review. Acta Oecologica, 20, 237-248. Geertz-Hansen, O. Sand-Jensen, K., Hansen, D.F., & Christiansen, A. (1993). Growth and grazing control of abundance of the marine macroalga, Ulva lactuca L. in a eutrophic Danish estuary. Aquatic Botany 46, 101-109. Greenway, M. (1976). The grazing of Thalassia testudinum in Kingston Harbour, Jamaica. Aquatic Botany 2, 117-126. Hedal, S. & Hansen, L.R. (2001). Roskilde Fjord 2000. Frederiksborg Amt. Hemminga, M.A., Cattrijsse, A. & Wielemaker, A. (1996). Bedload and Nearbed Detritus Transport in a Tidal Saltmarsh Creek. Estuarine, Coastal and Shelf Science 42, 55-62. Hernández, I., Peralta, G., Pérez-Llorens, J. L., Vergara, J., & Niell, F. X. (1997). Biomass and dynamics of growth of Ulva species in Palmones river estuary. Journal of Phycology 33, 764-772. Kamp-Nielsen, L. (1992). Benthic pelagic coupling of nutrient metabolism along an estuarine eutrophication gradient. Hydrobiologia, 235/236, 457-470. Lefeuvre, J.C. (1994). Comparative studies on salt marsh processes in the baie du Mont Saint-Michel: a multi-disciplinary study. In: Mitsch, W.J. (Ed.). Global Wetlands: Old World and New. (pp. 215-234). Amsterdam. Elsevier. Lillebø, A.I, Pardal, M.A., Flindt, M.R, Neto, J.M., Macedo, F., Martins, I. & Marques, J.C. (2002). Nutrient dynamics in the intertidal pools of the Mondego Estuary. IV. Possible contribution to dissolved inorganic phosphorus loading. In Aquatic Ecology of the Mondego River Basin. Global importance of local experience. Ed. Pardal, Marques & Graca. (pp. 287-300). Coimbra University Press. Martins, I., Oliveira, J.M., Flindt, M.R. & Marques, J.C. (1999). The effect of salinity on the growth of Enteromorpha intestinalis (Chlorophyta) in the Mondego estuary (west Portugal). Acta Oecologica. 20 (4) 259-265. Martins, I, Pardal, M.A., Lillebø, A.I., Flindt, M.R. and Marques, J.C. (2001). Hydrodynamics as a major factor controlling the occurrence of green macroalgal blooms on the influence of precipitation and river management. Estuarine, Coastal and Shelf Science. 52: 165-177. Menzies, R.J., Zaneveld, J.S., & Pratt, R.M. (1967). Transported turtle grass as a source of organic enrichment of abyssal sediments off North Carolina. Deep-Sea Res. 14, 111-112. Menzies, R.J., & Rowe, G.T. (1969). The distribution and significance of detrital turtle grass Thalassia testudinum on the deep sea floor off North Carolina. Internationale Revue der Gesamten Hydrobiologie 54, 217-222. Nixon, S.W. (1980). Between coastal marshes and coastal waters – A review of twenty years of speculation and research on the role of salt marshes in estuarine productivity. In P. Hamilton & K.B. MacDonald (Eds.), Estuarine and wetlands processes (pp. 437-525). New York: Plenum Press.

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Odum, E.P. (1968). A research challenge: evaluating the productivity of coastal and estuarine water, Proceedings of the 2nd Sea Grant conference (pp. 63-64). Kingston: University of Rhode Island. Odum, W.E., Fischer, J.S., & Pickral, J.C. (1979). Factors controlling the flux of particulate organic carbon from estuarine wetlands. In R.J. Livingston (Ed.), Ecological processes in coastal and marine systems (pp. 69-80). New York: Plenum Press. Packham, J.R., & Willis, A.J. (1997). Ecology of Dunes, Salt marsh and Shingle. Chapman & Hall, London. 335 pp. Pedersen, M. F. (1995). Nitrogen limitation of photosynthesis and growth: comparison across aquatic plant communities a Danish estuary (Roskilde Fjord). Ophelia, 41, 261-272. Salomonsen, J., Flindt, M.R., & Geertz-Hansen, O. (1997). Significance of advective transport of Ulva lactuca for a biomass budget on a shallow water location. Ecological Modelling, 102, 129-132. Salomonsen, J., Flindt, M.R., & Geertz-Hansen, O. (2000). Modelling advective transport of Ulva lactuca (L) in the sheltered bay, Møllekrogen, Roskilde Fjord, Denmark. Hydrobiologia, 397, 241-252. Sfriso, A., Pavoni, B., Marcomini, A., & Orio, A. A. (1992). Macroalgae, nutrient cycles, and pollutants in the Lagoon of Venice. Estuaries,15, 517-528. Sfriso, A., (1995). Temporal and spatial responses of growth of Ulva rigida C. Ag. To environmental and tissue concentrations of nutrients in the Lagoon of Venice. Botanica Marina, 38, 557-573. Sfriso, A. & Marcomini, A. (1997). Macrophyte Production in a Shallow Coastal Lagoon. Part 1: Coupling with Chemical-Physical Parameters and Nutrient Cencentration in Waters. Marine Environment Research 44(4), 351-375. Thomson, C.E.L. & Amos, C.L. (2002). The impact of mobile Disarticulated Shells of Cerastoderma edulis on the Abrasion of a Cohesive Substrate. Estuaries 25, 204-214. Umgiesser, G. & Zampato, L. (1997). Hydrodynamic and salinity modelling of the Venice channel network with coupled 1-D-2-D mathematical models. Ecological modelling 138, 75-85. Umgiesser, G. & Flindt, M.R., Amos, C.L. & Bergamasco, A. (2001). Exporting the phytobenthos-reaction modeling approach: Roskilde Fjord case study. Proceedings from the ELOISE-congress, Bavena, Italy. Vermat, J.E. & Sand-Jensen, K. (1987). Survival, Metabolism and growth of Ulva lactuca under winter conditions: a laboratory study of bottlenecks in the life cycle. Marine Biology 95, 55-61. Viaroli, P., Pugnetti, A., & Ferrari, I. (1991). Ulva rigida growth and decomposition processes and related effects on nitrogen and phosphorus cycles in a coastal lagoon (Sacca di Goro, Po River Delta). 25th Proc European Marine Biology, 77-83. Viaroli, P., Bartoli, M., Bondavalli, C., Christian, R.R., Giordani,G., & Naldi, M. (1996). Macrophyte communities and their impact on benthic fluxes of oxygen, sulphide and nutrients in shallow eutrophic environments. Hydrobiologia, 329, 105-119. Viaroli, P. Bartoli, M., Fumagalli. I. & Giordani G (1997). Relationship between benthic fluxes and macrophyte cover in a shallow brackish lagoon. Water Air and Soil Pollution, 99 (1-4), 533-540. Zieman, J.C., Thayer, G.W., Robblee, M.B., & Zieman, R.T. ( 1979). Production and export of sea grasses from a tropical bay. In R.J. Livingston (Ed.), Ecological processes in coastal and marine systems (pp. 21-33). New York: Plenum Press.

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M.R. Flindt & F.Ø. Andersen: Institute of Biology, University of Southern Denmark, Campusvej 55, DK-5230 Odense M, Denmark. J. Neto & M.A. Pardal: Institute of Marine Research, Department of Zoology, University of Coimbra, Portugal. C.L. Amos: Southampton Oceanographic centre, University of Southampton, England. A. Bergamasco: CNR, Messina, Sicily, Italy. C.B. Pedersen: Freshwater Biological Laboratory, University of Copenhagen, Helsingørsgade 51, DK-3400 Hillerød, Denmark.

JENS KJERULF PETERSEN

GRAZING ON PELAGIC PRIMARY PRODUCERS - THE ROLE OF BENTHIC SUSPENSION FEEDERS IN ESTUARIES

1. INTRODUCTION For more than two decades it has been recognized that benthic suspension feeders, by filtering particles out of the water, may have a major impact on the overlying water column. Benthic suspension feeders comprise a large group of both passive and active filter feeders, but only the impact of actively pumping macro-zoobenthic organisms will be dealt with. Of these are bivalves like mussels, oysters and clams the most investigated benthic suspension feeders. Other important groups are ascidians, sponges, polychaetes and bryozoans. By actively pumping water past a filtering apparatus, benthic suspension feeders are able to capture particles from the size of big bacteria of 1-2 µ (sponges and ascidians) or small phytoplankters or other particles of about 4 µ (bivalves) 100% efficiently. The activity of benthic suspension feeders is influenced by a number of environmental characters like temperature and particle concentration. Within e.g. the temperature range of a given suspension feeder, increase in temperature will result in increased clearance rate (e.g. Petersen & Riisgård 1992). With increased particle load, clearance will on the other hand be reduced above a certain threshold level. The different species of benthic suspension feeders differ to some extent with regard to how the respond in relation to environmental parameters. Some species, like most bivalves, have the ability to sort particles and are thus well adapted to the turbid environment of many estuaries. These eco-physiological aspects of benthic suspension feeding are, however, not the topic of my presentation. It is rather the ecological impact of the grazing by benthic suspension feeders on the biological structure of the water column. Two different but related approaches have been used to document the importance of benthic suspension feeders in shallow coastal waters: “Ask the animals” is an ecophysiological view of the potential impact of suspension feeding populations of macro zoobenthos, whereas “ask the water” is an ecological approach focused on studying the water column.

1.1

Ask the animals

By “asking the animals”, physiological parameters like filtration rate, growth or energy budgets of individual benthic suspension feeders are extrapolated to population level. By determining filtration rate of a dominant benthic suspension feeder as a function of size, temperature and concentration of particulate organic 129 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 129-152. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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material (POM), a model of filtration in relation to environmental conditions can be established. Combined with estimates of population size, geographical extension and size structure of the population in consideration, an estimate of benthic grazing potential in a given area can be calculated. Reported grazing potentials may be high, often more than one time the volume of the investigated ecosystem per day (Jordan & Valiela 1982, Loo & Rosenberg 1989, Petersen & Riisgård 1992, Petersen et al. 2002). An example of this kind of approach is the study (Petersen & Riisgård 1992) of the ascidian Ciona intestinalis in Kertinge Nor, a shallow Danish cove. The area of extension and the density of the ascidian population were determined several times by scuba diving and photogrammetrical analysis of stereophotographs. Clearance rates of the ascidians were determined in the laboratory as a function of size and temperature. From population density, size structure and water temperature estimates of grazing capacity could be estimated. As seen in table 1, the potential grazing impact of the ascidians in the cove was always high but variable due to both individual capacities of the ascidians as well as their population density. In other systems dominated by organisms with a more stable population structure, like perennial mussel, the variation in benthic grazing potential is likely less than in this highly dynamic annual species. A more integrative approach is to extend the population description to include secondary production and compare this with either grazing potential or primary production. It has thus been shown that secondary production by infaunal suspension feeding bivalves could be of the same magnitude as the primary production in a shallow semi-exposed habitat (Möller et al. 1985, Loo & Rosenberg 1996). However, in the more exposed Laholm Bay with deeper water, the same infaunal species did not fully exploit their grazing potential and secondary production was about 50% of primary production (Loo & Rosenberg 1989). Using estimates of secondary production, as an indication of benthic control is in fact one step beyond the classical “ask the animals” approach. The primary question for benthic ecologists has often been to determine limiting factors for zoobenthos populations (see e.g. Fréchette & Lefaivre 1990, Fréchette et al. 1992, Wildish & Kristmanson 1997). From this point of view, food can be a limiting factor whether seen on the small scale of a mussel bed (Frechette & Bourget 1985a, Fréchette & Bourget 1985b) or on a larger ecosystem level (Beukema & Cadée 1986, Smaal & van Stralen 1990, Smaal et al. 2001). For both levels the concept is to understand how food supply has the potential to structure populations. The next natural step is to apply the concept of food limitation to a manageable tool and estimate or model carrying capacity of those coastal areas used for cultivation or fishery of edible benthic suspension feeders like mussels, clams and oysters. It is, however, important that the carrying capacity characteristic of a system involves both physical (mixing) and biological (primary production) properties. 1.2

Ask the water

By asking the water main focus is put on water column parameters and it is the study of these that lead to the conclusions regarding benthic control.

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Table 1. Population clearance of the ascidian Ciona intestinalis in Kertinge Nor based on individual clearance rate and population abundance. T1/2 is the mean residence time for an algal cell over the ascidian population and Q is the quotient of the ascidian filtration capacity to the water volume of the cove.

Time

Nov 90 Mar 91 May 91 Sep 91 Dec 91 Apr 92 Jun 92 May 94 Jun 94 Aug 94 Oct 94 May 95 Jun 95 Sep 95 Apr 96 May 96 Jul 96 Sep 96

Temperature

Weight

Clearance rate

Abundance

Population clearance

T1/2

Q

°C

mg ind-1

ml min-1

ind m-2

l h-1 m-2

h

d-1

5

86

6.5

450

176

8

1.1

6

57

6.6

76

30

46

0.2

13.5

169

31.4

20

38

37

0.2

12.5

56

13.5

237

192

7

1.2

3.5

49

2.8

277

47

29

0.3

8.5

90

12.5

139

105

13

0.6

20

295

70.1

33

139

10

0.9

12

55

12.9

169

131

11

0.8

14.5

381

59.8

75

268

5

1.6

22.5

23

14.0

448

375

4

2.3

10

59

11.1

168

111

12

0.7

12

69

15.0

118

107

13

0.7

15.5

112

46.6

152

425

3

2.6

15

24

16.2

36

35

40

0.2

10

30

12.4

69

51

27

0.3

13

254

39.8

27

65

21

0.4

18

15

14.2

448

382

4

2.4

16

47

27.3

579

949

1

5.8

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In this approach, phytoplankton biomass and productivity are thus compared with available nutrients and various loss factors like zooplankton grazing or advection and other limiting factors like light extinction. The result is then that the loss and limiting factors cannot account for the difference between potential phytoplankton biomass/production and actually experienced levels of e.g. phytoplankton biomass in a given estuary. The only remaining explanation is that large populations of benthic suspension feeders, most often bivalves, act as a phytoplankton sink. A variation over the same theme is changes in time in phytoplankton concentration coincident with invasion of a suspension feeder. A classic example is the studies of San Francisco Bay (Cloern 1982, Nichols 1985, Alpine & Cloern 1992). Using a simple model describing change of phytoplankton concentration where dispersive transport and zooplankton grazing balance growth rate, calculated concentrations of phytoplankton were much higher than actually observed concentrations in South San Francisco Bay (Cloern 1982). This led to the conclusion that some other factor had to control phytoplankton biomass. By analysing benthic fauna composition and estimating the filtration capacity of the benthic suspension-feeders, primarily clams, it was shown that these had the capacity to clear the water column more than once per day and concluded that benthic infauna thus control phytoplankton in South San Francisco Bay. It could further be shown that invasion of a non-indigeneous clam in the northern part of the bay, resulted in persistently low levels of chlorophyll a (fig. 1, (fig. 28b) and a 5-fold reduction of primary production (Alpine & Cloern 1992). Similarly Hily (1991) stated that in the eutrophic Bay of Brest, phytoplankton biomass remained low in summer and fall most likely due to huge populations of benthic suspension feeders that have the potential of filtering 30% of the bay volume per day. In a similar approach but covering a longer time period, Chauvaud et al supported the view that benthic suspension feeders, and in particular the gastropod Crepidula fornicata, mitigate the effects of eutrophication by keeping phytoplankton biomass low despite increased nutrient input to the Bay of Brest. Møhlenberg (1995) demonstrated in a shallow nutrient rich Danish cove how stratification stimulated phytoplankton growth, whereas vertical mixing resulted in loss of phytoplankton biomass. The decreases in phytoplankton biomass were unaccounted for by changes in net community growth rate of the phytoplankters and were explained by grazing by a large population of mussels. 1.3

Enclosure experiments

Another approach to the determination of potential benthic control of phytoplankton has been the use of enclosures or mesocosms. The assumption is that in situ studies are difficult to reduce to a series of conclusive measurements and that laboratory experiments cannot necessarily be translated into natural ecological events. Hence by enclosing natural estuarine water and manipulating with expected controlling factors like nutrient supply and/or benthic grazing more specific but ecological relevant answers could be obtained. A number of different mesocosm experiments have been carried out with the aim studying the effect of benthic grazing on the phytoplankton community. The systems encompass a huge range in enclosed volume, duration, mixing and how suspension feeding is induced (Riemann et al. 1988, Sullivan et al. 1991, Olsson et al. 1992, Graneli et al. 1993, Prins et al. 1995). The experiments tend

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to give variable results due to unforeseen incidences in individual enclosures, but the general pattern is that benthic suspension feeding reduces overall levels of phytoplankton biomass. Through excretion of waste products, the suspension feeders can further enhance phytoplankton growth rates (Dame 1996) and lead to a succession in the planktonic community towards small opportunistic species (Riemann et al. 1988). Thus, suspension feeders, supporting the notion that they play a key role in shallow estuarine habitats, can mediate both top-down and bottom-up control of phytoplankton.

Figure 1. Time series of chlorophyll a concentration and mean abundance of the nonindigenous clam Pomatocorbula amurensis before and after its introduction to northern San Francisco Bay (from Cloern 2001).

1.4

Freshwater experiences

In freshwater research, benthic grazing on phytoplankton had in general received less attention until the invasion of zebra mussels (Dreissena polymorpha) in North American freshwater lakes, rivers and streams. Since their first appearance in the mid 1980’s several studies have been carried out demonstrating a number of effects (Rutherford et al. 1999): a) reduced levels of phytoplankton or chlorophyll a concentration irrespective of eutrophication status (e.g. Fahnenstiel et al. 1995a); b) increased water clarity stimulating growth of aquatic macrophytes (Fahnenstiel et al.

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1995b) and changing stratification patterns (Yu & Culver 2000); c) increase in the frequency and duration of blue-green algal blooms (Fahnenstiel et al. 1995a, Nalepa & Fahnenstiel 1995); d) a general shift in energetic pathways from the pelagic to the benthic organisms (Rutherford et al. 1999) thereby e) affecting zooplankton biomass and composition (Lavrentyev et al. 1995, Horgan & Mills 1999, Idrisi et al. 2001); and f) altering the benthic communities (Lozano et al. 2001). The studies of the impact of zebra mussels on ecosystems and accompanying studies on other freshwater benthic suspension feeders have followed more or less the same pattern as marine investigations with focus either on the bivalves and their grazing capacity or on the water column parameters. Due partly to the enforcement of the study objective to the scientific community by the invasion of the mussels and partly to the blessings of access to pre- and post-invasion data, there has been a tendency towards more integrated studies, that clearly demonstrates the ecosystem effects of benthic grazing.

Figure 2. The concentration boundary layer which develops over a blue mussel reef in a unidirectional flow, Z, water depth; X, mussel bed length; C0, initial seston concentration, from Wildish & Kristmanson (1997).

2. PHYSICAL FORCING Despite the differences in scientific approach it is evident that benthic suspension feeders play a dynamic role in the structure and function of coastal ecosystems. In addition, some species are of economic interest and the harvest or cultivation of these species requires management. Structurally, benthic suspension feeders can influence the bathymetry of creeks, provide new habitats in the form of dense epibenthic beds, increase bottom roughness and foul all hard substrata including man made installations. Functionally, they have the potential of mitigating effects of eutrophication by limiting phytoplankton and seston concentrations in the water column and thereby alter trophic structures. But some of these roles are speculative due to the shortcomings of the scientific approaches. When asking the animals, it is thus obvious that laboratory derived filtration rates are not necessarily applicable to natural conditions. Hence, even in shallow systems the grazing potential is not realized as evidenced by the presence of chlorophyll a in areas with a benthic grazing

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potential exceeding the water volume of the system up to several times. The attempts to duplicate the complexity of the estuarine ecosystem in mesocosms or enclosures also suffer from weaknesses, notably the lack of sufficient replication, their vulnerability to unforeseen and chaotic incidences in one treatment/enclosure and the problems with physical mixing and edge effects due to the walls. And asking the water has so far been done more rarely, giving partly correlative answers or been local rather than global in significance. These shortcomings are very much related to the physical dimension of the habitats of marine benthic suspension feeders, i.e. water movement and stratification. For that reason, freshwater experiences cannot be directly applied to coastal areas due to the absence of tidal currents and weaker and less persistent (only thermal) stratification patterns in freshwater systems. If the analogy of benthic suspension feeders in a eutrophic estuary is a vacuum-cleaner and that these unlike a real vacuum-cleaner are sessile, it is not trivial how, how fast and how much of the carpet is swept over the vacuum-cleaner if one wants a clean floor.

2.1

Benthic suspension feeding in tidally mixed areas

From the management and research of exploited bivalve stocks it has become evident, that these animals are dependent on flow of water to supply them with food, i.e. hydrodynamically limited (Wildish & Kristmanson 1997). It is intuitively obvious that in still water, benthic suspension feeders will rapidly deplete the part of the water column they can access and thus rapidly become food limited. A sustainable size of a population of benthic suspension feeders is thus dependent on the rate of food supplied by advection, besides from what can be obtained by bio-mixing resulting from exhalent currents (Larsen & Riisgård 1997). Assuming a balance between turbulent mass transfer of seston and consumption at the boundary in a well developed boundary layer, Wildish & Kristmanson (1979) proposed that bed consumption rate is dependent on benthic grazing potential: NA = BRαC

(1)

where NA is the food flux in g m-2 h-1, B is the biomass of suspension feeders in g m, R is clearance capacity in m3 g-1 h-1 and α is a dimensionless arbitrary expression of filtration efficiency, indicating the degree of re-filtration and/or irregular filtration patterns. Flux of organic material to the bottom can be defined by free stream seston concentration and current velocity:

2

NA = γufree (C0 – C’)

(2)

where ufree is free stream current velocity in m h-1, C0 and C’ respectively upstream and effective food concentration in g m-3 and γ is a dimensionless hydrodynamic parameter related to bottom roughness. Equations 1 and 2 can be combined: (C0 – C’)/C’ = BRα/γufree

(3)

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The right-hand side of the equation is the ratio of filtration capacity to the turbulent mixing intensity, or the seston depletion index (Wildish & Kristmanson 1997). When it exceeds a certain threshold value, depletion effects can be assumed to occur (Herman et al. 1999). Upstream or at the edges in bidirectional tidal flows, bed consumption rate and flux of organic matter to the bottom may not be in balance and the seston concentrations near the bottom will vary with depth and distance from the leading edge. But a depletion effect will gradually build up resulting in the formation of an increasing concentration boundary layer (fig. 3) that can be adequately described by equation 3. Further model studies (Wildish & Kristmanson 1984, Muschenheim 1987, Frechette et al. 1989) have shown that the concentration boundary layer will depend on current velocity, mussel bed path length and water depth. In general, the models predict that decreasing current speed and increasing mussel bed length will lead to an increased concentration boundary layer, finally reaching the surface, and to larger vertical gradients in seston concentration. Both vertical and horizontal seston profiles over experimental mussel, found in flume studies (Wildish & Kristmanson 1984, Butman et al. 1994), have more or less accurately confirmed the model predictions. Some evidence in favour of this general model has also been presented. A number of studies have thus shown that mussels near the edges of mussel beds have higher growth rate and higher meat contents than mussels in the centre of patches (Newell 1990). Similarly, Fréchette & Bourget (1985b) demonstrated that mussels raised 1 m above the bottom grew significantly better than mussels on the bottom and explained it by differences in food concentration (Frechette & Bourget 1985a). Few others have been able to demonstrate vertical seston depletion (fig. 2) above dense beds of benthic suspensions-feeders (Frechette et al. 1989, Muschenheim & Newell 1992), but the results were ambiguous due to e.g. increased vertical mixing by wind (Frechette & Bourget 1985) or varying current speeds (Frechette et al. 1989). Measurements of near-bed vertical depletion of chlorophyll a rarely give uniform results due to resuspension and biodeposition. Demonstration of vertical profiles is, however, not limited by practical problems only. The model assumptions of e.g. homogeneous roughness will not hold in natural mussel beds, not even in infaunal suspensions-feeder beds (O'Riordan et al. 1995). Inhomogeneous roughness can thus emerge from differences in 3D-structure or gaps in mussel beds (Frechette et al. 1989), but is also a result of exhalent jet currents that act as solid objects protruding from the bed and resulting in an increase in bed roughness of up to 2 orders of magnitude compared to a flat bed (Duren et al. 2002). Hence, these animals can significantly affect their food supply by altering their physical environment (Herman et al. 1999). Horizontal depletion of phytoplankton or seston has been demonstrated in tunnels and in situ flumes (e.g. Asmus et al. 1992). However, enclosing a body of water in a tunnel or a flume may produce artificial and unrealistic hydrodynamic effects, unless care is taken in the design of the flume or tunnel. Early field observations have also demonstrated horizontal depletion over populations of bryozoans and sponges (Buss & Jackson 1981) or in shallow coves (Carlson et al. 1984), but in situ studies are few and rarely demonstrate effects on an ecosystem level. Noren et al. (1999) and Haamer & Rodhe (2000) showed that water passing a mussel bed in the Sound between Sweden and Denmark was depleted of chlorophyll a and phytoplankton cells in the

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entire height of the fully mixed water column over a several km long more or less continuous mussel bed, but experimental studies on a larger scale are in general rare.

100

10

1

0,1 0

0,5

1

1,5

2

2,5

3

Chl. a, µg/l

Figure 3. Depletion of chlorophyll a as a function of height (cm) over a mussel bed in macrotidal Oosterschelde. Relation between logarithm of depth and concentration of chlorophyll a is indicated by a line. Redrawn from Møhlenberg & Petersen (1998).

2.2

Suspension feeding in micro-tidal areas

In micro-tidal areas with low current velocities of 5-10 cm s-1 or less many of the same characteristic features, like concentration boundary layers or horizontal profiles, of coastal areas with higher current velocities are seen. Vertical profiles of chlorophyll a above benthic suspension feeder beds have thus been observed in a number of studies (Riisgård et al. 1996a, Riisgård et al. 1996b, Vedel 1998, Dolmer 2000b, Dolmer 2000a, Petersen et al. 2000). In Kertinge Nor, Riisgård et al. (1998) demonstrated the occurrence of density driven currents in periods with large differences in density at the borders of the cove. In the cove proper this resulted in stratification with relative well defined, but opposing, currents above and below the pycnocline with current velocities around 1 cm s-1. Below the pycnocline, but just above the population of Ciona intestinalis, depletion of chlorophyll a in the passing water mass could be observed. Advective transport generated by tide or density driven currents is, however, not the only structuring factor in micro-tidal shallow areas. Wind generated mixing is thus an important means of providing food to dense populations of benthic suspension feeders. In a shallow semi-enclosed basin over mussel beds it could thus be shown

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that stratification lead to vertical gradients in chlorophyll a and build up of phytoplankton biomass, whereas destratification leads to uniform chlorophyll a concentrations in the entire water column and decreasing biomass (Møhlenberg 1995). It was further shown that vertical mixing was primarily caused by larger wind events where mean wind of 4-6 m s-1 through 12 hours was sufficient to break down stratification. In Kertinge Nor it could be demonstrated that filtration by the dense population of ascidians resulted in vertical profiles of chlorophyll a in the water column even during fully mixed conditions. It could further be demonstrated that the steepness of the seston depletion was dependent on the degree of mixing and duration of respectively stratified and mixed conditions (fig. 4). Stratification was found to be dependent primarily on wind conditions and to a lesser extent on advective currents. Further, the height of the concentration boundary layer was dependent on whether the ascidians were attached to eelgrass leaves, and thus lifted into the water column, or were lying on the bottom in clumps. Other studies have demonstrated the importance of vertical mixing to phytoplankton blooms in shallow estuaries with huge biomass of benthic suspension feeders (Cloern 1991, Koseff et al. 1993). An important implication of the occurrence of dense populations of benthic suspension feeders in micro-tidal and/or unexposed areas is that flow does not necessarily restrict these populations in the classical sense. If sufficient food is produced due to eutrophication and that the food can be made accessible to the benthic community by mixing due to wind/waves, dense populations can be maintained as seen in e.g. Danish estuaries (Conley et al. 2000). 2.3

Plant canopies

Plant canopies like seagrass meadows or kelp forests are generally believed to alter the current pattern by decreasing current velocity within the canopy but increasing it just above and on the sides of the canopy. Concurrent with the decrease in current speed within the canopy or seagrass bed, turbulence is increased especially at the top of the leaf canopy. The reduced flow inside the canopy may explain enhanced sedimentation rates (Gacia et al. 2002) and reduced resuspension (Terrados & Duarte 2000). The microcosm of particles within a seagrass canopy has been shown to have an increased biovolume of particles compared to the above canopy concentrations, but with a larger fraction of detrital material of both angiosperm and planktonic origin (Duarte et al. 1999). Plant canopies may thus act as both sources and sinks of particles. The effects of plant canopies on benthic suspension feeders are debated and previous results are not unequivocal (Wildish & Kristmanson 1997). Some results show increased growth, others decreased growth compared to outside canopies. The contradicting results may have several explanations. Differences in growth response can be a result of physiological differences in the ability of different suspension feeders to digest and utilize seagrass or macro-algae material or differences between plant species in nutritious value. Alternatively, attachment of the suspension feeders to the leaves of the plants could have lifted the benthic suspension feeders differently up from the bottom and thus differently away from the food-limited concentration boundary layer in the different investigations.

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3. EUTROPHICATION AND BENTHIC SUSPENSION FEEDING Eutrophication may interact with benthic suspension feeders at several levels: 1) increased nutrient input may lead to increased primary production and standing stock of phytoplankton biomass, thereby potentially enhancing benthic secondary production or carrying capacity; 2) benthic suspension feeders may mitigate adverse effects of eutrophication like increased turbidity and oxygen depletion in bottom waters by grazing on phytoplankton cells and thus increase water transparency and diminish sedimentation; and 3) by suspension feeder removal of phytoplankton cells, stored nutrients will to some extent become readily available to production through excretion and may thus alter plankton composition. 3.1

Food limitation and carrying capacity

It can intuitively be appreciated that on the level of the patch of benthic suspension feeders, food limitation is predominantly a result of physically mediated food supply. If turbulence, vertical mixing and horizontal advection are sufficiently high, competitive interactions will be of minor importance and the size of the patch will depend on the physical forcing. On a system scale, benthic suspension feeder biomass and production will, however, be limited of both physical constraints and system (primary) productivity (Heip et al. 1995) or productivity of adjacent systems. There are various sources of evidence for the tight coupling between system primary productivity and secondary production of benthic suspension feeders. It has thus been shown that biomass of both overall zoobenthos (Beukema & Cadée 1986, Beukema et al. 2002) - and specifically suspension feeder (Beukema & Cadée 1997) - biomass and production could be related to pelagic phytoplankton biomass and production. It could further be shown that growth of a dominant suspension feeding bivalve was related to plankton cell concentration during the growing season (Beukema & Cadée 1986, Beukema & Cadée 1991, Beukema et al. 2002). Similarly, a tight correlation between fluctuations in annual primary production and mussel growth, indicated by condition index, was demonstrated in one of the main mussel production sites in Holland, the Oosterschelde (Smaal & van Stralen 1990, Smaal et al. 2001). Benthic secondary production has also been correlated directly to nutrient supply in e.g. Ria de Arosa, where growth and condition of cultivated mussels are closely related to the annual average upwelling index (Blanton et al. 1987), indicating the degree of nutrients transported from deep Atlantic waters to the Galician estuaries through upwelling. Further, Josefson & Rasmussen (2000) demonstrated dependency of benthic biomass on nutrient load from land run-off (fig. 5), with increased nutrient loading resulting in increased biomass of particularly suspension feeders that constituted >90% of the biomass. Above a nutrient load of approximately 125 g Ntot m-2 yr-1 biomass decreased, probably due to increased occurrence of oxygen depletion or persistent stratification. The relation was derived for a number of different estuaries and was further improved, if residence time of the estuaries were taken into consideration. On an ecosystem level, carrying capacity can be defined as the population size that can be supported by a given system and will depend on water residence time and primary

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production of the system in relation to the suspension feeder biomass and its clearance capacity (Dame & Prins 1998).

Chl. a depletion, µg l -1 cm-1

3 2,5 2 1,5 1 0,5

-100

-80

-60

-40

0 -20 0 hours

20

40

60

80

Figure 4. Chlorophyll a depletion (slope of the linear relation between depth and concentration of chl. a) as a function of hours of mixed conditions before sampling. Negative xvalues indicate hours of mixed conditions and positive x-values indicate hours of stratified conditions prior to sampling (Petersen et al unpublished data)

A number of different coastal ecosystems have been compared with regard to these parameters (Smaal & Prins 1993, Heip et al. 1995, Dame & Prins 1998). The systems encompass a huge range in physical structure with regard to area, average depth, volume, tidal range and residence time and primary production. In general the systems fell in a range from small fast systems with short residence time supporting denser populations to larger systems with longer residence times and lower biomass (Dame & Prins 1998). A surprisingly good correlation between volume specific biomass of benthic suspension feeders and water residence time of a number of primarily tidal estuaries could thus be demonstrated (Smaal & Prins 1993, Heip et al. 1995). Including variables like primary production time, defined as annually averaged phytoplankton biomass to primary production, and clearance time, defined as the time needed for the suspension feeder population to clear the entire water column, some of the systems were outliers (Dame & Prins 1998).

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Figure 5. Plot of mean total biomass (AFDW) of benthic biomass against total area specific Nload for 14 Danish estuaries (From Josefson & Rasmussen 2000, printed with permission from Elsevier).

Outliers were systems with low primary production time due to e.g. high turbidity like the Bay of Marennes-Oléron, or systems with low clearance time in relation to primary production time due to e.g. large areas without benthic suspension feeders as in (the present) Chesapeake Bay. When recalculating the data by Josefson & Rasmussen (2000), it can further be seen that shallow Danish estuaries sustain a higher volume specific biomass of benthic suspension feeders than what can be inferred from the relation found by Heip et al. (1995). Both the outliers and the Danish estuaries point to the same conclusion. Carrying capacity on the system level will depend on both a biological component, i.e. the primary production, and a physical component, that cannot be reduced to residence time, but must include some measure of contact rate between the water column and the suspension feeding populations. The physical contact rate will depend on mixing characteristics like current speed and bottom roughness, but also on wind regime and bathymetry i.e. the ratio of shallow shoals to deeper channels. In addition, contact rate will depend on the physical placement of the suspension feeders, whether as a normal benthic population, attached to macro-algae or seagrasses or cultivated on suspended culture units. A special feature of some systems is that clearance time is shorter than primary production time or water residence time. Thereby it is indicated that for the populations to be sustained, food must be supplied either by advection or from

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unaccounted food sources. Resuspended material has been shown to be of importance to benthic suspension feeder growth (Emerson 1990, Smaal & Zurburg 1997, Coma et al. 2001). The resuspended material may be composed of epibenthic micro-algae and detritus including material of macro-phytobenthic origin and may potentially constitute a major part of the food supply to benthic suspension feeders (Héral 1993). For budget calculations and in relation to primary production time and thus carrying capacity of estuaries it is worthwhile noting, that only rarely is epibenthic primary production included in estimates of system primary production. The contribution of non-phytoplankton material to suspension feeder diet and how this relates to water column mixing and primary production must be included in carrying capacity models for specific areas that are commercially exploited.

3.2

Benthic suspension feeders as eutrophication control

In the comparison of different ecosystems mentioned above, some of the ecosystems had lower clearance time compared to water residence time (Dame 1996). The implication of this is not only that food for the benthic suspension feeders potentially have to originate from outside the system, but also that the benthic suspension feeders potentially regulate system phytoplankton biomass and production. If in addition, clearance time is approximately equal to primary production time, then the systems will be even more susceptible to benthic control (Dame & Prins 1998, Prins et al. 1998). Evidence of top-down control by benthic suspension feeders under eutrophic conditions has been demonstrated in various ways. In a series of experiments with land-based, stirred mesocosms (Prins et al. 1995, Prins et al. 1999, Escaravage & Prins 2002) it could thus be shown that even under conditions with high external nutrient load, mussel grazing may exert an effective control on phytoplankton biomass when mussel clearance capacity corresponds to 20-35% of the volume of the mesocosms per day, but that structuring effects were negligible at a clearance capacity of 15% or less (Prins et al. 1995). An important point of the experiments was that with a strong top-down control, the ecosystem was independent of increased nutrient loading and phytoplankton biomass remained low. A similar conclusion was pointed out by Herman & Scholten (1990) in a modelling study, showing that increasing nutrient loading under high suspension feeding pressure only marginally changed phytoplankton concentration, implying that benthic grazing can act as eutrophication control. As a consequence thereof, substantial changes in suspension feeder biomass can be expected to have dramatic consequences for water column parameters, notably chlorophyll a. In the absence of benthic suspension feeders, grazing on phytoplankton will be performed by zooplankton. The developmental lagphase of zooplankton renders them, however, less efficient as phytoplankton control (Herman & Scholten 1990, Prins et al. 1998). The lower metabolic rates of benthic suspension feeders in contrast enable them to survive periods of low food, and they can thus be present and ready to ingest when phytoplankton blooms. On the scale of estuaries, studies are few but invasion of non-indigenous species has been shown to reduce average phytoplankton concentrations dramatically (Nichols 1985, Alpine & Cloern 1992). In the northern San Francisco Bay mean chlorophyll a

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concentration was normally dependent on mean river flow, but in 1976-77 and again since 1986 that relation was no longer valid and chlorophyll a concentrations were dramatically reduced (Alpine & Cloern 1992, Cloern 1996). In both cases the reduction could be related to invasion of clams, in 1976-77 primarily Mya arenaria and in 1986 by Potamocorbula amurensis, since neither nutrient nor light limitation could explain the decrease. On a shorter time scale, breakdown of stratification has been shown to lead to decrease in phytoplankton biomass (Cloern 1991, Koseff et al. 1993, Møhlenberg 1995). Similarly, decoupling of benthic grazing due to die off of a large part of the benthic suspension feeders (Riisgård et al. 1995) or to stratification (Cloern 1991, Koseff et al. 1993, Møhlenberg 1995) result in increase in phytoplankton biomass. And it has been shown that biomass of mussels is a better predictor of chlorophyll a concentration in Danish estuaries than any quantity related to nutrient fluxes and concentrations (Kaas et al. 1996). Chlorophyll a was thus inversely related to mussel biomass in these estuaries, which also could be demonstrated in 15 Canadian estuaries (Meeuwig 1999). Data on suspension feeder control of phytoplankton biomass and production primarily originate from well-mixed shallow ecosystems with high benthic biomass. These conditions along with requirements of no nutrient or light limitation of phytoplankton production are the basic assumptions of the Officer et al. (1982) model that use an idealized Lotka-Volterra approach of population interactions: τg = H/RB = τp = 1/µp

(4)

where τg is the grazing and τp is production time scale, H is the system water depth in m, R is specific clearance rate of the benthic suspension feeders in m3 g-1 d-1, B is biomass of benthic suspension feeders in g m-2 and /µp is specific phytoplankton net growth rate in d-1. Equilibrium between benthic suspension feeders and phytoplankton is reached when the grazing time-scale equals the production timescale, which in the case of South San Francisco Bay can be reached with a biomass of suspension feeders of a few hundred g m-2 resulting in a phytoplankton biomass of a few µg chlorophyll a l-1. This is a simple but realistic situation for many well-mixed estuaries (Herman et al. 1999). Focusing on phytoplankton bloom formation in the presence of benthic suspension feeders, Koseff et al. (1993) showed that vertical mixing of the water column is a critical process, which was not accounted for by Officer et al. (1982). In non-stratified waters with ample nutrient and light availability, the possibility of a bloom will depend on the ratios of the production time-scale to the grazing time-scale and to the vertical (turbulent) mixing time-scale (Koseff et al. 1993). Benthic suspension feeders can thus prevent phytoplankton blooms if mixing is vigorous and grazing potential is sufficiently high compared to primary production. In a more detailed, but qualitatively similar model, Lucas et al. (1998) showed that even a mild form of permanent stratification affects the system and renders benthic grazing unimportant in predicting bloom formation. Benthic control is thus confined to shallow areas, though it may be significant down to 15 m depending on actual conditions of light and stratification (Lucas et al. 1998). Bloom formation is, however, not only dependent on the local production-loss balance but is also affected by large-scale horizontal transport mechanisms (Lucas et al. 1999a, Lucas et al. 1999b).

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Apart from the physical mixing, an important prerequisite for benthic suspension feeders to act as eutrophication control is the open nature of the estuarine systems. If there are no effective removal mechanisms, there will be an ever-increasing level of phytoplankton when nutrients accumulate in a system, despite high grazing pressure (Heip et al. 1995). Removal can be mediated trough different mechanism like removal of excretion products by tidal currents, in the form of temporarily stored nutrients released during autumn and winter (Heip et al. 1995) or by harvesting the suspension feeders. In the highly eutrophic Mariager Fjord empirical modelling based on monitoring data has shown that phosphorous has become limiting for primary production during spring and early summer, while nitrogen is limiting in late summer and fall (Nordjyllands Amt & Århus Amt 2002). The local authorities have set local goals to reduce nutrient loading beyond the national goals in order to minimize the risk for major anoxia events. Since additional limiting of nutrient load is difficult for technical, economical or political reasons, the local authorities have promoted a plan for cultivating blue mussels, Mytilus edulis, with the aim of removing nutrients. With an annual harvest of 6,000 t live mussels, the local goal can be reached, showing that cultivating suspension feeders can be a realistic management tool in mitigating effects of eutrophication. 3.3

Nutrient feed-back and ecosystem structuring

Benthic suspension feeders not only remove phytoplankton and other particulate organic matter from the water column. Specific organic constituents in the form of phytoplankton cells and detritus consumed as food are processed and deposited as faeces and pseudofaeces or excreted as fundamental nutrients. As an example, Jordan & Valiela (1982) made a nitrogen budget for the ribbed mussel, Geukensia demissa, in a salt marsh. They showed that at the level of the marsh-estuarine ecosystem, the mussels filtered 1.8 times the particulate nitrogen exported yearly from the marsh by tidal flushing. Of the N filtered from the water column, approximately half was absorbed and half was deposited as faeces or pseudofaeces. Of the absorbed nitrogen, 55% was excreted as ammonia. It was calculated that the marsh mussels released more ammonia into the system than any other component and that since nitrogen limits the productivity of the marsh, re-mineralization by the mussels may ultimately increase productivity. In the Askö area of the Baltic Sea, Kautsky & Wallentinus (1980) measured in situ excretion of blue mussels and found that the regenerated nutrients constitute a substantial input to both the pelagic and benthic system. The excretion from mussels could explain why many of the benthic red and brown macroalgae could extend their growth maximum after nutrients are depleted in the water mass. Thus, on one side mussels increase retention of nutrients within an estuarine system, by filtering particulate material and depositing it as biodeposits (Kautsky & Evans 1987) or long-lived mussel biomass. On the other side nutrients stored in particulate planktonic material are transformed to readily accessible inorganic nutrients by being processed by suspension feeders, thereby fuelling and thus promoting primary production (Asmus & Asmus 1991).

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Figure 6. Amount of nitrogen mineralized by mussel beds in central Oosterschelde compared to total benthic and pelagic nitrogen mineralization (median and 10-90% range) from the Oosterschelde ecosystem model SMOES (from Prins & Smaal 1994).

Apart from laboratory studies (e.g. Bayne 1976), excretion and regeneration of nutrients have been studied in mesocosms (e.g. Doering et al. 1987) and in situ in chambers or plastic bags (e.g. Kautsky & Wallentinus 1980, Jordan & Valiela 1982) and in tunnels or flumes (e.g. Asmus et al. 1992). An extensive review can be found in Dame (1996). In general, suspension feeders excrete ammonia and phosphate, whereas mineralization of biodeposits leads to release of nitrate, nitrite and silicate (Prins & Smaal 1994). Direct excretion can account for 17-94% of total ammonia and 37-85% of total phosphate release from mussel beds (Prins & Smaal 1994) and in stable mussel and oyster beds, nitrogen release approximately balances input, whereas a part of the phosphorous apparently is retained (Prins et al. 1998). The importance of nutrient cycling by suspension feeders to phytoplankton production will depend on general nutrient conditions in the estuary and mixing conditions, but when turnover time of nutrients are lower than water retention time of the estuary, it is probable that it is significant to the system (Dame 1996). Using model runs, Prins & Smaal (1994) made an estimate of the amount of nitrogen mineralized by mussel beds compared to total system mineralization (fig. 6) and found that it constitutes approximately 50%. In addition, below suspended bivalve culture units, the organic

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enrichment of the sediments can lead to anaerobic conditions and thereby promote denitrification (Kaspar et al. 1985). Bivalve cultures could thus function as a sink of nitrogen not only through harvest of the bivalves but also due to loss of nitrogen to the atmosphere. Grazing by suspension feeders and the accompanying nutrient release predominantly as easily accessible nutrients can be expected to alter phytoplankton species composition and succession patterns. In general it can be expected that faster growing small algal species will be favoured (Furnas 1990, Riegman et al. 1993) from grazing, but also enhanced light availability in the water column from decreased POM concentrations can be expected to influence phytoplankton composition and turnover (Dame 1996). Increased phytoplankton growth (Doering et al. 1987) and shifts in community composition towards smaller, faster growing species (Riemann et al. 1988, Olsson et al. 1992, Graneli et al. 1993, Prins et al. 1995) have been demonstrated in mesocosm experiments. In the field, Noren et al. (1999) demonstrated changes in phytoplankton composition in a water mass passing a mussel bed, but field experiments from marine areas are few. Experiences from freshwater systems are more abundant and invasion of zebra mussels has been shown to cause complex changes in phytoplankton species composition and community structure (e.g. Fahnenstiel et al. 1995a, Lavrentyev et al. 1995, Smith et al. 1998, Rutherford et al. 1999) though results are not identical between systems with regard to e.g. post-dreissenid dominance of cyanobacteria (Smith et al. 1998). There is also evidence of effects of benthic grazing on the zooplankton community, although more scarce. Effects may be mediated through competition for food or by direct grazing of benthic suspension feeders on zooplankton. Mesocosm experiments have shown that micro-zooplankton populations can be directly affected by bivalve grazing (Horsted et al. 1988, Sullivan et al. 1991, Prins et al. 1995) and some indications of ingestion of copepods or their development stages by mussels have been reported (Kimmerer et al. 1994, Davenport et al. 2000). Freshwater zebra mussels have been shown to affect ciliates (Lavrentyev et al. 1995) and daphnids (Horgan & Mills 1999, Idrisi et al. 2001) though results are not clear. Filtration and ingestion of larger zooplankton forms must depend on their ability to escape inhalant siphon currents (Titelman 2001, Jakobsen 2002) and to the degree these are interfered by turbulence (Singarajah 1975).

4. WHERE TO GO FROM NOW? In summary, we now have a strong sense that benthic suspension feeders can act as a filter in relation to eutrophication of shallow estuaries (Cloern 2001). This sense stems from a series of different types of approaches that despite their individual strengths and weaknesses together are convincing in demonstrating the importance of benthic suspension feeders in the flux of particulate matter and dissolved nutrients. In addition to the methodological problems discussed, it is hopefully apparent from the present review, that the knowledge of the ecological role of benthic suspension feeders in estuaries are limited both geographically, historically and quantitatively. In

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an extensive review of ecological importance of bivalves (Dame 1996, Dame & Prins 1998), 15 estuaries were listed, from which bivalve impact could be appreciated. Generally speaking, the vast majority of information stem from a few selected and not necessarily representative estuaries and definitely with an under representation of micro-tidal areas. As our knowledge is limited geographically, it is also apparent that most information was gathered up to around 1990 and that the last decade has produced little and mostly review-like information, especially on the part dealing with benthic grazing impact on phytoplankton biomass and production. There is thus a need of a proliferation of studies expanding into other geographical areas with different physical and chemical forcing, having other benthic suspension feeders than mussels as key organism and using different scientific approaches, preferably asking both the water and the animals. The existing literature gives ample proof that changing locality or organism provides new insight. Recommendations for future work can be summarized on 3 levels: 1) scientifically a number of questions need to be addressed. The “α-factor”, i.e. the transformation of potential grazing capacity to actual grazing in relation to turbulent conditions needs further attention on all scales: individual, patch and estuary. Impact of benthic grazing (and re-mineralization) on planktonic (including importance for fish populations) and benthic biodiversity is at the moment very little investigated both theoretically and practically. Food sources for benthic suspension feeders and the relative importance of these under various mixing conditions also need further attention, especially the importance of epibenthic micro-algae. Finally, the current controversies on basic physiological rates like clearance and ingestion in relation to POM concentration need to be resolved. 2) For monitoring, new approaches like remote sensing of surface waters from above water or ships of opportunity and modelling need to address physical mixing and benthic grazing in areas with dense populations of benthic suspension feeders. 3) Managers of cultured or fished populations of especially bivalves need to address environmental issues in order to be able to maintain a sustainable exploitation. These issues are the community added value from mussel cultures as means of mitigating effects of eutrophication, as well as the damaging effects on the benthic environment below these units and the possible cascading effects that changes in planktonic diversity potentially can have on e.g. the populations and recruitment of commercial fish in estuaries. Further, when using models as management tool, these should be carefully examined for how physical forcing is taken into account. 5. REFERENCES Alpine, A. E., & Cloern, J. E. (1992). Trophic interactions and direct physical effects control phytoplankton biomass and production in an estuary. Limnology and Oceanography, 37(5), 946-955. Asmus, H., Asmus, R. M., Prins, T. C., Dankers, N., Frances, G., Maass, B., et al. (1992). Benthic-pelagic flux rates on mussel beds: Tunnel and tidal flume methodology compared. Helgoländer Meeresuntersuchungen, 46(3), 341-361. Asmus, R. M., & Asmus, H. (1991). Mussel beds: limiting or promoting phytoplankton. Journal of Experimental Marine Biology and Ecology, 148, 215-232. Bayne, B. L. (1976). Marine mussels: their ecology and physiology. Cambridge, U.K.: Cambridge University Press. Beukema, J. J., & Cadée, G. C. (1986). Zoobenthos responses to eutrophication of the Dutch Wadden Sea. Ophelia, 26, 55-64.

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Beukema, J. J., & Cadée, G. C. (1991). Growth rates of the bivalve Macoma balthica in the Wadden Sea during a period of eutrophication: relationships with concentrations of pelagic diatoms and flagellates. Marine Ecology Progress Series, 68, 249-256. Beukema, J. J., & Cadée, G. C. (1997). Local differences in macrozoobenthic response to enhanced food supply caused by mild eutrophication in a Wadden Sea area: Food is only locally a limiting factor. Limnology and Oceanography, 42(6), 1424-1435. Beukema, J. J., Cadée, G. C., & Dekker, R. (2002). Zoobenthic biomass limited by phytoplankton abundance: evidence from parallel changes in two long-term data series in the Wadden Sea. Journal of Sea Research, 48, 111-125. Blanton, J. O., Tenore, K. R., Castillejo, F., Atkinson, L. P., Schwing, F. B., & Lavin, A. (1987). The relationship of upwelling to mussel production in the rias on the western coast of Spain. Journal of Marine Research, 45, 497-511. Buss, L. W., & Jackson, J. B. C. (1981). Planktonic food availability and suspension-feeder abundance: evidence of in situ depletion. Journal of Experimental Marine Biology and Ecology, 49, 151-161. Butman, C. A., Fréchette, M., Geyer, W. R., & Starczak, V. R. (1994). Flume experiments on food supply to the blue mussel Mytilus edulis L. as a function of boundary-layer flow. Limnology and Oceanography, 39(7), 1755-1768. Carlson, D. J., Townsend, D. W., Hilyard, A. L., & Eaton, J. F. (1984). Effect of an intertidal mudflat on plankton of the overlying water column. Canadian Journal of Fisheries and Aquatic Sciences, 41, 1523-1528. Cloern, J. E. (1982). Does the benthos control phytoplankton biomass in South San Francisco Bay. Marine Ecology Progress Series, 9, 191-202. Cloern, J. E. (1991). Tidal stirring and phytoplankton bloom dynamics in an estuary. Journal of Marine Research, 49, 203-221. Cloern, J. E. (1996). Phytoplankton bloom dynamics in coastal ecosystems: a review with some general lessons from sustained investigation of San Fransisco Bay, California. Reviews of Geophysics, 34(2), 127-168. Cloern, J. E. (2001). Our evolving conceptual model of the coastal eutrophication problem. Marine Ecology Progress Series, 210, 223-253. Coma, R., Ribes, M., Gili, J.-M., & Hughes, R. N. (2001). The ultimate opportunists: consumers of seston. Marine Ecology Progress Series, 219, 305-308. Conley, D. J., Kaas, H., Møhlenberg, F., Rasmussen, B., & Windolf, J. (2000). Characteristics of Danish estuaries. Estuaries, 23(6), 820-837. Dame, R. F. (1996). Ecology of marine bivalves. An ecosystem approach. Boca Raton: CRC Press. Dame, R. F., & Prins, T. C. (1998). Bivalve carrying capacity in coastal ecosystems. Aquatic Ecology, 31, 409-421. Davenport, J., Smith, R. J. J. W., & Packer, M. (2000). Mussels Mytilus edulis: significant consumers and destoyers of mesozooplankton. Marine Ecology Progress Series, 198, 131-137. Doering, P. H., Kelly, J. R., Oviatt, C. A., & Sowers, T. (1987). Effect of the hard clam Mercenaria mercenaria on benthic fluxes of inorganic nutrients and gases. Marine Biology, 94, 377-383. Dolmer, P. (2000a). Algal concentration profiles above mussel beds. Journal of Sea Research, 43(2), 113119. Dolmer, P. (2000b). Feeding activity of mussels Mytilus edulis related to near-bed currents and phytoplankton biomass. Journal of Sea Research, 44(3-4), 221-231. Duarte, C. M., Benavent, E., & Sánchez, M. d. C. (1999). The microcosm of particles within seagrass Posidonia oceanica canopies. Marine Ecology Progress Series, 181, 289-295. Duren, L. A. v., Hendriks, I., Petersen, J. K., Larsen, J., Sikkelerus, P. B. C. v., Herman, P. M. J., et al. (2002). Combined flume and field experiments on boundary layer hydrodynamics over mussel beds. Paper presented at the ASLO Conference, June 2002, Victoria BC. Emerson, G. W. (1990). Influence of sediment disturbance and water flow on the growth of the soft-shell clam, Mya arenaria L. Canadian Journal of Fisheries and Aquatic Sciences, 47, 1655-1663. Escaravage, V., & Prins, T. C. (2002). Silicate availability, vertical mixing and grazing control of phytoplankton blooms in mesocosms. Hydrobiologia, 484, 33-48. Fahnenstiel, G. L., Bridgeman, T. B., Lang, G. A., McCormick, M. J., & Nalepa, T. F. (1995). Phytoplankton productivity in Saginaw Bay, Lake Huron: effects of zebra mussel (Dreissena polymorpha) colonization. Journal of Great Lakes Research, 21(4), 465-475. Fahnenstiel, G. L., Lang, G. A., Nalepa, T. F., & Jahnengen, T. H. (1995). Effects of zebra mussel (Dreissena polymorpha) colonization on water quality parameters in Saginaw Bay, Lake Huron. Journal of Great Lakes Research, 21(4), 435-448.

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Petersen, J. K., Møhlenberg, F., Jonsson, P. R., Mathieu, P.-P., & Burchard, H. (2000). Grazing impact by benthic suspension-feeders in a microtidal estuary. Paper presented at the ASLO Conference 2000, Copenhagen. Petersen, J. K., & Riisgård, H. U. (1992). Filtration capacity of the ascidian Ciona intestinalis and its grazing impact in a shallow fjord. Marine Ecology Progress Series, 88, 9-17. Petersen, J. K., Stenalt, E., & Hansen, B. W. (2002). Invertebrate re-colonisation in Mariager Fjord (Denmark) after a severe hypoxia. II. Blue mussels (Mytilus edulis L.). Ophelia, 56(3), 215-226. Prins, T. C., Escaravage, V., Smaal, A. C., & Peeters, J. C. H. (1995). Nutrient cycling and phytoplankton dynamics in relation to mussel grazing in a mesocosm experiment. Ophelia, 41, 289-315. Prins, T. C., Escaravage, V., Wetsteyn, L. P. M. J., Peeters, J. C. H., & Smaal, A. C. (1999). Effects of different N- and P-loading on primary and secondary production in an experimental marine ecosystem. Aquatic Ecology, 33, 65-81. Prins, T. C., & Smaal, A. C. (1994). The role of the blue mussel Mytilus edulis in the cycling of nutrients in the Oosterschelde estuary (The Netherlands). Hydrobiologia, 282-283(0), 413-429. Prins, T. C., Smaal, A. C., & Dame, R. F. (1998). A review of the feedbacks between bivalve grazing and ecosystem processes. Aquatic Ecology, 31, 349-359. Riegman, R., Kuipers, B. R., Noordeloos, A. A. M., & Witte, H. J. (1993). Size-differential control of phytoplankton and the structure of plankton communities. Netherlands Journal of Sea Research, 31(3), 255-265. Riemann, B., Nielsen, T. G., Horsted, S. J., Bjørnsen, P. K., & Pock-Steen, J. (1988). Regulation of phytoplankton biomass in estuarine enclosures. Marine Ecology Progress Series, 48, 205-215. Riisgård, H. U., Christensen, P. B., Olesen, N. J., Petersen, J. K., Møller, M. M., & Andersen, P. (1995). Biological structure in a shallow cove (Kertinge Nor, Denmark) - Control by benthic nutrient fluxes and suspension-feeding ascidians and jellyfish. Ophelia, 41, 329-344. Riisgård, H. U., Jensen, A. S., & Jürgensen, C. (1998). Hydrography, near-bottom currents and grazing impact of the benthic filter-feeding ascidian Ciona intestinalis in a Danish fjord. Ophelia, 49(1), 1-16. Riisgård, H. U., Jürgensen, C., & Clausen, T. (1996). Filter-feeding ascidians (Ciona intestinalis) in a shallow cove: implications of hydrodynamics for grazing impact. Journal of Sea Research, 35(4), 293300. Riisgård, H. U., Poulsen, L., & Larsen, P. S. (1996). Phytoplankton reduction in near-bottom water caused by filter-feeding Nereis diversicolor - implications for worm growth and population grazing impact. Marine Ecology Progress Series, 141, 47-54. Rutherford, E. S., Rose, K. A., Mills, E. L., Forney, J. L., Mayer, C. M., & Rudstam, L. G. (1999). Individual-based model simulations of a zebra mussel (Dreissena polymorpha) induced energy shunt on walleye (Stizostedion vitreum) and yellow perch (Perca flavescens) populations in Oneida Lake, New York. Canadian Journal of Fisheries and Aquatic Sciences, 56, 2148-2160. Singarajah, K. V. (1975). Escape reactions of zooplankters: effect of light and turbulence. Journal of the Marine Biological Association of the United Kingdom, 55, 627-639. Smaal, A. C., & Prins, T. C. (1993). The uptake of organic matter and the release of inorganic nutrients by bivalve suspension feeder beds. In R. F. Dame (Ed.), Bivalve filter feeders in estuarine and coastal ecosystem processes (pp. 273-298). Berlin: Springer. Smaal, A. C., & van Stralen, M. R. (1990). Average annual growth and condition of mussels as a function of food source. Hydrobiologia, 195, 179-188. Smaal, A. C., van Stralen, M. R., & Schuiling, E. (2001). The interaction between shellfish culture and ecosystem processes. Canadian Journal of Fisheries and Aquatic Sciences, 58(5), 991-1002. Smaal, A. C., & Zurburg, W. (1997). The uptake and release of suspended and dissolved material by oysters and mussels in Marennes-Oleron Bay. Aquatic Living Resources, 10(1), 23-30. Smith, T. E., Stevenson, R. J., Caraco, N. F., & Cole, J. J. (1998). Changes in phytoplankton community structure during the zebra mussel (Dreissena polymorpha) invasion of the Hudson River (New York). Journal of Plankton Rsearch, 20(8), 1567-1579. Sullivan, B. K., Doering, P. H., Oviatt, C. A., Keller, A. A., & Frithsen, J. B. (1991). Interactions with the benthos alter pelagic food web structure in coastal waters. Canadian Journal of Fisheries and Aquatic Sciences, 48, 2276-2284. Terrados, J., & Duarte, C. M. (2000). Experimental evidence of reduced particle resuspension within a seagrass (Posidonia oceanica L.) meadow. Journal of Experimental Marine Biology and Ecology, 243, 45-53. Titelman, J. (2001). Swimming and escape behaviour of copepod nauplii: implications for predator-prey interactions among copepods. Marine Ecology Progress Series, 213, 203-213.

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Vedel, A. (1998). Phytoplankton depletion in the benthic boundary layer caused by suspension-feeding Nereis diversicolor (Polychaeta): grazing impact and effect of temperature. Marine Ecology Progress Series, 163, 125-132. Wildish, D. J., & Kristmanson, D. D. (1979). Tidal energy and sublittoral macrobenthic animals in estuaries. Journal of Fisheries Research Board of Canada, 36, 1197-1206. Wildish, D. J., & Kristmanson, D. D. (1984). Importance to mussels of a benthic boundary layer. Canadian Journal of Fisheries and Aquatic Science, 41, 1618-1625. Wildish, D. J., & Kristmanson, D. D. (1997). Benthic suspension feeders and flow. New York: Cambridge University Press. Yu, N., & Culver, D. A. (2000). Can zebra mussels change stratification patterns in a small reservoir? Hydrobiologia, 431, 175-184.

6. AFFILIATIONS J.K. Petersen: Department of Marine Ecology, National Environmental Research Institute, P.O. Box 358, DK-4000 Roskilde, Denmark.

JUST CEBRIÁN

GRAZING ON BENTHIC PRIMARY PRODUCERS 1. INTRODUCTION The importance of herbivores as controls of nutrient storage as producer biomass depends to a great extent on the percentage of primary production they consume. This is so because the quantity of primary production that remains stored as producer biomass decreases as consumption by herbivores increases. Hence, herbivores may keep low levels of producer biomass through intense consumption (Carpenter 1986, Valentine & Heck 1991, Cebrian & Duarte 1994). Similarly, the contribution of herbivores to producer-bound nutrient recycling through faeces excretion and exudation also increases with the percentage of primary production consumed. Herbivore excreta are usually nutrient-rich in relation to producer detritus and thus are quickly decomposed by detritivorous organisms. Thus, large percentages of primary production consumed can significantly increase the rate of nutrient recycling in the system (Elser & Urabe 1999, Sterner & Elser 2002). The existing literature indicates that the intensity of herbivory on benthic producers can be very variable. For instance, the percentage of primary production removed by herbivores ranges from negligible values to ca. 100% in communities of microphytobenthos (i.e. benthic microalgae that live in the sediment; Montagna 1984, Baird & Ulanowicz 1993), macroalgae (Mann 1972, Hagen 1995, Ferreira et al. 1998) and seagrasses (Cebrian & Duarte 1998, Valentine & Heck 1999, Valentine et al. 2000). It therefore seems that herbivores can play a variable role as controls of the dynamics of producer biomass and nutrient recycling in coastal communities. Unfortunately, there have been few comparisons of the variability in herbivory within and across types of benthic producers (Valiela 1995, Duarte & Cebrian 1996, Alongi 1998, Cebrian 2002). In addition, most of the existing comparisons are not extensive. This has impaired our ability to elucidate if there are any patterns in how herbivory varies within and across types of benthic producers and to identify the controls and consequences of that variability. It is clear that herbivores may play an important role in the storage and trophic transference of carbon and nutrients in coastal systems, but our knowledge of how much and why that role differs across communities of benthic producers is poor. There is another reason that justifies the importance of understanding the extent and controls of herbivory across communities of benthic producers: the change in dominant benthic producers that often results from increased anthropogenic eutrophication. Coastal eutrophication is one of the most wide-spread, pervasive human-induced environmental disturbances (Nixon 1995, Valiela 1995, Cloern 2001). It results from changes in land use that accompanies the occupation of coastal areas by human populations. Urbanization and deforestation into agricultural and farm-lands of the coastal watershed leads to higher rates of nutrient input into receiving coastal waters. In turn, if seagrasses are the dominant producers in the receiving waters, increased nutrient loading rates often lead to seagrass decline and 153 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 153-185. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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the proliferation of loose and/or epiphytic macroalgae. The proximate cause for that shift seems to be that macroalgae can generally take up water-column nutrients much faster than seagrass can (Sand-Jensen & Borum 1991, Duarte 1995, Pedersen & Borum 1997); thereby quickly building large stocks of biomass that can impose severe shading and adverse biochemical conditions on seagrasses and lead to their decline. There are numerous examples of the replacement of seagrasses by macroalgae as dominant benthic producers following increased eutrophication (e.g. Sfriso 1987, Reise et al. 1989, Rosenberg et al. 1990, Ciszewski et al. 1992, Bombelli & Lenzi 1996, Niell et al. 1996, Nienhuis 1996, Hauxwell et al. 2003), and that process is analysed at length in Chapter 3. Increased eutrophication also promotes the accumulation of phytoplankton in the water-column, which can exert substantial shading on benthic producers and further contribute to seagrass decline (Valiela et al. 1992, 2000b, Taylor et al. 1995a, b, 1999). In addition, some recent evidence suggests that large phytoplankton accumulations under high eutrophication may also limit the growth of benthic macroalgae through intense light attenuation, and even cause macroalgal decline (Funen Island Council 1991, Taylor et al. 1995a, b, 1999, Duarte 1995, Schramm 1996). Thus, while it is well known that the eutrophication of pristine environments often leads to the replacement of seagrasses by macroalgae as dominant benthic producers, it seems that further eutrophication reaching intense levels may lead to macroalgal decline and the replacement by sediment flats dominated by benthic microalgae. Regardless of what outcomes may occur under intense eutrophication, it is clear that increasing loading rates affect the composition of assemblages of benthic producers and may shift the dominant types. Hence, elucidating the extent and controls of the variability in herbivory among different types of benthic producers is important to understand the changes in herbivory that follow the replacement of dominant benthic producers under increasing eutrophication and the consequences for coastal systems. This chapter presents a summary of the extent and some of the controls and consequences of herbivory on benthic primary producers under pristine conditions (i.e. not affected by anthropogenic eutrophication). It also analyses how increasing eutrophication, through the promotion of benthic algal growth and associated seagrass decline, may affect the intensity of herbivory on benthic communities. The chapter is thus divided in two parts. In the first part, using an extensive compilation of the literature, I examine the variability in herbivory within and across pristine communities of microphytobenthos, macroalgae and seagrasses. I also discuss some of the controls and consequences of herbivory on benthic producers under pristine conditions. In the second part, I first comment on several considerations as to how the replacement of seagrasses by algae as dominant benthic producers following increased eutrophication may affect herbivory. I then summarize what several experiments comparing the response of macrofaunal densities to macroalgal blooms following enhanced eutrophication have taught us. I finish by discussing the changes in herbivory on benthic producers observed in two well-known ecosystems that have undergone intense eutrophication over the past few decades, Waquoit Bay (Massachusetts, USA) and the Venice Lagoon (Italy). The chapter focuses on temperate systems, but reference to tropical systems is also made when pertinent and possible. In choosing the tone of the chapter, I have intended to (1) review existing information relevant to the goals of the paper, (2) present hypotheses regarding the

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control of herbivory in pristine communities and how eutrophication may affect herbivory, and (3) suggest some recommendations for future research. 2. GRAZING ON MARINE BENTHIC PRODUCERS UNDER PRISTINE CONDITIONS: EXTENT, CONTROL AND CONSEQUENCES The question of whether herbivores are important controls of the biomass of benthic producers (i.e. top-down control) has received considerable attention over the past decades (Neckles et al. 1994, Valentine & Heck 1999, Heck et al. 2000, Worm et al. 2000, Lotze & Worm 2002). As a consequence, a wealth of reports quantifying herbivory on communities of benthic microalgae, macroalgae, and seagrasses are now available (for extensive reviews see Cebrian et al. 1998, Cebrian 1999, 2002). Those reports tell us a clear message: the extent of herbivory on marine benthic producers is highly variable, ranging from negligible levels to the removal of almost all available producer biomass. For instance, meio- (<500µm) and macroherbivores (>500µm) may remove a large percentage of the production of microphytobenthic communities (Nicotri 1977, Baird & Ulanowicz 1993), but in other instances herbivory only represents a minor loss for the community (Admiraal et al. 1983, Montagna 1984). The extent of herbivory may even be more variable in communities of marine macroalgae. Tropical communities often support intense consumption by herbivores and, as a consequence, they are mainly composed of turf and/or unpalatable (i.e. with chemical defences against herbivores) species. In fact, tropical macroalgal communities are the template for most seminal examples of herbivore control of the primary production and composition of marine communities (Samarco et al. 1977, Hay 1983, Hay 1984, Hatcher & Larkum 1983, Morrison 1988). In contrast, it seems that herbivores, by removing a smaller percentage of primary production, generally play a lesser role on the dynamics of temperate macroalgal communities (Newell et al. 1982, Alongi 1998). However, herbivory may also be intense in some temperate macroalgal communities, such as kelp beds, where large grazing events leading to the denudation of extensive kelp areas following the outburst of sea urchin populations have been often documented (Lawrence 1975, Breen & Mann 1976, Dayton 1985, Hagen 1995). Existing reports show a similarly broad range in herbivory on seagrasses. Although many authors have found that herbivory accounts for a small loss of seagrass production (Thayer et. al 1984, Mukai & Nojima 1985, Nienhuis & Groenendijk 1986, Cebrian et al. 1996, Cebrian & Duarte 1998), herbivores may also remove a large percentage of the production of tropical and temperate seagrass communities (Greenway 1976, Keller 1983, Vermaat & Verhagen 1996, Valentine & Heck 1999, Kirsch et al. 2002). The wide range in herbivory found for communities of benthic microalgae, macroalgae and seagrasses imply an important corollary; the role of herbivores as controls of producer biomass dynamics and nutrient recycling, which largely depends on the percentage of production removed, may differ greatly for each community type. A logical question arises then: are there any patterns in that variability within and across community types? In other words, does the extent of herbivory, and thus the importance of herbivores as top-down controls of producer biomass and nutrient recycling, vary in a certain way within and across community types? And if it does, what are the factors responsible for those patterns? Answering these questions requires extensive comparisons of representative ranges and frequencies of herbivory

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values reported in the literature. Unfortunately, this kind of comparisons has been scarce. Here, I capitalize on previous literature compilations (Cebrian et al. 1998, Cebrian 1999, Cebrian 2002) and add recently published data to search for patterns in the extent and control of herbivory within and between communities of microphytobenthos, macroalgae and seagrasses. The values compiled are representative of the community considered (i.e. they integrate the most abundant producers and herbivores in the community), correspond to natural conditions (i.e. communities not deliberately impacted by human disturbances such as eutrophication), and cover at least one year of observations or the growing season for annual producers. The data set is available at http://ecosystemslab.disl.org under "data sets". The compilation done confirms previous results that herbivory on benthic producers may be very variable. Indeed, herbivores may remove small or large percentages of primary production in communities of microphytobenthos (Fig. 1A), macroalgae (Fig. 1B) and seagrasses (Fig. 1C). The results, however, reveal an obvious difference between seagrasses and the two types of algal producers; the percentage of production removed is strongly skewed to the right for seagrasses, whereas this is not the case for microphytobenthos and macroalgae. Indeed most (>80%) of the seagrass communities compiled lose a relatively modest percentage (<10%) of production to herbivores (Fig. 1C). Thus, in spite of the large variability found within each community type, communities of microphytobenthos and macroalgae tend to lose higher percentages of primary production to herbivores than do seagrass communities (non-parametric multiple comparison Q test; microphytobenthos vs. macroalgae, P>0.10; microphytobenthos vs. seagrasses, P<0.01; macroalgae vs. seagrasses, P<0.01). This difference indicates that in general herbivores, by removing a larger percentage of primary production, play a greater role in the dynamics of producer biomass and nutrient recycling in communities of microphytobenthos and macroalgae than in seagrass communities. Accordingly, reports of significant top-down effects by herbivores on benthic micro- and macro-algae are abundant (e.g. Littler et al. 1983, Hackney et al. 1989, Steneck & Dethier 1994, Bennett et al. 1999, Blanchard et al. 2001), but much less so for seagrasses (Tribble 1981, Valentine & Heck 1991, Heck & Valentine 1995). As it is expected from the large variability found in the percentage of production consumed, absolute herbivory also varies greatly within communities of microphytobenthos (Fig. 1D), macroalgae (Fig. 1E) and seagrasses (Fig. 1F). For each community type, absolute consumption by herbivores ranges over three orders of magnitude (i.e. from <1 to ca. 1000 g C m-2 yr-1). However, and unlike the percentage of production consumed, absolute consumption is only marginally smaller in seagrass communities than in communities of microphytobenthos and macroalgae (Kruskal-Wallis, P=0.06). In other words, in spite of the fact that seagrass communities generally have smaller percentages of production removed by herbivores than do communities of microphytobenthos and macroalgae, the absolute flux of producer biomass channelled to herbivores varies little among the three community types. The explanation for these results lies in the differences in primary production found between the three communities. Seagrass communities tend to reach larger levels of primary production than many benthic micro- and macroalgal communities (Cebrian 2001). Higher levels of primary production in seagrass communities partially compensate for lower percentages of production removed by

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herbivores and, as a consequence, absolute consumption is only marginally smaller in seagrass communities than in communities of benthic micro- and macroalgae. These patterns suggest an important corollary; among communities of benthic producers, production of herbivore biomass (as indicated by the absolute flux of producer biomass transferred to the herbivore compartment) should not be as variable as the extent of herbivore control of producer biomass and nutrient recycling (as indicated by the percentage of primary production removed by herbivores). Many factors can contribute to the large variability found in the extent of herbivory on benthic producers. An extensive treatment of those factors is beyond the scope of this chapter because the aim is to summarize information that can be instrumental in predicting and understanding the changes in herbivory associated with the replacement of dominant benthic producers due to eutrophication. Here I comment on some factors that seem most relevant to that aim. One prominent factor that seems responsible for large differences in herbivory is the ample temporal and spatial variability that herbivore abundance often exhibits. Differences in recruitment success (Valiela 1995, Mann 2000), predator abundance (Vadas & Steneck 1995, Sala et al. 1998), degree of shelter in the habitat (Himmelman 1984, van Tamelen 1996) and frequency and duration of anoxic conditions (Sagasti et al. 2001, Gray 2002) may lead to large temporal and spatial differences in the abundance of herbivores in benthic communities. In turn, fluctuating herbivore densities may entail substantial variability in the extent of producer biomass consumed. Indeed, higher herbivore abundances have been experimentally associated with larger levels of consumption in communities of microphytobenthos (Admiraal et al. 1983, Blanchard et al. 2001), macroalgae (Hauxwell et al. 1998, Ruesink 2000) and seagrasses (Valentine & Heck 1991, Macia 2000). One of the best-known examples of control of herbivore abundance by predation and resulting effects on producers is the interaction between kelps, sea urchins (Strongylocentrotus sp.) and sea otters (Enhydra lutris) reported in the North American Pacific Coast (Estes & Palmisano 1974, Foster & Schiel 1988). When sea urchins are abundant, they remove large amounts of kelp biomass and may even denude extensive beds. However, under intense predation by sea otters, sea urchin abundance remains low and kelps flourish. Similar trophic cascade effects where the extent of herbivory is controlled by the intensity of predation on the herbivore have also been reported in communities of microphytobenthos (Posey et al. 1995, 1999) and seagrasses (Macintyre et al. 1987, Jackson 1997, Valentine & Heck 1999). Differences in the type of dominant herbivores in the community may also be conducive to substantial variability in the intensity of herbivory among marine benthic communities. For instance, in the seagrass communities where they are still present, large vertebrate herbivores such as dugongs, manatees, green turtles and waterfowl remove large quantities of seagrass and epiphyte biomass needed to support the high metabolical demands imposed by their large body mass (Ogden et al. 1983, Lanyon et al. 1989, de Iongh et al. 1995, Masini et al. 2001).

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Figure 1. The histograms of the percentage of primary production consumed by herbivores (plots A-C) and absolute consumption by herbivores (plots D-F) in communities of microphytobenthos, macroalgae and seagrasses. Dashed lines represent medians.

Similarly, green turtles and herbivorous fish, such as parrotfishes and surgeonfishes, often consume large quantities of algal biomass in tropical beds and coral reef communities (Hixon & Brostroff 1996, Musick & Limpus 1997). Consumption rates of producer biomass per individual are far greater for large vertebrate herbivores than for smaller invertebrate herbivores, and thus communities with different ratios of vertebrate vs. invertebrate herbivore abundance should also differ in the intensity of consumption of producer biomass. The feeding specificity of the dominant

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herbivores in the community may also contribute to the large variability in herbivory observed among marine benthic communities. Some herbivores are general feeders and consume a wide range of producer types, whereas some other herbivores are more specific and strongly select a limited number of producers and/or target certain parts of the producers. Sea urchins are often an example of extremely generalist herbivores in seagrass meadows. They may consume seagrass blades and sheaths, epiphytic micro- and macro-algae on the seagrass blades and rhizomes, and the macroalgae that grow on bare patches interspersed in the meadow (Bergin 1987, Klinger et al. 1994, Valentine & Heck 2001). On the contrary, dugongs select a number of seagrass species (De Iongh et al. 1995, Preen 1995) and green turtles target the basal part of turtlegrass (Thalassia testudinum) leaf blades (i.e. younger, more tender parts, Thayer et al. 1984, Zieman et al. 1984). Similar examples exist for other benthic communities. In sediment flats, harpacticoid copepods, nematods, ostracods and other important groups of herbivores often differ greatly in their choice and use of benthic microalgae as a food resource (Carman & Thistle 1985, Carman & Fry 2002). In temperate macroalgal communities, sea urchins often consume a variety of host and epiphyte species (Bergin 1987, Verlaque 1987) whereas other invertebrate grazers, such as isopods and amphipods, tend to be more selective (Salemaa 1987, Pavia et al. 1999, Karez et al. 2000). Therefore, it can be expected that differences in the relative abundance of generalist vs. selective herbivores and in the availability of preferred producer diets may also induce substantial variability in the intensity of herbivory among marine benthic communities. Another factor that may also be responsible for differences in herbivory among benthic communities is the nutrient content of the dominant producers. There is now extensive experimental evidence that the growth rates of many herbivores are limited by the nutrient content of their diets (Sterner & Hessen 1994, Elser et al. 2000a, Sterner & Elser 2002, Stelzer & Lamberti 2002, Urabe et al. 2002). Many lab and field manipulations have shown that aquatic and terrestrial herbivores that feed on producers with enhanced nitrogen and phosphorus contents grow faster. The rationale for these observations is that the growth rate of organisms is highly dependent on the concentrations of RNA and enzymes in the cells (Elser et al. 1996, 2000b, Stener & Elser 2002). Hence, because RNA is rich in phosphorus and enzymes are rich in nitrogen, higher nitrogen and phosphorus contents in the diet allow for a larger production of RNA and enzymes and, thus, faster growth rates. In turn, faster herbivore growth rates imply higher rates of biomass build-up and, thus, they should be conducive to larger rates of consumption of producer carbon. In fact, recent comparisons encompassing a wide range of community types, from marine to freshwater to terrestrial, have shown that, at that large across-system scale, communities composed of producers with higher nutrient contents support larger rates of consumption by herbivores (Cebrian et al. 1998, Cebrian 1999). Nevertheless, when the range of comparison is reduced to marine benthic communities, the evidence for the association between higher producer nutrient contents and faster consumption rates by herbivores seems equivocal. Most tests of this association in marine benthic communities refer to seagrass communities because the low nutrient contents of seagrass leaves relative to other benthic producers (Duarte 1992, Nielsen et al. 1996) and the high fiber content in the leaves (Bjorndal 1980, Dawes & Lawrence 1980) have made many authors hypothesize that these characteristics limit the intensity of herbivory on seagrasses. That hypothesis has

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been confirmed for some vertebrate herbivores, such as the parrotfish, dugongs and green sea turtles that select species or blade parts with higher nitrogen contents (Bjorndal 1980, Zieman et al. 1984, Williams 1988, McGlathery 1995, Preen 1995). However, recent papers comparing different seagrass species and herbivore populations have found no association between leaf nitrogen content and herbivory intensity (Cebrian & Duarte 1998, Mariani & Alcoverro 1999). Even more discrepant is the fact that certain herbivores, such as the pink sea urchin (Lytechinus variegatus), seem able to compensate for low nutritional quality in seagrass leaves and consume larger quantities of the leaves with lower nitrogen contents (Valentine & Heck 2001). Observations of the nutrient contents and consumption rates of benthic micro-and macroalgae also indicate controversial results. For instance, in coral reefs and other tropical algal communities most herbivores feed intensively on small turf algae that normally have higher nutrient contents than most other algae (Hay 1984, Russ & Alcala 1989, McClanahan et al. 1996). In contrast some macroalgae that also have high nutrient contents suffer little or no herbivory due to the presence of deterrent compounds (Hay & Fenical 1996, Hay 1997) or because of restricted accessibility by herbivores (Carpenter 1986, Alongi 1998, Mann 2000). To shed some light on whether a significant relationship between producer nutrient contents and herbivory intensity exists for marine benthic communities, I have compiled literature values of producer nitrogen and phosphorus contents and consumption by herbivores in communities of microphytobenthos, macroalgae and seagrasses. These values meet the same criteria as the values of absolute consumption and percentage of primary production consumed compiled for Figure 1 (i.e. representative of the community considered, referring to natural conditions, and encompassing one year of observations or the growing season for annual producers), and are also available at http://ecosystemslab.disl.org under "data sets". The results show that, when the three types of community are compared, larger percentages of primary production consumed by herbivores tend to be associated with higher nitrogen (Fig. 2A, Pearson correlation coefficient=0.40, P<0.01) and phosphorus contents (Fig. 2B, Pearson correlation coefficient=0.49, P<0.01), with seagrasses showing lower percentages consumed and nutrient contents in relation to microphytobenthos and macroalgae. However, the association does not hold within any of the community types; when communities of microphytobenthos, macroalgae and seagrasses are examined separately, the percentage of production consumed is independent of the nitrogen and phosphorus contents in the producer (P>0.10 for all within-community Pearson correlation coefficients). Neither are there any significant correlations between absolute herbivory and producer nitrogen and phosphorus contents either across or within community types (Figs. 2C and D, P>0.10 for all Pearson correlation coefficients). Taken together, these results convey a clear message; the role of producer nutrient content as a control of herbivory in benthic communities seems, at the best, modest. Significant, albeit not strong, associations are found between the percentage consumed and producer nutrient contents, but not with absolute herbivory. In addition, those associations only occur when different types of benthic community are compared and a wide range of percentage (from <0.1 to ca. 100%) and nitrogen (from 1 to 10 %DW) and phosphorus (from 0.1 to 1 %DW) values are encompassed. At scales smaller than comparing little consumed, nutrientpoor seagrass species with highly consumed, nutrient-rich macroalgae and

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microphytobenthos, it seems that factors other than producer nutrient contents are responsible for the variability in herbivory.

Figure 2. The relationships between (A) the percentage of primary production consumed by herbivores and nitrogen producer content, (B) the percentage of primary production consumed by herbivores and phosphorus producer content, (C) absolute consumption by herbivores and nitrogen producer content and (D) absolute consumption by herbivores and phosphorus producer content in communities of microphytobenthos (open squares), macroalgae (open circles) and seagrasses (filled circles). Data have been log-transformed to comply with the assumptions of correlation analyses (see text for statistical results).

Despite all the many factors that can affect herbivory in marine benthic communities, the absolute consumption of producer biomass remains closely associated with the magnitude of primary production both across and within community types (Fig. 3A). The compilation of published values shows high correlation coefficients across the three types of community (Pearson correlation coefficient=0.81, P<0.01) and within communities of microphytobenthos (Pearson correlation coefficient=0.96, P<0.01), macroalgae (Pearson correlation coefficient=0.94, P<0.01) and seagrasses (Pearson correlation coefficient=0.59, P<0.01). Thus, primary production appears as a good predictor of the absolute flux of producer biomass transferred to herbivores in marine benthic communities. The reason is entirely mathematical. Both across and within types of community, the percentage of production consumed varies little in comparison with the variability in primary production (Fig. 3B). For instance, across

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community types the percentage consumed varies from 10 to ca. 100% for production values comprised between 0.1 and 100 g C m-2 yr-1, and from 1 to ca. 100% for production values between 100 and ca. 1000 g C m-2 yr-1. Thus, whereas primary production ranges about 4 orders of magnitude across community types, the percentage consumed only ranges about 2 orders of magnitude. Much larger ranges in primary production than in the percentage consumed are also found within communities of microphytobenthos, macroalgae and seagrasses. As a consequence, absolute consumption, which corresponds to the product of primary production and the percentage consumed, remains closely associated with primary production across and within the three community types compared. Thus, regardless of the type of marine benthic community compared, it seems that more productive communities transfer a larger flux of producer biomass to herbivores and, thus, they should support higher levels of herbivore production. Another corollary is that, since absolute consumption is strongly associated with primary production but independent of producer nutrient content, primary production and producer nutrient content should also be independent in marine benthic communities. This is supported by the data compiled (P>0.05 for all Pearson correlation coefficients between primary production and nitrogen and phosphorus contents across and within community types) and by other past reports (Cebrian et al. 1998, Cebrian 1999).

3. TOWARDS AN UNDERSTANDING OF HOW EUTROPHICATION-INDUCED CHANGES IN BENTHIC PRODUCERS ASSEMBLAGES MAY AFFECT HERBIVORY As mentioned above, there are now numerous observations that increasing eutrophication of pristine shallow environments dominated by seagrasses often leads to their replacement by loose and epiphytic macroalgae (e.g. Valiela et al. 1992, 1997a, 2000b, Short et al. 1995, Niell et al. 1996, Nienhuis 1996, Raffaelli et al. 1998, Hauxwell et al. 2001, 2003). These macroalgae are frequently species of opportunistic green filamentous genera such as Enteromorpha, Chaetomorpha, Ulva and Cladophora (Schramm & Booth 1981, Lapointe & O'Connell 1989, Jeffrey et al. 1992, McComb & Humphries 1992, Jeffrey 1993, Dion & Bosec 1996, Romero et al. 1996), although some other brown and red genera, such as Porphyra, Gracilaria, Polysiphonia, Pilayella, and Ectocarpus (Fletcher 1996a, Peckol & Rivers 1995) may also respond quickly to enhanced nutrient availability. In some instances, however, factors such as unfavourable temperatures (Jeffrey et al. 1992, Nienhuis 1996), low light irradiance (Taylor et al. 1995a, b, 1999), exposure to wind and wave action (De Vries et al. 1996, Raffaelli et al. 1998) and elevated grazing (Neckles et al. 1994, Heck et al. 2000, Worm et al. 2000) may impair, and even abort, the formation of macroalgal blooms under enhanced nutrient delivery. The regulation of the appearance and duration of macroalgal blooms is not well understood. It seems that increased nutrient availability is a necessary condition, but not sufficient, for the build-up of macroalgal biomass (Raffaelli et al. 1998).

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Figure 3. The relationships between (A) absolute consumption by herbivores and primary production and (B) the percentage of primary production consumed by herbivores and primary production in communities of microphytobenthos (open squares), macroalgae (open circles) and seagrasses (filled circles). Data have been log-transformed to comply with the assumptions of correlation analyses (see text for statistical results).

Indeed, several authors have shown that blooms only occur when favourable conditions of potentially limiting factors accompany enhanced nutrient availability (Aubert 1990, DeVries et al. 1996, Fletcher 1996b, Poole & Raven 1997). The nature and controls of macroalgal blooms are examined in Chapter 3 and also in adequate recent reviews (Valiela et al. 1997a, Rafaelli et al. 1998, Cloern 2001). Here, I focus on the effects that those blooms and subsequent seagrass decline may have on the intensity of benthic herbivory. To do that, I first discuss some potential implications of the replacement of seagrasses by macroalgae on local herbivore populations and expected consequences on the intensity of herbivory. I then compare those expectations with the results of several case studies and search for trends in the response of herbivores to algal build-up and seagrass decline following eutrophication. Results from various mesocosm experiments also suggest that intense shading imposed by large phytoplankton accumulations under high eutrophication may limit macroalgal blooms, and even lead to their replacement by sediments dominated by benthic microalgae (Taylor et al. 1995a, b, Taylor et al. 1999). These results obtained in laboratory mesocosms are yet to be demonstrated in field conditions but, here, I also discuss how the replacement of macroalgal canopies by bare sea beds inhabited by microphytobenthos could affect benthic herbivory.

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3.1 Potential Effects of Shifting Assemblages of Benthic Producers on Herbivores 3.1.1 Changes in Producer Nutrient Content Seagrass leaves tend to have low nutrient contents in relation to macroalgal tissues. This was first showed in two extensive comparisons done in the last two decades (Atkinson & Smith 1983, Duarte 1992) and I now confirm it with newly compiled data (available at http://ecosystemslab.disl.org under data sets). As mentioned above, the data compiled refer to the community level (i.e. include the most important producers in the community), correspond to natural communities (i.e. not deliberately altered by human impacts such as eutrophication) and integrate several seasons of observations. The compilation shows that, in spite of substantial variability within each producer type, seagrass leaves tend to have lower nitrogen (Figs. 4B and C) and phosphorus (Figs. 4E and F) contents than macroalgae (non-parametric Q multiple comparison test, P<0.05 for both tests). Hence, it may be expected that a replacement of seagrass by macroalgae as dominant producers under increasing eutrophication enhances the nitrogen and phosphorus contents of the producer community. In addition, under high nitrogen availability bloom-forming macroalgal genera may increase their nitrogen contents to levels that stand high in relation to most other macroalgae (e.g. > 4%DW for Polysiphonia, Ulva and Cladophora, Peckol et al. 1994, Campbell 2001; >3%DW for Enteromorpha, Barr & Rees 2003), which further supports the hypothesis of an overall increase in producer nitrogen content as a result of macroalgal pile-ups and seagrass decline. Higher nutrient availability can also increase the nutrient content of seagrass leaves (Fourqurean et al. 1997, van Katwijk et al. 1999), so declining but richer seagrass leaves under increasing eutrophication could also contribute to the higher overall nutrient contents of the producer assemblage. If the replacement of seagrasses by bloom-forming macroalgae does indeed lead to higher nutrient contents in the producer assemblage, and based on the results of Figures 2A and B, it could then be expected that herbivores exert a greater pressure on the newly formed macroalgal canopy by removing a larger percentage of primary production. No significant changes, however, would necessarily be expected in absolute consumption since a concomitant decrease in total primary production could offset the increase in the percentage consumed (Figs. 2C and D; Borum & SandJensen 1996, Cebrian 1999). At any rate, the association between larger percentages of production removed and higher producer nutrient contents found across producer types (Figs 2 A and B) is not strong and it is possible that other concomitant deleterious effects of macroalgal build-ups on herbivory override the expected positive effect of enhanced producer nutrient contents. Some of those possible negative effects are discussed below. My compilation also shows that microphytobenthos tends to have higher nitrogen (Figs. 4A and B) and phosphorus (Figs. 4D and E) cell contents than do macroalgal tissues (non-parametric Q multiple comparison test, P<0.05 for both tests). Values of nitrogen and phosphorus contents in microphytobenthic cells are scarce, but the result is robust in spite of having few values available for the comparison. Hence, and based again on the association between larger percentages of production removed and

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higher nutrient contents found across different types of benthic producers (Figs. 2A and B), it may seem that a hypothetical replacement of macroalgal canopies by sediment microalgae due to intense shading by phytoplankton that would occur under intense eutrophication could also be conducive to higher intensities of herbivory. This logic, however, is fallacious, since the association between larger percentages consumed and higher producer nutrient contents does not hold when only macroalgae and microphytobenthos are compared (P>0.05 for both correlation coefficients). A replacement of macroalgae by microphytobenthos under intense eutrophication undoubtedly seems to lead to enhanced nutrient contents in the producer community, but we do not have a strong empirical basis to predict how that enhancement can affect herbivory. 3.1.2 Changes in Structure A wealth of reports have demonstrated that the structurally complex web of vertical shoots, leaves and attached epiphytes in seagrass meadows offers numerous refuge possibilities to many organisms (e.g. Orth 1977, Heck & Orth 1980, Heck & Crowder 1991, Sogard, 1992, Edgard & Shaw 1995). Indeed, seagrass meadows are considered as one of the most important habitats in coastal ecosystems (Orth et al. 1984, Hemminga & Duarte 2000, Williams & Heck 2001). Some organisms remain in the meadows throughout their entire life span (i.e. permanent residents), whereas some others use the meadow only during certain stages of their life cycles (i.e. temporal residents). Notorious among those temporal residents are the juveniles of many fish species with commercial and recreational interest. Upon recruitment from offshore waters, the juveniles find shelter and abundant food in the meadows and, after reaching a certain size, swim to deeper waters (Coles et al. 1993, Ros & Moser 1995, Rooker et al. 1998). Seagrass meadows thus play an important role as "nurseries" for numerous fish species (Heck et al. 2003). The dense macroalgal canopies that are formed under increasing eutrophication are also structurally complex and can thus offer shelter to many organisms. Indeed, several studies have shown that increasing algal abundance reduces predation rates on mobile epifauna. For instance, Isaksson et al. (1994) and Pihl et al. (1995) showed that increasing cover by Enteromorpha and Cladophora depressed the predation rates of the crustaceans Crangon crangon and Carcinus maenus by cod (Gadus morhua). Bluecrab (Callinectes sapidus) may also find refuge from predation in mats formed by the macroalgae Ulva lactuca (Wilson et al. 1990). More recently, Norkko (1998) examined how algal blooms may influence the trophic relationships between infaunal preys and epibenthic predators and provided experimental evidence that blooms of Ectocarpus and Pilayella offer substantial refuge to the infaunal amphipod Corophium volutator and polychaete Nereis diversicolor from the epibenthic predator Crangon crangon. Dense macroalgal blooms may also be valuable habitats for a number of fish species, such as the four-spined stickleback (Apeltes quadracus, Raposa & Oviatt 200) and tautog (Tautoga onitis, Dorf & Powell 1997).

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Figure 4. The histograms of nitrogen (plots A-C) and phosphorus (plots D-F) contents in microphytobenthos, macroalgae and seagrasses. Dashed lines represent medians.

Overall, thus, it is not straightforward how increasing algal abundance and declining seagrass abundance following eutrophication may affect the degree of refuge offered by the producer assemblage. Based on the arguments presented above, it may seem that the structural complexity gained with the building of dense macroalgal blooms can compensate for the loss due to declining seagrass, with the overall level of refuge offered by the producer community thus being little altered. Unfortunately, empirical tests of this hypothesis are lacking. Common methods to measure structural complexity in seagrass and macroalgal canopies have not been developed. Fractal geometry may prove as a good solution (Sugihara & May 2000, Chaplin 2001), but much research is needed to decide how the technique can be commonly applied to seagrass and macroalgal canopies. Elucidating how the replacement of seagrasses by macroalgae under increasing eutrophication affects the degree of refuge provided by the producer assemblage to organisms is an important question to understand how eutrophication, one of the most pervasive world-wide human impacts, may affect the dynamics of coastal ecosystems. Future research would do well in addressing that question. Under further eutrophication, the potential loss of macroalgal canopies due to intense shading by phytoplankton and subsequent replacement by bare bottom

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populated by microphytobenthos would certainly decrease drastically the extent of refuge offered to organisms. It may be expected that the faunal population would then shift from being dominated by epibenthic species to being dominated by infaunal species that live and hide in the sediment. This hypothesis is in accordance with recent reports of the response of faunal populations to the experimental removal of macroalgal canopies (Franz & Friedman 2002). 3.1.3 Changes in Biochemical Conditions Perhaps one of the best documented processes that often results from the accumulation of large macroalgal blooms is the occurrence of anoxic conditions within the algal canopy. In dense algal canopies, photosynthesis is usually restricted to the top layers of the canopy due to the intense self-shading. This was first shown by Peckol and Rivers (1996), who found that, in thick canopies of Cladophora vagabunda and Gracilaria tikvahiae in Waquoit Bay (Massachusetts, USA), 90% and 70% of incident light was intercepted in the first 2cm of the canopy (Fig. 5A). Part of the oxygen produced in the top centimetres of the canopy diffuses downward and it is rapidly taken up by shaded, non-photosynthesizing algae. In addition, macroalgal canopies, in virtue of their intertwined texture, retain many of the detrital particles that settle from the water column (Hull 1987, Worcester 1995), which serve as substrata for bacterial growth and further consume the oxygen that diffuses from the top layers. As a consequence, oxygen concentrations are reduced to hypoxic/anoxic levels within the top centimetres of the canopy even on sunny days and high photosynthetic activity of the top layer (Fig. 5B). The problem is worsened at night, when anoxia occurs throughout most of the canopy. Anoxic conditions can even be found at dawn above the canopy up in the stratified water column after cloudy, hot summer days (D'Avanzo & Kremer 1994). Many lab experiments and field surveys have shown that macroalgal blooms may induce hypoxic and/or anoxic conditions within the canopy and overlying water column. The large macroalgal blooms observed in Venice Lagoon (Italy) are a classical example; the dynamics of macroalgal biomass in the Lagoon displays large variability among seasons and locations, but when a large biomass of Ulva accumulates, water-column anoxia may occur on hot summer days under cloudy conditions or high water-column turbidity (Sfriso et al. 1992). The large canopies of Ulva lactuca that accumulate in much of the shoreline of Jamaica Bay (New York, USA) in the summer also produce anoxic conditions within the canopy, and sometimes in the water column, during the night and early morning hours (Franz & Friedman 2002). In turn, hypoxic/anoxic conditions are extremely harmful for most organisms that inhabit the algal canopy. Some organisms have a higher tolerance to oxygen deprivation (Sagasti et al. 2001), but in general, most fauna readily migrates to the top canopy layer when oxygen becomes limiting or, if exposed to oxygen deprivation for a sustained period, they perish (Rafaelli et al. 1998, Tagliapietra et al. 1998, Osterling & Pihl 2001, Franz & Friedman 2002, Gray et al. 2002). Not all macroalgal blooms are conducive to anoxic conditions. In some cases, the blooms occur as sparse drifting algal masses that are relatively well flushed and exposed to continuous air exchange with the atmosphere (Rafaelli et al. 1998). Under these conditions, hypoxic/anoxic conditions are rarely encountered within the algal mass

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(Thybo-Christensen et al. 1993, Norkko & Bonsdorff 1996, Norkko et al. 2000). However, if the blooms occur as thick macroalgal canopies that remain stagnant on the bottom and subject to hydrographical isolation through stratification, anoxic/hypoxic conditions and subsequent detrimental effects for the resident fauna can be expected. Oxygen depletion is not the only harmful condition that organisms face within the canopies of algal blooms because the highly reduced environment that occurs with anoxia favours the formation of sulphides and ammonia (Gray et al. 2002) In fact, Nedergaard et al. (2002) reported high hydrogen sulphide concentrations within canopies of Ulva lactuca. There have been no direct measurements of ammonia concentrations within canopies of bloom-forming algae, but McGlathery et al. (1997) and Krause-Jensen et al. (1999) measured high ammonium concentrations within canopies of Chaetomorpha linum, and Hauxwell et al. (2001) measured high concentrations within mixed canopies of Cladophora vagabunda and Gracilaria tikvahiae (Fig. 5B). In turn, elevated ammonium concentrations should also increase the concentration of ammonia since the two products are in equilibrium in the marine environment (Gray et al. 2002). Sulphides and ammonia are extremely toxic to many marine organisms and can inflict substantial mortality even at low concentrations (Gray et al. 2002). In addition, in a highly reduced environment the effects of anoxia may interact with the effects of sulphide and ammonia and cause increased damage. For instance, hypoxic tolerance in three species of marine gastropods (Littorina ziczac, Neritina vorginea and Olivella vereuxii) decreased in the presence of hydrogen sulphide (Hiroki 1978). Therefore, it seems that steady, thick macroalgal canopies are, except in the upper layers, a hostile environment for many organisms.

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Figure 5. (A) Light attenuation through mats of Gracilaria tikvahiae (open squares) and Cladophora vagabunda (filled squares) in Waquoit Bay (Massachusetts, USA). Light attenuation is expressed as the percentage of irradiance measured at the mat surface that remains at a given mat depth. From Peckol & Rivers (1996). (B) Profiles of oxygen and ammonium concentrations throughout the water column and natural ("no algae" treatment) and experimentally thickened ("algae" treatment) algal mats in the subestuaries Sage Lot Pond and Hamblin Pond of the Waquoit Bay estuarine system . In Sage Lot Pond, natural and experimental algal mats were 2 and 25cm tall, respectively, and in Hamblind Pond they were 9 and 18 cm tall. The algal mat consisted of a mixture of G. tikvahiae and C. vagabunda. In each subestuary, dusk and dawn concentrations were taken at a number of depths (i.e. labelled as "distance from the sediment" in the figure) through the water column and canopy of two natural and two experimentally thickened mats. Average values and standard deviations are plotted (see legend in figure). Measurements were done after a typical sunny summer day. Dashed lines represent the height of thickened algal mats. Adapted from Hauxwell et al. (2001).

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3.2 Overall Effects of Eutrophication-Induced Changes in Benthic Producer Assemblages on Herbivory: Lessons from Case Studies Based on the arguments elaborated above, we can expect diverse and counter effects of eutrophication-driven macroalgal build-ups and seagrass decline on benthic herbivory. One direct positive effect seems to be an overall increase in the nutrient contents of the producer community. Anoxic conditions and elevated sulphide and ammonia concentrations within the algal canopy, however, may impose severe deleterious consequences on associated fauna and herbivory. It is not straightforward how the replacement of seagrass meadows by macroalgal mats may alter the structural complexity of the community and degree of refuge offered to organisms. Both types of producer community are structurally very complex and important habitats for many organisms. It thus seems that, when replacing seagrass meadows by macroalgal mats, the effects on associated fauna due to changes on structure would be modest in comparison with the effects of enhanced producer nutrient contents and adverse biochemical conditions. A logical question arises; what is the net effect of the replacement of seagrass meadows by macroalgal blooms on benthic herbivory? In other words, how do two of the most important consequences of that replacement, i.e. an increase in the nutrient content of the producer community but adverse biochemical conditions, interact and what are the overall implications on the extent of benthic herbivory? Unfortunately, very little work has been done to answer that question, even though elucidating how increased eutrophication, through changes in the dominant assemblages of benthic producers, may affect herbivory is important to understand associated changes in secondary production and nutrient recycling and storage in coastal ecosystems. Many authors have examined how macroalgal blooms on bare substratum affect faunal diversity and abundance (e.g. Rafaelli et al. 1998, Bolam et al. 2000, Norkko et al. 2000, Osterling & Pihl 2001, Cardoso et al. 2002, Franz & Friedman 2002), and a few authors have assessed the extent of herbivory on macroalgal blooms (Hauxwell et al. 1998, Sfriso & Marcomini 1997), but no comparison of how herbivory changes from seagrass-dominated to macroalgaldominated communities following increased eutrophication, and of the controls responsible for those changes, has been done to date. Here, I attempt to shed some light on that question by (1) examining a number of reports of the response of macrofauna to macroalgal blooms on bare sediment and the role of anoxia on that response and (2) analysing how herbivory on macroalgae has changed in two well known ecosystems, Waquoit Bay (Massachusetts, USA) and Venice Lagoon (Italy), where increased eutrophication has caused the persistence of dense macroalgal mats and intense seagrass loss. 3.2.1 Macrofaunal response to macroalgal blooms on sediment flats. A good number of authors have examined how macroalgal build-ups on intertidal and subtidal sediments affect infaunal and epifaunal communities (see review by Rafaelli et al. 1998). Those authors have focussed on such aspects as changes in abundance (Osterling & Pihl 2001, Franz & Friedman 2002), diversity (Bolam et al. 2000, Norkko et al. 2000), growth (Sogard 1992, Cardoso et al. 2002), recruitment (Bonsdorf et al. 1996, Cardoso et al. 2002), and trophic relationships (Sogard & Able

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1991, Lopes et al. 2000). No report, however, has addressed how macroalgal accumulations on bare substratum alter the intensity of benthic herbivory. Here, I discuss existing information on the development of algal blooms on bare substratum and effects on local fauna that can help predict what the consequences on herbivory could be. In what has become a classical review, Rafaelli et al. (1998) documented that macroalgal blooms that develop on sediment flats can have contrasting effects on the local fauna depending on the type of dominant fauna at the bloom onset, the size and nature (drifting or steady) of the bloom, season and, to a lesser extent, the type of bloom-forming algae. However, despite the high variability in the response of macrofauna to macroalgal blooms, their analyses pointed to some apparent trends. Macroalgal blooms, particularly those that remain anchored to the bottom, often reduced the abundance of infauna (i.e. organisms living and/or hiding in sediment burrows) and surface feeders, except for organisms that can thrive under highly reduced conditions such as the polychaete Capitella capitata. Macroalgal blooms, on the contrary, frequently increased the abundance of epifaunal species (i.e. organisms living above the sediment), which readily colonized and found abundant shelter and food in the new algal substratum. Overall, thus, one of the messages that stemmed from the review by Rafaelli et al. (1998) is that macroalgal blooms tend to be detrimental for most infaunal species but beneficial for many epifaunal species. Nevertheless, research done since the review by Rafaelli et al. (1998) was published has provided evidence that macroalgal blooms may also be detrimental for epifaunal species through the generation of persistent anoxia. For instance, Franz & Friedman (2002) did a series of algal removal/addition experiments to demonstrate that blooms of Ulva lactuca drastically reduced the abundance of epibenthic copepods and attributed that reduction to the anoxic conditions registered within the algal mats at night. Similarly, Tagliapietra et al. (1998) showed opposite cycles between benthic macrofaunal abundance, including many epibenthic species, and the abundance of Ulva in the Palude della Rosa lagoon (Italy), with precipitous decreases in faunal abundance following the formation of algal blooms. They also attributed those changes to the severe anoxic conditions brought about by the blooms. In contrast, Norkko et al. (2000) found high levels of epifaunal abundance in drifting algal mats of Ectocarpus siliculosus and Pilayella littoralis, which could even exceed the high levels typically found in seagrass meadows. Drifting algal mats are often highly mobile and are thus relatively well flushed in relation to large, steady algal canopies. Indeed, Norkko et al. (2000) reported intermediate oxygen levels within the drifting mats and only found hypoxic conditions at the algal-sediment interface. Thus, the lack of prolonged anoxia within the drifting mats, along with the enhanced levels of refuge, food and surface for recruitment provided by the mats, probably explain the high epifaunal abundance reported by the authors. Additional evidence of the importance of anoxia in determining the overall effect of macroalgal blooms on epifaunal populations is provided by Osterling & Pihl (2001). They subjected subtidal faunal communities to moderate (i.e. mesh-bags filled with algae simulating floating patches with loose attachment to the sediment) and intense (i.e. cages filled up with algae simulating steadier, thicker canopies) algal disturbance and found higher epifaunal abundances at moderate disturbance. They suggested that differences in hypoxic/anoxic conditions were probably an important factor explaining the results.

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Hence, when one compares reports of positive and negative effects of macroalgal blooms on epifaunal abundance, the occurrence of anoxia and probably other harmful chemical reagents such as ammonia and sulphides within the algal canopy seems to be of paramount importance in dictating the net consequences of the bloom; if anoxia persists, an overall negative effect on the abundance of the epifaunal community seems to prevail in spite of all the other possible concomitant positive effects of the bloom (i.e. enhanced refuge, food and structure for recruitment). A number of epifaunal species that thrive in algal canopies are herbivorous (Hicks & Coull 1983, Salemaa 1987, Geertz-Hansen et al. 1993). Many other species are primarily detritivorous but can also consume living tissues of the algae (Rafaelli et al. 1998, Norkko et al. 2000). Thus, it may be hypothesised that the extent of herbivory on macroalgal blooms is also primarily determined by the intensity and persistence of oxygen reduction; if persistent anoxic conditions occur, herbivory, via depressed herbivore abundance, would also be drastically reduced. In contrast, in comparatively well-oxygenated canopies, such as those of drifting mats, algal consumption by herbivores could be substantial. The two case studies discussed below represent a template to further explore this hypothesis.

3.2.2. Changes in herbivory on macroalgal mats in Waquoit Bay (Massachusetts, USA) and the Venice Lagoon (Italy). The estuarine system of Waquoit Bay comprises a number of subestuaries that have undergone contrasting eutrophication over the last decades (Valiela et al. 1997b, Valiela et al. 2000a). Some subestuaries are protected land and are thus surrounded by Spartina marshes and forestlands. Those subestuaries receive small nutrient loading rates from their watersheds. In contrast, some other subestuaries have been greatly deforested mainly through intense urbanization and the construction of golf courses. As a consequence, those subestuaries receive large nutrient loads from their watersheds. Finally, some other subestuaries have only undergone moderate urbanization and, consequently, they receive intermediate loads. Interestingly, the physical and hydrographical differences among the subestuaries of the Waquoit Bay system are small in comparison to the large variability in nitrogen loading rates, which range from 5 to 400 Kg nitrogen ha-1 yr-1. Therefore, the system represents a convenient scenario to test how increasing eutrophication may affect the ecology of coastal waters. One of the consequences of intense nitrogen loading has been the replacement of native eelgrass (Zostera marina) meadows by massive accumulations of the algae Cladophora vagabunda and Gracilaria tikvahiae (Valiela et al. 1992, Valiela et al. 2000b, Hauxwell et al. 2003). In the subestuaries subject to the highest loading, eelgrass has disappeared completely and the algae form dense, thick canopies that may reach a maximum height of ca. 80cm (Hauxwell et al. 2001). On the contrary, in the pristine subestuaries, eelgrass meadows are still luxurious and the algal canopy is sparse and only 2-3 cm tall on average. Hauxwell et al. (1998) investigated how larger algal canopies produced with increasing eutrophication affected the intensity of herbivory on the algae. To do so, they measured herbivory in three subestuaries subject to contrasting nitrogen loading rates and, thus, with

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different algal canopies. Sage Lot Pond was the pristine environment, with a loading rate of 5 Kg nitrogen ha-1 yr-1 and a maximum canopy height of 10 cm; Quashanet River had intermediate values of loading rate (300 Kg nitrogen ha-1 yr-1) and maximum canopy height (20 cm), and Childs River had the highest values (410 Kg nitrogen ha-1 yr-1 and 75 cm, respectively). The results obtained by Hauxwell et al. (1998) represent an exemplary case of how the positive and negative effects entailed by macroalgal build-ups may affect the intensity of herbivory on the algae. The seven-fold increase in maximum canopy height observed from Sage Lot Pond to Childs River undoubtedly entailed a substantial increase in refuge for many epifaunal species. Furthermore, as a consequence of higher nutrient availability, the two algal species increased their internal nutrient contents with higher nitrogen loading (Peckol et al. 1994), which should also enhance their nutritional quality for herbivores. Accordingly, Hauxwell et al. (1998) found that the individual consumption rates for most species of herbivorous amphipods examined increased from Sage Lot Pond to Quashanet River to Childs River (Fig. 4 in Hauxwell et al. (1998)). Nevertheless, they also found that the total abundance of herbivorous amphipods and isopods within the algal canopy decreased abruptly as nitrogen loading and height of the canopy increased, with the summer abundance at the pristine Sage Lot Pond being four-fold higher than that at the eutrophic Childs River. Those differences overrode the opposite trend found for individual consumption rates and, as a consequence, the total absolute consumption of algal biomass (in g DW per year per bottom m2) and the percentage of algal production consumed decreased from Sage Lot Pond to Quashanet River to Childs River. Later work by the same authors (Hauxwell et al. 2001, 2003) has shown that the anoxic conditions created at night within the thick algal canopies at Quashanet River and Child River, in contrast with the absence of anoxia in the sparse, thin canopies at Sage Lot Pond, are probably responsible to a great extent for the differences in herbivore abundance and intensity of herbivory found across the three subestuaries (Fig. 6). The Venice Lagoon (Italy) provides another example of the importance of anoxic conditions in regulating the extent of herbivory on eutrophication-driven macroalgal mats. The Venice Lagoon experienced substantial anthropogenic eutrophication throughout the 70's and 80's via untreated sewage effluents and inputs from agricultural fields in the watershed (Orio & Donazzolo 1987, Sfriso et al. 1992). This caused the replacement of beds of seagrasses (Zostera marina, Z. noltii and Cymodocea nodosa) and bulky, slow-growing red and brown macroalgae (Cystoseira fimbriata, C. barbata and Dictyopteris membranacea) by mats of green filamentous algal species (Ulva rigida, Enteromorpha sp. and Cladophora sp; Pignatti 1962, Sfriso 1987, Sfriso et al. 1989) in central parts of the Lagoon where the water is slowly renewed by offshore influx (Battiston et al. 1983). In turn, those changes in benthic producer assemblages have entailed profound consequences on the local fauna, which has also undergone substantial changes in diversity and abundance (Giordani Soika & Perin 1970, Tagliapietra et al. 1998). Sfriso et al. (1992) showed that one of the most important effects induced by macroalgal mats on the faunal populations in the central part of the Lagoon was the frequent occurrence of hypoxic/anoxic conditions. They sampled several stations in the central part of the Lagoon and found that, in the stations with the highest nutrient inputs and least exposed to flushing by offshore water, large mats of Ulva and

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Enteromorpha accumulated in the summer months and anoxic conditions occurred under reduced light intensity (i.e. cloudy days) and/or water-column stratification. They also found large mortality events of fish and benthic animals associated with those anoxia events. Sfriso et al. (1992) did not measure the extent of herbivory on the macroalgal mats during well-oxygenated and anoxic conditions, but it is obvious that herbivory would have been drastically reduced, if not completely eliminated, under anoxic conditions. Interestingly, though, a few years later Sfriso and Marcomini (1997) measured herbivory on the Ulva stands of one of the previously most polluted stations and where they had found the highest occurrence of anoxic events (Lido station). They also compared those herbivory values with those measured in a southern station dominated by eelgrass (Z. marina) that had undergone little eutrophication. Sfriso and Marcomini (1997) found intense herbivory on the Ulva mats (65% of the algae production consumed), but negligible values on the seagrass. What is very relevant to this discussion is that, in contrast to the frequent occurrence of anoxia recorded a few years earlier, they did not find any anoxic conditions within the algal mat at the Lido station. They attributed the lack of anoxic conditions to reduced levels of Ulva abundance in relation to the levels measured a few years earlier. In their conclusion, they mentioned that "in the absence of anoxic conditions, the benthic herbivores associated with Ulva drastically reduced the biomass produced by this species whereas, in areas dominated by rhizophytes, herbivores feed primarily on the macroalgae growing within the Zostera shoots" 4. CONCLUSIONS AND SUGGESTIONS FOR FUTURE RESEARCH The analyses presented here confirm past observations that herbivory on benthic producers can be very variable. However, the results reveal a strong tendency for seagrass communities to have a smaller percentage of primary production removed by herbivores in comparison with communities of microphytobenthos and macroalgae. These differences suggest that herbivores should generally play a lesser role in the dynamics of producer biomass and in the recycling of producer-bound nutrients in the former communities. The results also suggest that the lower nutrient contents found for seagrass leaves in relation to the contents in macroalgal tissues and microphytobenthic cells may be partially responsible for the smaller percentages consumed in seagrasses. Yet, the association between higher producer nutrient contents and larger percentages of production consumed found across types of benthic producer is not strong and other factors, such as herbivore abundance, size or feeding specificity, may be more important at explaining the variability in percentage consumed across benthic producers.

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Figure 6. Hypothesized causal relationships between increased eutrophication and the changes observed in algal herbivory in subestuaries of the Waquoit Bay system. (A) Higher nutrient loading rates are conducive to larger macroalgal mats (macroalgal biomass, mostly accounted for by C. vagabunda, is expressed in g dry weight per square meter; Sage Lot Pond is the "low N estuary", Quashanet River is the "mid N estuary" and Childs River is the 'high N estuary". Adapated from Hauxwell et al. 1998). (B) Larger algal canopies induce anoxic conditions within the canopy (see Fig. 5B for details; adapted from Hauxwell et al. (2001)). (C) Increased anoxia within the algal canopies leads to depressed grazer abundance (legends for estuaries as in Figure 6A; adapted from Hauxwell et al. (1998)). (D) depressed grazer abundance results into lower leves of herbivory on the algae, both expressed as absolute consumption (in g algal dry weight per bottom square meter per day, Y-axis on the plot) and as percentage of algal production consumed (the ratio between absolute consumption and algal production, which is also exprssed in g algal dry weight per bottom square meter per day and corresponds to the X-axis in the plot) (legends for estuaries as in Figure 6A; the 1:1 line has been depicted for comparison; adapted from Hauxwell et al. (1998))

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More research is needed to elucidate why seagrasses tend to lose a lower percentage of production to herbivores when compared with microphytobenthos and macroalgae. However, when herbivory is regarded as an absolute flux, seagrasses tend to show only marginally lower values than do the two types of algal producers because the higher values of primary production typically found for seagrasses partially compensate for the smaller percentages of production removed. Indeed, primary production stands out as a strong indicator of the variability in absolute herbivory when benthic producers of the same type or different types are compared; when benthic micro- and macroalgae and seagrasses are examined in concert or separately, larger absolute consumption is strongly associated with higher primary production. The reason is entirely mathematical; both with and across producer types, the range in primary production largely exceeds that in the percentage consumed and, as a consequence, absolute consumption remains closely associated with primary production. These results provide a basis to formulate hypotheses as to how the replacement of seagrass communities by mats of filamentous algae that usually follows anthropogenic eutrophication may affect the intensity of benthic herbivory. This is an important question to understand how this pervasive, world-wide environmental impact alters secondary production and carbon and nutrient cycling in marine ecosystems but, unfortunately, the question is yet to be investigated. My results suggest that the overall increase in producer nutrient content entailed by the replacement of seagrasses by filamentous algae as dominant producers may be conducive to higher percentages of primary production consumed by herbivores. In turn, that would promote the role of herbivores as controls of carbon and nutrient recycling and storage as producer biomass. Changes in absolute consumption and implications on secondary production would instead be determined by the balance between the changes in primary production and percentage consumed that would occur with the shift from seagrasses to macroalgae. Nevertheless, abundant evidence indicates that the adverse biochemical conditions that are usually found in eutrophication-driven thick macroalgal mats may be of paramount importance in determining the response of herbivory. Numerous reports of the effects of macroalgal blooms in sediment flats on associated fauna and the history of research in the highlyeutrophic ecosystems of Waquoit Bay (Massachusetts, USA) and Venice Lagoon (Italy) suggest that, if the macroalgal mats that outcompete seagrass beds under enhanced eutrophication induce persistent anoxia and elevated ammonia and sulphide concentrations within the mat, the extent of herbivory, via decreased faunal abundance, would be drastically limited to values even lower than those on the former seagrass community. Much less evidence exists to hypothesise how a possible replacement of macroalgal mats by microphytobenthos-dominated bare sediment due to severe phytoplankton shading under intense eutrophication may affect the levels of benthic herbivory. Profound changes in producer nutrient content, spatial structure, biochemical conditions and other factors affecting herbivory will follow the replacement of macroalgal mats by bare bottom inhabited by microphytobenthos. Those changes may have contrasting effects on local herbivore populations, and the possible overall impact on the intensity of benthic herbivory is not straightforward. As the replacement of macroalgal mats by sediment flats dominated by microphytobenthos seems plausible under substantial phytoplankton shading,

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6. AFFILIATION J. Cebrián: Dauphin Island Sea Laboratory, Department of Marine Science, University of Southern Alabama, 101 Bienville Boulevard, Dauphin Island, AL 36528, USA.

GARY THOMAS BANTA, MORTEN FOLDAGER PEDERSEN AND SØREN LAURENTIUS NIELSEN

DECOMPOSITION OF MARINE PRIMARY PRODUCERS: CONSEQUENCES FOR NUTRIENT RECYCLING AND RETENTION IN COASTAL ECOSYSTEMS

1. INTRODUCTION Coastal marine plant communities are, in contrast to the open sea where the only primary producers are phytoplankton, made up by a number of different plant groups that each contribute to total autotrophic biomass and production (Borum & SandJensen, 1996); Hauxwell and Valiela, Chapter 3). Coastal plant communities rank among the most productive biomes on earth and large amounts of nutrients (especially N and P) pass therefore through this compartment (Pedersen et al., chapter 1). Nutrients assimilated during plant growth are immobilized temporarily in living biomass and detritus until they are mineralized through grazing or decomposition. Incomplete or very slow decomposition of detrital components that are resistant to decomposition may, however, lead to long term or permanent (i.e. over decades, centuries) storage of organic matter and associated nutrients. Whether detritus-bound nutrients are mineralized and recycled quickly or whether they become stored in slowly degradable detritus over longer time scales is controlled by the rates and patterns of the decomposition of plant detritus. The percentage of net primary production that is lost through grazing versus decomposition differs among major groups of marine primary producers (i.e. phytoplankton, macroalgae and seagrasses) (Cebrián, 1999; Duarte, 1995; Duarte & Cebrián, 1996). Hence, slow-growing macroalgae and seagrasses seem to lose a much smaller proportion of net primary production to grazing than fast-growing macro- and microalgae and thus the immediate fate of organic matter produced by the slower growing macrophytes is therefore to end up as detritus. A large proportion of the primary production may therefore enter the detrital pathway when coastal ecosystems are dominated by macrophytes. The rate by which plant detritus decomposes depends on the chemical composition, nutrient content and carbon quality of the litter (i.e., degradability or “quality” of detritus) and therefore decomposition rates vary substantially among plant types ranging from uni-cells to trees (Enríquez, Duarte, & Sand-Jensen, 1993). Comparative decomposition studies of marine plants have also shown that detritus from phytoplankton and fast-growing macroalgae generally decay faster than that from slow-growing, perennial macroalgae and seagrasses (e.g., Buchsbaum, Valiela, Swain, Dzierzeski, & Allen, 1991; Schmidt, 1980; Twilley, Ejdung, Romare, & Kemp, 1986). These variations suggest that nutrients are recycled much faster in systems dominated by the former groups (Duarte, 1995). This conclusion is based on 187 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 187-216. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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the assumption that N and P are lost from the litter at the same rate as dry weight or carbon. It is however evident, from the rich literature on decomposition of terrestrial plant litter, that loss rates of N and P may differ substantially from those of dry weight or carbon (e.g., Staaf & Berg, 1982). For example, nutrients may be mineralized faster than organic matter are decomposed (i.e. net mineralization) or they may be lost more slowly than organic matter are decomposed due to nutrient assimilation by the associated microflora (i.e. net immobilization during decomposition). A proper evaluation of mineralization of detritus-bound nutrients through decomposition should therefore preferably be based upon observed changes in the N and P pools of the detritus and not on mass changes in biomass. Such data are unfortunately rare for marine plants. The purpose of this chapter is to compare patterns of decomposition and mineralization among different groups of marine primary producers in order to evaluate whether dominance of certain plant groups (i.e. fast-growing ephemeral versus slow-growing persistent types) affect the fate – that is, mineralization versus long term storage - of detritus-bound nutrients in coastal marine ecosystems. Based on a large data-set extracted from the literature, we will first test whether general patterns of decomposition differ between major groups of marine plants. We intend then to compare how major nutrients (N and P) are released during decomposition of litter from different marine plants using data obtained by the authors. In order to compare published data on decomposition in a systematic manner, we have chosen to standardize these data using Westrich and Berner’s (1984) multi-G decomposition model. Where possible, we apply this model specifically to carbon (C) and nutrient (N and P) pools of the detritus to compare how materials, especially major nutrients, are recycled via decomposition.

2. DECOMPOSTION Decomposition of plant detritus can be viewed as occurring in stages (e.g., Aber & Melillo, 2001), starting with a rapid leaching phase where labile matter (low molecular-weight sugars and small amino-acids) are abiotically lost from the newly dead cells, followed by a slower loss of material caused by microbial activity (i.e., decomposition). Substrate quality and thus, susceptibility to microbial activity, varies among the different compounds contained in plant litter and these compounds are therefore often decomposed at different rates. Small amino-acids and low molecularweight sugars are thus decomposed faster than high molecular-weight sugars and cellulose, which are decomposed faster than lignin. Some of these compounds may further contain fractions that are resistant to microbial degradation and plant litter therefore often contains a refractory fraction that is only decomposed over very long time scales. Decomposition and mineralization rates thus vary widely among detritus originating from different plants groups because the composition (i.e., content of nutrients, structural tissues etc.) of such litter may differ substantially. Decomposition of litter is most often described as an exponential decrease in biomass (G) with time (Olson, 1963):

Gt = G0 e − k t

(eq. 1)

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where t is time and k is the litter specific rate constant. This equation represents a simplified description of decomposition since it is assumed that all substrates contained in the litter are processed simultaneously and completely. This is of course not true (see above) and a better alternative to this model is therefore Westrich and Berner’s multi-G model (as modified in Westrich & Berner, 1984). Briefly, this model assumes that detritus is made up by several organic fractions (G1, G2, etc.) that decompose at different rates (ki) albeit at the same time. Thus, the overall pattern of decomposition is described as the sum of the patterns of each component plus eventually a refractory pool of matter (R) that does not appear to decompose within the time studied:

Gt = G0 e − k0 t + G1 e − k1 t + G2 e − k2 t + ... + Gi e − ki t + R

(eq. 2)

Often, only one actively decomposing pool of organic matter can be deduced from the decomposition data available in addition to an eventual refractory pool and in such cases the model simplifies to:

Gt = G0 e − k t + R

(eq. 3)

where the amount of organic matter remaining at time t, Gt, is a function of the exponential decay, with rate constant k, of a decomposable pool, G0, and a refractory pool, R, of organic matter. In this chapter we will focus on the k and R parameters as these indicate, respectively, how rapidly organic matter is decomposed (or associated nutrients are mineralized) and, to what extent organic matter (or nutrients) may be preserved in refractory detritus. It is not, however, common practice to report both these parameters (i.e., k and R) since most studies apply the simple model of exponential decay (equation 1). The application of a model without a refractory pool (R) may, however, have important consequences for the estimation of decomposition rates as can be illustrated by the example in Figure 1. In this example, weight losses of litter (i.e., decomposition) proceeded for a period of approximately 150 days after which further changes occurred very slowly. Thus, in this example there was a refractory, or residual, pool of organic matter corresponding to ca. 13% of the original sample weight. Two different models – one with a residual component included (equation 3) and one without (equation 1) – were fitted to the same data and decomposition rates (k) were estimated. The estimated value of k was 3.7 d-1 when a model including a residual fraction was applied whereas it was 2.9 d-1 when a model without a residual fraction was applied. Hence, the estimated decomposition rate (k) becomes severely underestimated (by more than 20% in this case) when a refractory pool is present, but not considered, since an exponential decomposition model without a refractory term must approach zero by some time.

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% Remaining

100

Gt = G0e-kt + R

100

75

G0 = 86%

G = G0e-kt

75

-1

k = 3.7% d

50

k = 2.9% d-1

50

25

25

R = 14% 0

0

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Days

150

200

0 0

50

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200

Days

Figure 1. Application of simple G decomposition models (Westrich & Berner, 1984) with (left) and without (right) a parameter describing the refractory pool (R).

The larger the size of the refractory pool, the larger the error related to estimates of k. This can lead to poor model fits and large errors for estimates for decomposition rate constants. We have therefore only included data in this review for which we were able to determine both the decomposition rate constant (k) and the size of a refractory pool (R) if present. While the focus of this chapter is nutrient remineralization through decomposition, there is remarkably little data collected on nutrient release from marine plant detritus during decomposition and even fewer studies where rates of mineralization are compared across marine plant types. Most studies on decomposition of marine primary producers report decomposition rates, i.e. temporal changes in biomass (expressed in units of dry weight or carbon) and not mineralization rates (see, e.g., Enríquez et al., 1993). We take therefore two different approaches in this chapter. First, we briefly review the literature available on decomposition for marine primary producers and compare and interpret these data in terms of its potential consequences for nutrient release or storage, assuming that the element composition remains constant during decomposition (i.e., that the rate and degree of N and P mineralization is the same as that of C). Next, we present the limited data set of directly measured changes in N and P pools during decomposition for selected plant groups to test whether N and P are released faster or slower than organic matter is decomposed. We will also present some of our own data to examine the stoichiometry of C, N and P during the decomposition of a range of coastal macrophytes. Finally, we will try to illustrate the consequences of these differences in decomposition patterns for nutrient retention on coastal ecosystems based on two case studies based on two Danish estuaries which differ in macrophyte communities. The information reviewed in this chapter should be interpreted in light of what are the consequences of changes in coastal plant communities for nutrient cycling, a main theme of this book, via the detrital decomposition pathway.

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3. COMPARING PATTERNS OF DECOMPOSITION AND NUTRIENT RELEASE DURING DECOMPOSITION AMONG MARINE PRIMARY PRODUCERS Decomposition parameters k and R for different groups of coastal marine primary producers that could be extracted from the available literature are summarized in Figure 2. Only data sets where it was possible to assess the size of R (whether =0 or not) were included in this review. Decomposition rate constants (k) range from 0.0006 to 0.71 d-1 across the entire data set and vary substantially within each plant group. Some general patterns among the three plant groups are evident, however. The median decomposition rate for of seagrass litter (median value = 0.020 d-1) is significantly lower than that of detritus originating from macro- and microalgae (median values = 0.042 d-1 and 0.046 d-1, respectively) (Kruskal Wallis, p = 0.003). The decomposition rates for microalgae and macroalgae were quite similar, however, both in terms of the range observed and media values, and were not distinguishably different. It is evident from Figure 2 that the detritus from all the major plant groups often contain an apparently refractory pool of detritus (R). This is not always the case, however, as R was equal to 0 (i.e., no refractory pool could be detected) for nearly half (46%) of the litter decomposition curves we fitted. The size of R ranged widely from 0% to 86% of the total detrital biomass across the entire data set. Surprisingly, macroalgae had significantly smaller refractory pools (median = 0%) than either seagrasses (median = 16%) or microalgae (median = 33%) (Kruskal Wallis, p = 0.006). The latter two groups did not differ significantly in the size of the refractory pool due to the surprisingly high refractory pools estimated for microalgae. Part of the explanation for the relatively large refractory pools for microalgae may be due to the fact that studies of microalgae decomposition were generally shorter in duration than those for larger macrophytes and decomposition may not be fully completed, making it difficult to estimate the final stages of decomposition. This may lead to an overestimate of R. Another factor, however, that may lead to an underestimate of R for macrophyte detritus (both macroalgae and seagrasses) is that the litter bag technique typically used for studying macrophyte detritus is inherently different than the techniques used for studying the decomposition of microalgae (e.g., culture studies, flask incubations, radioactive labeling experiments). With the litter bag technique, once particles are reduced to below the mesh size of a litter bag they are likely to be lost and counted as decomposed, even if they may not truly be fully decomposed. In contrast, the techniques used with microalgae are more likely to fully account for all particulate microalgae detritus until it is completely decomposed. Still, it is apparent that all detritus types, including microalgae, often contain refractory pools which may be significant when considering long-term nutrient retention.

Phytoplankton Phytoplankton Phytoplankton Benthic microalgae

Twilley et al., 1986 Lee & Fisher, 1992 Andersen, 1996 Andersen & Kristensen, 1992

Yes Yes Yes No*

No No No No

No

Phytoplankton

Westrich & Berner, 1984

Yes

Phytoplankton

Garber, 1984

Yes

No

No Yes

No

Yes

Thalassiosira angstii Scrippsiella trochoidea Skeletonema costatum Chaetoceros tricornutum Skeletonema costatum seston Natural community (dominated by the diatom Leptocylindrus spp.) Chlorella spp. Thalassiosira psuedonana Skeletonema costatum Benthic penate diatoms

Phytoplankton

N data? No Yes

R >0? Yes Yes

Species estuarine phytoplankton Scenedesmus sp.

Primary producer Phytoplankton Phytoplankton

Study Jewell & McCarty, 1971 Otsuki & Hanya, 1972a; Otsuki & Hanya, 1972b R. C. Newell, Lucas, & Linley, 1981

No No No No

No

Yes

No

No

P data? No No

Table 1. Studies included in our review of decomposition of estuarine primary producers (Figs. 2&3). Indicated is what type of primary producer was studied, whether it was possible to explicitly determine that the refractory fraction (R) was >0 (in some cases* it was necessary to assume R=0 to fit the data). Also indicated is whether changes in nutrient content (N & P) during decomposition were reported.

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G.T. Banta, M.F. Pedersen and S.L. Nielsen

Macroalgae

Macroalgae Macroalgage Macroalgae

Macroalgae Macroalgae Macroalgae Macroalgae

Rice & Tenore, 1981

Birch, Gabrielson, & Hamel, Gabrielson, Birch, & Hamel, 1983 Twilley et al., 1986 Buchsbaum et al., 1991

Kristensen, Andersen, & Blackburn, 1992

Christensen et al., 1994 Kristensen, 1994

Bourgues, Auby, deWit, & Labourg, 1996

Chaeteromorpha linum Ulva lactuca Fucus vesiculosus Monostroma obscurum

Ulva lactuca Ulva lactuca Gracilaria tikvahiae Fucus vesiculosus Chondrus crispus

Macroalgae

Josselyn & Mathieson, 1980

1983;

Fucus vesiculosus Macrocystis intergrifolia Nerocystis luetkeana Ascophyllum nodosum Fucus vesiculosus Gracelaria foliifera Hypnea musciformis Sargassum filapendula Spatoglossum schroederi Cladophora albida

Macroalgae Macroalgae

Hunter, 1976 Albright, Chocair, Masuda, & Valdés, 1980

Species

Primary producer

Table 1 (continued)

Yes

Yes

No Yes

Yes No

Yes No* Yes No

Yes Yes

Yes

Yes

Yes

Yes No

N data?

No No

Yes

Yes

No*

Yes No

R >0?

Yes

Yes No

No No

Yes No

Yes

No

Yes

No No

P data?

Decomposition of marine primary producers 193

Macroalgae

Banta, Pedersen unpublished Macroalgae Seagrasses Seagrasses Seagrasses

Seagrasses Seagrasses Seagrasses Seagrasses Seagrasses

Pedersen, unpublished

Godshalk & Wetzel, 1978 Josselyn & Mathieson, 1980 Rice & Tenore, 1981

Pellikaan, 1982 Rublee & Roman, 1982 Gallagher, Kibby, & Skirvin, 1984 Kenworthy & Thayer, 1984 Kristensen, 1994

Nielsen,

Macroalgae Macroalgae Macroalgae

Brouwer, 1996 Bourgues et al., 1996 Paalme, Kukk, Kotta, & Orav, 2002

&

Primary producer

Table 1 (continued)

Zostera marina Zostera marina Thalassia testudinum Syringodium filiform Zostera marina Thalassia testudinum Zostera marina Zostera marina Zostera marina Ruppia maritima

Desmarestia anceps Monostroma obscurum Cladorphora glomerata Pilayella littoralis Ulva lactuca Polysiphonia fucoids Fucus vesiculosus Ceramium rubrum Sargassum muticum Halidrys sliquosa

Species

No Yes Yes Yes No

Yes Yes No Yes Yes

Yes Yes Yes

Yes Yes

No Yes Yes Yes Yes

Yes

Yes Yes Yes

N data?

Yes

No Yes Yes

R >0?

No

Yes No No

No Yes No

Yes Yes

Yes

Yes Yes Yes

P data?

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G.T. Banta, M.F. Pedersen and S.L. Nielsen

Nielsen,

Seagrasses Seagrasses Seagrasses Seagrasses Seagrasses Seagrasses

Halophila stipulacea Ruppia maritima Zostera marina Zostera noltii Cymodocea nodosa Cymodocea nodosa Zostera marina Zostera noltii Posidonia oceanica Enhalus acoroides Zostera marina (including roots)

Seagrasses Seagrasses Seagrasses Seagrasses

Peduzzi & Herndl, 1991 Christensen et al., 1994 Bourgues et al., 1996 Mateo & Romero, 1997 Holmer & Olsen, 2002 Banta, Pedersen & unpublished

Thalassia testudum

Seagrasses

S. Y. Newell, Fell, Statzelltallman, Miller, & Cefalu, 1984 Wahbeh & Mahasneh, 1985 Twilley et al., 1986 Buchsbaum et al., 1991 Hemminga & Nieuwenhuize, 1991

Species

Primary producer

Table 1 (continued)

No

Yes Yes Yes Yes No Yes

Yes Yes No Yes

R >0?

Yes No Yes Yes Yes Yes

No Yes Yes Yes

Yes

N data?

No Yes Yes Yes No Yes

No Yes No Yes

No

P data?

Decomposition of marine primary producers 195

196

G.T. Banta, M.F. Pedersen and S.L. Nielsen 100

10

80

1

R (%)

k (d-1)

60 0.1

40

0.01

20 0.001

0

0.0001

e

a alg cro Mi

ae alg cro Ma

e

s se ras ag Se

a alg cro Mi

ae alg c ro Ma

s se

ras ag Se

Figure 2. Box plot of decomposition parameters (k and R) for different groups of marine primary producers based on changes in biomass (dry weight) or carbon (C). Note log scale for k-values (left). Boxes encompass 25 and 75 quartiles while median values are shown as horizontal lines within each box. Error bars represent the 5 and 95 percentiles. Data outside the 5 and 95% percentiles are plotted as single observations. Note log scale for k (left). The number of observations is 38, 58 and 20 for seagrasses, macroalgae and microalgae, respectively. The studies included are listed in Table 1.

The studies reviewed here show that detritus from micro- and macroalgae are decomposed faster than detritus originating from seagrasses, which corresponds very well to the findings of Enríquez et al. (1993) and Cebrián (1999) although their datasets on marine plants were smaller than the one presented here. In addition, it is evident from our review that macroalgae have a tendency to decompose more completely than other detritus types. Thus, the slower and more incomplete decomposition of litter from slow-growing macrophytes (seagrasses in particular) may lead to a larger accumulation of dead organic matter in systems dominated by seagrasses when compared to systems dominated by macro- and microalgae provided that the net production per unit area is the same (Cebrián, 1999; Duarte & Cebrián, 1996). Seagrasses especially may therefore act as a sink for carbon due to their slow decomposition and the preservation of refractory organic matter. As previously mentioned, nutrient mineralization patterns should be evaluated from changes in nutrient pools of the detritus during decomposition rather than from changes in biomass of that detritus because nutrients may be lost at different rates than biomass is decomposed. There are, however, fewer available data sets on changes in pool size of major nutrients during decomposition so estimates for the decomposition parameters (k and R) expressed in units of N and P are therefore less common. The data we could find where it was possible to estimate k and R for N pools are presented in Figure 3. The k-values based on temporal changes in N-pools span a similar range as those based on changes in biomass (Figure 2) although it was

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not possible to detect any significant differences among plant groups (Kruskal Wallis, p = 0.30). This is despite the fact that it appeared that N in microalgae decomposes more rapidly (median = 0.068 d-1) compared to seagrasses and macroalgae (median = 0.040 d-1 and 0.038 d-1, respectively). The highest values of k based on N tend to be higher than those based on changes in biomass (kC), but it is difficult to make a proper comparison between the two data sets as there are fewer studies where we could assess N dynamics. 100 80

1

R (%)

k (d-1)

60

0.1

0.01

40 20 0

0.001

e

e

a alg cro Mi

lga

a cro Ma

es

ss

ra ag Se

e

e

a alg cro Mi

lga

a cro Ma

es

ss

ra ag Se

Figure 3. Box plot of decomposition parameters (k and R) based on changes in nitrogen (N) pools in different groups of marine primary producers. Note log scale for k-values (left). Boxes encompass 25 and 75 quartiles while median values are shown as horizontal lines within each box. Error bars represent the 5 and 95 percentiles. Data outside the 5 and 95% percentiles are plotted as single observations. Data where extremely high k-values (> 1 d-1) indicated leaching, not decomposition losses, were excluded. The number of observations is 29, 29 and 6 for seagrasses, macroalgae and microalgae, respectively. The studies included are listed in Table 1.

Examining the size of the refractory N-pool (R) in those studies, we observed that macroalgae (median = 0%) again had significant smaller refractory N-pools than microalgae (median = 31%) (Kruskal Wallis, p = 0.04). Seagrasses had intermediate sizes of refractory N-pools (median = 21%) and were not significantly different than the other two plant groups. This was similar to the pattern we observed for organic matter pools. There were even fewer studies where we could determine the decomposition parameters for P-pools (Fig. 4). It was not possible to detect significant differences among plant groups for either k-values or R (Kruskal Wallis p = 0.51 and 0.15, respectively) for P data, given in part the fewer data points. The patterns suggested in Figure 4 are different than those observed for organic matter (Fig. 2) or N (Fig. 3), however, with kP’s being greatest for microalgae (median = 0.13 d-1) and RP’s being lowest (median = 4 %) compared to macroalgae and seagrasses. This pattern is more typical of what one would have expected due to differences in composition (i.e., degree of structural components such as cellulose and lignin) of the different plant

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groups. As expected, microalgae lose P more rapidly and completely compared to macrophytes. Macroalgae and seagrasses lose P at similar rates (k medians = 0.039 d1 and 0.047 d-1, respectively) while that latter group has the largest amount of P retained in refractory pools (R medians = 16% and 25%, respectively). 10

100 80

1

R (%)

k (d-1)

60 0.1

40 20

0.01

0 0.001

e

a alg cro Mi

e

lga

a cro Ma

es

ss

ra ag Se

e

a alg cro Mi

e

lga

a cro Ma

es

ss

ra ag Se

Figure 4. Box plot of decomposition parameters (k and R) based on changes in phosphorous (P) pools in different groups of marine primary producers. Note log scale for k-values (left). Boxes encompass 25 and 75 quartiles while median values are shown as horizontal lines within each box. Error bars represent the 5 and 95 percentiles. Data outside the 5 and 95% percentiles are plotted as single observations. Data where extremely high k-values (> 1 d-1) indicated leaching, not decomposition, losses were excluded. The number of observations is 15, 24 and 4 for seagrasses, macroalgae and microalgae, respectively. The studies included are listed in Table 1.

Comparing the decomposition dynamics for the different elements within coastal marine plant groups, it seems that C, N and P are decomposed and mineralized at similar rates for macroalgae as k-values for macroalgae in Figures 2-4 are similar. It appears that P may be retained somewhat more than C or N as the median value of R for P (16%) is larger than for either organic matter (C) or N (both medians = 0%). There is a great deal of variability in R values for macroalgae, however, indicating a wide range of nutrient retentions in refractory pools for macroalgae. In contrast to macroalgae, the decomposition dynamics for C, N and P are not similar for either seagrasses or microalgae. Mineralization rates for N and P are generally faster (i.e., higher k-values) than for C for both plant groups, especially for P. For seagrasses there is also a tendency for there to be a larger refractory pool of N and P compared to C. That is not the case for microalgae, however, where there are smaller refractory pools, especially of P. This conclusion is based on precariously few data sets, however, and should be considered cautiously. In any case, it is clear that mineralization dynamics for nutrients cannot be easily predicted based on patterns for organic matter decomposition for seagrasses and microalgae, while it appears that one might have better success with macroalgae.

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Trying to summarize from this review of the literature, we conclude that the mineralization dynamics of nutrients do differ among coastal marine plant groups. The decomposition of microalgae is generally more rapid than for macrophytes and nutrient mineralization rates for microalgae are more rapid than organic matter decomposition rates. There does appear to be the potential for nutrient retention (N in particular) in a refractory component of microalgae detritus and this is a phenomenon that warrants further research. Seagrasses on the other hand decompose more slowly although they do exhibit higher rates of nutrient mineralization (which still are lower than microalgae, though) compared to rates of C decomposition. Seagrasses thus have a greater potential to retain nutrients than microalgae, both due to the slower rates of mineralization and to significant refractory pools. Macroalgae are quite variable, but can characterized as having relatively rapid rates of organic matter decomposition and nutrient mineralization and containing small refractory pools. Thus, they appear to have a much lower potential to retain nutrients than the other plant groups. These conclusions are greatly generalized given the fact that they are based on a wide range of data sets collected for many species under differing circumstances and should be examined more specifically, based on the specific plant groups in various environments. Still, they do present an overall hypothesis worth considering about the importance and significance of the various plant groups in coastal environments in terms of their role for nutrient retention.

4. CHANGES IN C, N AND P COMPOSITION DURING DECOMPOSITION The data reviewed in Figures 2-4 suffer from substantial variation within each plant group which makes specific comparisons of decomposition dynamics somewhat unreliable. Part of the variation can be attributed to species-specific variations within each group. For example, morphological, physical and functional properties that may serve to determine decomposition patterns vary enormously among different types of macroalgae and great variations in k and R-values are therefore to be expected. Part of the observed variation may further be due to the variable conditions under which these data were collected by different authors (i.e. in situ versus culture studies, aerobic versus anaerobic conditions, variable temperatures, etc.). These sources of variation make it difficult to precisely compare decomposition parameters (i.e. k and R) between groups and difficult to compare specific patterns of decomposition and mineralization for different plant groups. Overall, the results in Figures 2-4 do, however, suggest general differences among plant groups and do indicate that decomposition patterns for organic matter, N and P are not the same and need to be assessed individually. Surprisingly, relatively few studies have systematically compared the changes that occur in the C, N and P pools of detritus during decomposition in marine plants, which is in contrast to the extensive work published on decomposition of terrestrial and salt marsh detritus (e.g., Melillo, Aber, & Muratore, 1982; Staaf & Berg, 1982; Valiela et al., 1985). The following study was therefore designed to allow a direct comparison of decomposition and mineralization patterns among four species of marine macrophytes, which represent different functional types of the most common macrophytes in temperate (specifically Northern European) coastal waters. Briefly,

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we investigated the anaerobic decomposition of two species of ephemeral macroalgae (Ulva lactuca, Ceramium rubrum), one species of persistent macroalgae (Fucus vesiculosus) and one seagrass species (Zostera marina, both below- and aboveground organs were considered) using litterbags (1 mm mesh) which were buried in estuarine sediments in the laboratory for nearly two years. We did not include microalgae in this study, even though phytoplankton and other microalgae often contribute significantly to total production of organic matter in coastal ecosystems, because the litter bag technique is not well-suited to study the compositional changes during decomposition of such small cells. Long term changes in carbon biomass (i.e., decomposition) are shown in Figure 5 while decomposition rates and the estimated size of refractory pools are given in Table 2. Decomposition rates based on changes in carbon biomass varied substantially and were fastest for the two ephemeral species and slowest for the below-ground parts of Zostera marina while the decay rates of detritus from slowgrowing macroalgae Fucus vesiculosus and leaves of Zostera were intermediate between those extremes (Table 2).

% C remaining

100

100

75

75

Ulva Fucus Ceramium Ulva fit Fucus fit Ceramium fit

50 25

Zostera Zostera roots Zostera fit Zostera roots fit

50 25 0

0

0

100

200

300

400

500

0

100

200

300

400

500

Days

Figure 5. Long term changes in carbon biomass during anaerobic decomposition of detritus from three species of macroalgae (Ulva lactuca, Ceramium rubrum and Fucus vesiculosus; left) and the seagrass Zostera marina (above- and below-ground organs; right). Data are presented as % of the original sample size. Experimental temperature = 15° C and salinity = 15‰ (Banta, Pedersen and Nielsen, unpublished). The curves represent the fits of a 1-G decomposition model with a refractory pool (equation 3). Parameters for these fits are summarized in Table 2.

The pool size of refractory detritus differed also substantially between the detritus types examined - detritus from Ulva lactuca and Ceramium rubrum was completely decomposed after 9 months while there was 7%, of the original Fucus vesiculosus detritus left. The seagrass species studied, Zostera marina, contained even larger refractory pools with 15% and more than 50% of the detritus from Zostera leaves and Zostera roots and rhizomes, respectively, remaining after almost two years (Fig. 5, Table 2).

Species

Ulva lactuca Ceramium rubrum Fucus vesiculosus Zostera marina Zostera marina

Primary producer

Green macroalgae

Red macroalgae

Brown macroalgae

Seagrass (leaves)

Seagrass (roots & rhizomes)

0.026 ± 0.011 0 ± 26

0.040 ± 0.010 0±7

0.038 ± 0.014 0* 0.028 ± 0.005 7±4 0.033 ± 0.006 15 ± 3 0.013 ± 0.005 57 ± 4

k R k R k R k R

0.003 ± 0.012 86 ± 80

0.053 ± 0.012 14 ± 3

0.005 ± 0.007 0 ± 79

0.022 ± 0.010 0±9

N

C

Decomposition Parameter k R

0.006 ± 0.003 26 ± 14

0.055 ± 0.007 12 ± 2

0.019 ± 0.010 3 ± 11

0.029 ± 0.015 0*

0.039 ± 0.011 4±7

P

Table 2. Decomposition parameters k (d-1) and R (% of initial sample size) for selected estuarine macrophytes. Data were obtained by following long term changes in C, N and P pools under anoxic conditions. Parameters were obtained by fitting the data to a 1-G decomposition model (Westrich and Berner 1984) and are presented as mean ± SE. * No refractory pool was evident so R was set to 0 and a simple exponential decay was fitted for the data. (Banta, Pedersen & Nielsen, unpublished).

Decomposition of marine primary producers 201

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G.T. Banta, M.F. Pedersen and S.L. Nielsen

These patterns, as well as the specific decomposition parameters, are consistent with the general patterns we observed in our review of macrophyte decomposition studies (Figs. 2-4), although we are able to distinguish differences between the fast-growing ephemeral and slow-growing perennial macroalgae species in our study. Seagrass detritus, especially that from the underground parts, clearly stands out from the other plant types with lower rates of decomposition and much greater refractory pools. Given that the production of seagrasses roots and rhizomes often represents 20-50% of the total seagrass production (Duarte et al., 1998), the slower and less complete decomposition of this detritus type may have significant consequences for the accumulation of organic matter in shallow coastal ecosystems (Duarte & Cebrián, 1996). Dominance of seagrasses and, to some extent also by slow-growing macroalgae, may potentially have a similar effect on accumulation and burial of plant-bound nutrients and, hence, on nutrient retention at the ecosystem level.

Ulva Fucus Ceramium Ulva fit Fucus fit Ceramium fit

% N remaining

100 75

100 75

50 25

25

0

0

100

200

300

400

Ulva Fucus Ceramium Ulva fit Fucus fit Ceramium fit

100

% P remaining

Zostera Zostera roots Zostera fit Zostera roots fit

50

75 50

500

0

0

100

200

300

400

500

Zostera Zostera roots Zostera fit Zostera roots fit

100 75 50

25

25

0 0

100

200

300

400

500

0 0

100

200

300

400

500

Days

Figure 6. Long term changes in N (top) and P(bottom) pools during anaerobic decomposition of detritus from three species of macroalgae (Ulva lactuca, Ceramium rubrum and Fucus vesiculosus; left) and the seagrass Zostera marina (above- and below-ground organs; right). Data are presented as % of the original sample size. Experimental temperature = 15° C and salinity = 15‰ (Banta, Pedersen and Nielsen, unpublished). The curves represent the fits of a 1-G decomposition model with a refractory pool (equation 3). Parameters for these fits are summarized in Table 2.

Decomposition of marine primary producers

203

Whether or not mineralization will proceed with the same speed as decomposition depends on the rate and extent by which nutrients are released from detritus during decay. The decomposition patterns and parameters for N and P for the studied macrophytes are presented in Figure 6 and Table 2. While the overall patterns among macrophytes are similar to that of C, with the ephemeral macroalgae being decomposed more rapidly and completely than the slower growing macrophytes, there are noticeable differences in the decomposition dynamics of C, N and P for all of the macrophytes. Decay rates for N (kN) were generally lower than when expressed in units of carbon (Table 2). This was especially evident for detritus from Fucus and the below-ground parts of Zostera, which experienced a net increase in N-pool size during the initial stages of decomposition. Substantial losses of N from Fucus detritus appeared first after 2 months of decomposition, after which N was lost rapidly and almost completely (RN not significantly different from 0). The N-pool within detritus from Zostera roots/rhizomes remained, in contrast, nearly intact during 2 years of decomposition (figure 7) resulting in a very low decomposition rate (kN = 0.003 d-1) and a substantial refractory pool of detritus-bound N (RN = 86% of initial sample size). Surprisingly, the loss rate of N in detritus from Zostera leaves was high, but this detritus type contained also a significant refractory N pool (RN = 14%). Decay rates for P (kP) were generally slower than for C but faster than for N (Table 2) but rates for the different types of detritus were ranked as for N, i.e., Zostera leaves>Ulva>Ceramium>Fucus>>Zostera roots/rhizomes. Refractory pools of P were generally of the same size as those for C except in the case of Zostera roots/rhizomes where the refractory detritus pool was severely depleted in P. The fact that both N and P mineralization generally occurred more slowly than C decomposition in our study differs from the overall patterns observed in our review of the decomposition literature (Figs. 2-4). There N and P mineralization either occurs at similar rates (macroalgae) or more rapidly (microalgae and seagrasses) than C decomposition. Part of the explanation for this apparent discrepancy lies in the fact that the literature review allows general but not specific comparisons under a given set of conditions. One of the many factors that can vary both between and within species is their nutrient composition. Therefore we now will consider the results of our decomposition study based on the specific nutrient composition of the macrophytes studied. 5. NUTRIENT RATIOS – IMPLICATIONS FOR MINERALIZATION AND IMMOBILIZATION Changes in nutrient composition during decomposition are most easily analyzed and interpreted by examining changes in tissue nutrient ratios with time. Microbial decomposers are nutrient rich and balanced bacterial growth require substrates with an approximate C:N:P-ratio of about 106:12:1 (Goldman, Caron, & Dennett, 1987). The nutritional quality of the substrate may thus determine whether microbial activity becomes limited by lack of energy (i.e., carbon) or by nutrients and, hence, determine how the elemental composition of detritus will change during decomposition. Nutrient rich detritus may experience a net loss of nutrients relative to carbon because the decomposers become carbon-limited and excrete excess nutrients, while nutrient poor detritus may be enriched with nutrients because

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100

1000

80

800

C:P-ratio (molar)

C:N-ratio (molar)

decomposers contribute nutrients to the matrix of detritus and bacteria from the surrounding medium (Swift, Heal, & Anderson, 1979). In extreme cases, the nutrient pool size of detritus may remain almost intact or even increase while the total biomass of detritus decreases (e.g., Staaf & Berg, 1982). Different plant types differ considerably from each other in their chemical composition (e.g., Duarte, 1992). Microalgae and fast-growing macroalgae tend to be richer in both N and P than slow-growing macroalgae and seagrasses, because the former groups contain more actively metabolizing tissues and the latter contains more structural tissues. The same pattern was seen among the five detritus types included here – fresh detritus from Zostera marina and Fucus vesiculosus were deprived in both N and P relative to C, whereas detritus from Ulva and Ceramium were relatively rich on these nutrients (Fig. 5). Tissue concentrations of N and P are seasonally variable, but these plants were all sampled during mid-summer where N and P concentrations are generally low. Tissue concentrations of N and P and N:Pratios indicated that the detritus from Ceramium was well balanced relatively to the substrate requirements of bacteria, while Ulva lactuca and leaves of Zostera were marginally deprived of both N and P relative to C. Detritus from Fucus vesiculosus and the below-ground parts of Zostera were, in comparison, very poor in both nutrients, particularly N (both had N:P-ratios ≤10).

60 40 20

600 400 200

0

0 Ceramium

Ulva

Fucus

Zostera (leaves)

Zostera (roots)

Ceramium

Ulva

Fucus

Zostera (leaves)

Zostera (roots)

Figure 7. Initial C:N and C:P ratios (mean ± s.d.) of Ulva lactuca, Ceramium rubrum, Fucus vesiculosus and Zostera marina (above- and below-grounds organs, respectively) used in the experiment (Banta, Pedersen and Nielsen, unpublished).

The nutritional quality of detritus from the different plant types leads us to expect that losses of N and P during decomposition should follow losses of C in Ceramium, but be slower than for C (i.e. increasing C:N and C:P-ratios) among the other species and especially so for Fucus vesiculosus and the below-ground parts of Zostera. The data in Figure 8 show that the C:N-ratio of detritus from Ceramium rubrum and from leaves of Zostera marina did remain more or less constant during decomposition. Carbon and nitrogen were thus lost at approximate similar rates for these detritus types. Constant C:N-ratios and relatively fast decomposition rates for both these detritus types suggest that the quality of this detritus was relatively high as expected based on start C:N. The C:N-ratios decreased, in contrast, significantly

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with time in detritus originating from Ulva lactuca, Fucus vesiculosus and the roots and rhizomes of Z. marina indicating a more rapid loss of C relative to N or, alternatively, that microbial immobilization of N during the decomposition was able to enrich these N-poor detritus types. The tissue concentration of C tended to decrease with time for all three detritus types while N concentrations increased significantly (data not shown). The increase in N concentration within detritus from Fucus and from the below-ground parts of Zostera was even so large that the total pool size of nitrogen increased to more than 100% of the initial size during early stages of decomposition (Fig. 6, top). The detritus of Ulva, Fucus and below-ground parts of Zostera experienced thus a net enrichment of N during decomposition, which may be important since it may slow down the rates by which net mineralization from these detritus types occur. These changes during decomposition and colonization by microorganisms are consistent with starting C:N ratios observed for those detritus types With the exception of the below-ground portions of Zostera, there were no marked differences in the relative size of the refractory pools of C and N among the different macrophytes (Table 2). The refractory pool of Zostera root and rhizomes did contain substantially higher amounts of N than fresh detritus (much lower C:N ratio, Fig. 8), however, which may be important in relation to burial and retention of N. The C:P-ratios of detritus originating from Ceramium and Ulva decreased during decomposition (Fig. 8) showing that C was lost faster than P and/or that detritus was enriched with P by microflora associated with the detritus. The patterns were opposite for detritus from Fucus and Zostera leaves and root/rhizomes – C:P-ratios increased (albeit not significantly for Fucus) especially during the later stages of decomposition. Thus, P was lost much faster than C from the detritus. The decrease in C:P-ratios of detritus from Ceramium and Ulva corresponded well to the observation that fresh detritus of these plants were relatively depleted in P. Colonization of microbial decomposers should in theory contribute P to the mixture of detritus and microflora. Fucus and Zostera had even higher C:P-ratios than the two ephemeral algae, but experienced never the less a rapid loss of P during decomposition. Low N:P-ratios (range 10-15) indicated, however, that these plants were more severely depleted in N relative to P. P may therefore have been available in excess from the detritus for the microbes relative to N, which was limiting relative to P for their requirements, and therefore P may have been preferentially excreted (i.e. mineralized). Changes in N:P-ratios obviously reflect largely the changes in C:N and C:P-ratios, but they do provide insights into the relative nutrient limitations for microbes utilizing these detritus types. The N:P-ratio within detritus from Ceramium decreased slightly due to the weak immobilization of P while that of Ulva remained largely unaffected over time suggesting good nutrient balance for microorganisms. The N:P-ratios of detritus from Fucus and Zostera increased, in contrast, markedly due to the parallel immobilization of N and rapid losses of P. The immobilization of P and N, respectively, is a good indication of which nutrient is limiting for the decomposer microbes and therefore which nutrient is most likely to be retained during the decomposition of that detritus.

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70 60 50

C/N

70

Ulva Fucus Ceramium

50

40

40

30

30

20

20

10

10

0

0

0

100

200

300

400

500

2500

0

100

200

300

400

500

300

400

500

300

400

500

2500 Ulva Fucus Ceramium

2000

C/P

Zostera Zostera Roots

60

Zostera Zostera roots

2000

1500

1500

1000

1000

500

500

0

0

0

100

200

300

400

500

150

0

100

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100 Zostera Zostera roots

N/P

100 50

50

Ulva Fucus Ceramium

0

0

0

100

200

300

400

500

0

100

200

Days

Figure 8. C, N and P relationships (mean ± s.d.) during decomposition of detritus from three species of macroalgae and the seagrass Zostera marina (both leaves and roots/rhizomes). Details of the study are given in Figure5 (Banta, Pedersen & Nielsen, unpublished).

In overall terms, decomposition lead to a net enrichment of detritus by the most depleted nutrient (most often N) and rapid losses of replete nutrient species (often P in our study). Only few studies of decomposition among marine plants also include information on stoichiometric changes during decomposition. Results from those few studies vary in similar fashion as our results did. For example, Buchsbaum et al. (1991) and Schmidt (1980) found that N and C was released at approximately similar rates during decomposition of detritus from Fucus vesiculosus, Laminaria

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sp. and Zostera marina leaves under in situ conditions. In contrast, Rublee and Roman (1982) and Kristensen (1994) found that N was released slower than C in detritus from macroalgae (Ulva lactuca and Fucus vesiculosus) and seagrasses (Zostera marina, Ruppia maritima and Thalassia testudinum). Finally, Paalme et al. (2002) found that N was released more slowly and P faster than dry weight biomass was decomposed in detritus from the two ephemeral macroalgae Cladophora glomerata and Pilyella littoralis which is consistent with our results for Fucus and Zostera. The most likely explanation of these varying degrees of N and P release or immobilization is the starting nutrient status (i.e., C:N:P) of the detritus studied relative to the needs of decomposer microorganisms. Availability of nutrients in the environment may also play a role, but a systematic investigation of nutrient availability and coastal macrophyte decomposition is lacking. Finally, whether decomposition occurs aerobically or anaerobically also affects nutrient stoichiometry because anaerobic bacteria are generally less efficient in utilizing C and thus retain less of the nutrients remineralized per mole C during decomposition (Blackburn, 1979; Canfield, Kristensen, & Thamdrup, in press). Results from both our study and others do suggest that stoichiometric changes in detritus during decomposition can vary substantially, likely governed by the availability of internal and external nutrients. More work should be focused on this area if we are to better understand how the decomposition of macrophytes affects the recycling of nutrients in coastal ecosystems. If major nutrients are immobilized in detritus by microbial decomposers in significant amounts due to nutrient limitation this may have practical implications for the patterns of nutrient mineralization and recycling in coastal ecosystems. Substantial nutrient immobilization (N or P) may result in slower mineralization rates and refractory pools of detritus that are more nutrient rich when expressed in units of N or P instead of C. In contrast, if nutrients are available in excess of microbial needs, this may lead to faster nutrient mineralization and recycling rates and the formation of refractory pools of detritus depleted in nutrients relative to C in such environments. 6. DECOMPOSITION PATTERNS - IMPLICATIONS FOR NUTRIENT RETENTION The overall impression from our investigation of decomposition patterns of coastal primary producers is that the decomposition of fast growing, ephemeral algae occurs rapidly and completely with only slight differences in the rates of release of different components (C, N and P). The below-ground parts of seagrasses represent the opposite extreme where decomposition occurs very slowly and is less complete. Nutrients (especially N) are released much more slowly from seagrass roots and rhizomes when compared to C, indicating that limiting nutrients are immobilized during the decomposition of this poor quality detritus. Slower growing, perennial macroalgae and seagrass leaves represent an intermediate between these two extremes of decomposition with faster and more complete decomposition compared to seagrass roots/rhizomes, but slower than for the fast-growing ephemeral algae. This suggests that slow-growing macroalgae may play an intermediate role in nutrient retention, slowing the recycling of nutrients via decomposition compared to the case of fast growing, ephemeral macroalgae.

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The role of microalgae in terms of nutrient retention is less clear as it has not been studied as extensively or in comparable fashion to macrophytes. Decomposition of microalgae occurs rapidly, but the presence of a refractory component could lead to nutrient retention in preserved organic matter deposits. More attention should be focused on refractory components of microalgae detritus and its potential role in nutrient retention. Also, much more work should be done on the decomposition of benthic microalgae. This important group of coastal primary producers has been overlooked almost entirely, probably due in part to the difficulties in studying the decomposition of microalgae so closely associated with sediment particles. Given the extensive background information on decomposition and nutrient mineralization dynamics for coastal primary producers presented so far, we can now consider the consequences of changes in coastal “plant” communities due to, e.g., eutrophication, for nutrient recycling and retention via detritus and decomposition. As reviewed in Chapter 3, one of the main consequences of eutrophication is to cause a shift in primary producer dominance from slow-growing seagrasses and brown macroalgae to fast-growing, ephemeral macroaglae (greens and reds) and phytoplankton in coastal ecosystems. Based on the review in this chapter, such changes in coastal plant communities towards the latter groups would be expected to lead to faster and more complete nutrient recycling, i.e., less nutrient retention. Systems dominated by seagrasses, on the other hand, will both recycle nutrients more slowly and permanently retain a larger proportion of nutrients in those systems. Based on the expected changes in nutrient cycling, community changes of that type could have a reinforcing effect on the effects of eutrophication by ensuring that the excess nutrients entering eutrophied coastal ecosystems are recycled more rapidly through those systems and remain available to support the greater nutrient demands of those fast-growing primary producers (i.e., ephemeral macroalgae and phytoplankton). It is worth examining this general conclusion in a bit more detail in terms of the decomposition dynamics examined in this chapter, however. While it is true that there are significantly different rates of decomposition and nutrient mineralization among the different coastal primary producers, this, in and of itself, may have little effect on permanent nutrient retention and thus little effect on, for example, eutrophication. If detritus is completely mineralized, then all nutrients will ultimately be returned to the ecosystem via decomposition, albeit over different time scales. (We ignore here other ecosystem processes which remover nutrients from active cycling such as denitrification – see Chapter 10). Thus, the differences in rates of decomposition and nutrient mineralization (i.e., k’s) observed in this chapter are not expected to have as important ecological implications over the long term as might first be expected as k only determines how quickly detritus is decomposed and nutrient are mineralized, not whether they are retained. What is of more importance for nutrient retention is the presence of refractory pools of nutrients for different types of detritus. Such pools are especially prevalent in seagrasses, particularly in their underground components which make up a significant fraction of their biomass. The other macrophytes considered had only small refractory pools, although there was a great deal of variation. Phytoplankton appear to have a significant refractory pool based on the data available, but it is unclear what the ultimate fate of nutrients retained in refractory phytoplankton detritus is in the environment. Based on these observations, we suggest that the

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most important alteration on nutrient retention due to plant uptake and sequestering in biomass occurs when costal communities go from being dominated by rooted macrophytes (seagrasses) to other primary producers. This shift in community composition is one of the first effects of eutrophication (see Chapter 3) which may, in turn, make it difficult to reverse the effects of eutrophication later given the changes in nutrient cycling dynamics of the different primary producer communities. Unfortunately, few studies have examined the role of coastal “plant” community changes on nutrient dynamics from this perspective. It is an area worth further consideration in the future. 7. CASE STUDY – DETRITUS DYNAMICS IN TWO SMALL ESTUARIES DOMINATED BY DIFFERENT PRIMARY PRODUCERS The foregoing comparison shows that patterns of decomposition and mineralization differ substantially among different types of marine primary producers. Furthermore, we and others have inferred that such differences may have significant consequences for the turn-over, accumulation and, eventually, burial of detritus and associated nutrients such that more detritus and nutrients accumulate and become buried in systems dominated by slow-growing, perennial macrophytes (e.g., Cebrián, 1999; Duarte, 1995; Enríquez et al., 1993). Whether or not dominance of slow-growing, perennial macrophytes does in fact lead to a higher accumulation of detritus and retention of associated nutrients remains, however, to be systematically tested. The best way to evaluate this problem is to compare the detritus dynamics in systems dominated by different types of primary producers. We conducted a detailed survey based upon repeated sampling in two shallow estuaries in Isefjorden (Denmark) which provided data on seasonal changes in plant community composition, biomass, primary production, grazing losses and tissue nutrient concentrations. The daily production of detritus from each of the major plant types was estimated from measured production, grazing losses and changes in biomass between successive sampling events. Export losses were not accounted for in this study and the estimated production of detritus may therefore have been slightly over-estimated (see Chapter 4). Dynamic modeling was subsequently used to simulate the dynamics of plant-derived detritus and associated nitrogen over a year in both systems. Mineralization of nitrogen was simulated by use of the equations that describe decomposition of different types of plant litter expressed in units of nitrogen. Decomposition of microalgae was simulated on the basis of the decomposition patterns of Skeletonema costatum presented by Garber while decomposition of the remaining macrophytes was based on the parameters presented in Figures 5 & 6 and Table 2. The two estuaries, Vellerup Vig and Tempelkrogen, are small and shallow and they receive approximately the same amounts of nitrogen from external sources on an annual basis (Table 3). The plant communities of the two estuaries differed, however, considerably from each other. Total plant biomass and total primary production were greater in Vellerup Vig than in Tempelkrogen. Furthermore, the plant biomass in Vellerup Vig was dominated by perennial macrophytes (mainly seagrasses) whereas the plant biomass in Tempelkrogen was completely dominated

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by microalgae (phytoplankton and benthic microalgae) and ephemeral macroalgae belonging to the genera Ulva, Enteromorpha, Ceramium, Polysiphonia and Pilayella (Fig. 9, top).

Table 3. Approximate area, depth, annual nitrogen loading, total plant biomass and annual primary production in Vellerup Vig and Tempelkrogen, Isefjorden, Denmark.

Vellerup Vig 2

Area (m )

0,5×10

Average depth (m) -2

-1

External N-loading (g N m year ) -2

Plant biomass (g C m ) Total net primary (g C m-2 year-1)

production

6

Tempelkrogen

4,0×106

2,5 (max: 6 m)

1,5 (max: 4 m)

38

44

33,1

17,7

956

684

Total annual production of autochthonous detritus reached ca. 710 g C m-2 year-1 in Vellerup Vig and almost 480 g C m-2 year-1 in Tempelkrogen (Fig. 9, middle), corresponding to ca. 79 and 52 g detritus-bound N m-2 year-1 (Table 4). The production of detritus-bound nitrogen from slow-growing, perennial macrophytes was relatively small in both systems, even in Vellerup Vig where perennial macrophytes dominated plant biomass. This is partly because the production by these plant types contribute relatively little to total primary production (Fig. 9, bottom) but also because the detritus derived from these plants contained much less nitrogen than detritus from microalgae and ephemeral macroalgae. Seasonal variations in the standing stocks of detritus-bound nitrogen reflect the balance between inputs (i.e. production of detritus-bound nitrogen) and remineralization through decomposition. The amount of detritus-bound N in Vellerup Vig (Fig. 10) varied substantially across the season, ranging from a few grams of N m-2 in winter to more than 20 g N m-2 in April-May. The spring peak in detritus-bound nitrogen resulted from a substantial production of detritus, which appeared at the termination of a macroalgal bloom by the end of April. The detritus from these plants decomposed rapidly, however, and mineralization rates exceeded those of production during summer and fall, leading to a rapid decline in the standing stocks of detritus-bound nitrogen. The input of detritus-bound nitrogen from slow-growing, perennial macrophytes remained low relative to that from microalgae and ephemeral macroalgae, but this detritus decomposed slowly and incompletely, leading to a continuous accumulation of refractory detritus over time. The amount of nitrogen contained in detritus from these plants represented therefore more than 70% of the detritus-bound nitrogen by the end of the year (3.7 g N m-2, Table 4). The clear dominance of detritus from plant types that decompose rapidly and nearly completely resulted in low standing stocks of detritus-bound nitrogen by the end of

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the year. The amount of detritus-bound nitrogen in Vellerup Vig was therefore only 5 g N m-2 by the end of December, off which 62% was contained in the refractory component of the detritus.

Vellerup Vig

Tempelkrogen Total biomass: 17,7 g C m-2

Total biomass: 33,1 g C m-2 Perennial macrophytes 11%

Microalgae 24%

M icroalgae 34%

Ephemeral macroalgae 12% Perennial macrophytes 64%

Ephemeral macroalgae 55%

Total net production: 956 g C m-2 year-1 Perennial macrophytes 19%

M icroalgae 48%

Total net production: 684 g C m-2 year-1 Ephemeral macroalgae 39%

Microalgae 59%

Ephemeral macroalgae 33%

Total production of detritus: 707 g C m-2 year-1 Perennial macrophytes 25%

Ephemeral macroalgae 38%

Perennial macrophytes 2%

Microalgae 37%

Total production of detritus: 483 g C m-2 year-1 Perennial macrophytes 9%

Microalgae 52%

Ephemeral macroalgae 45%

Figure 9. The contribution of different plant types (microalgae, ephemeral macroalgae and perennial macrophytes) to total plant biomass, net primary production and production of detritus in Vellerup Vig (left column) and Tempelkrogen (right column), Isefjorden, Denmark (Pedersen, Banta and Nielsen, unpublished).

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Figure 10. Seasonal changes in the standing stock of detritus-bound nitrogen in Vellerup Vig (left) and Tempelkrogen (right), Isefjorden, Denmark. Temporal changes reflect the balance between production of detritus-bound nitrogen (from different plant types) and losses of nitrogen through mineralization. Data from Pedersen, Banta and Nielsen (unpublished)

The annual production of detritus-bound nitrogen in Tempelkrogen reached 52 g N m-2 and microalgae and ephemeral macroalgae provided by far most of that detritus (Table 4). Decomposition and mineralization was fast and the standing stocks of detritus-bound nitrogen remained always smaller and less seasonal variable than in Vellerup Vig. Maximum levels of detritus-bound nitrogen reached only 6-7 g m-2 in mid-summer and declined to 1.5 g N m-2 by the end of the year, corresponding to less than 3% of the amount produced annually (Fig. 10). The amount of nitrogen bound in refractory detritus from slow-growing perennial macrophytes by the end of December was only 0.4 g N m-2, or about 40% of the total standing stock (Table 4). The amount of detritus-bound nitrogen that accumulates within these systems is thus markedly higher when the plant community is dominated by slow-growing, perennial macrophytes rather than by fast-growing, ephemeral algae (including microalgae). Part of this difference can be attributed to differences in the amount of detritus produced, but most is due to the slow and incomplete turn-over of detritus originating from perennial plants. In the present case, for example, the production of detritus-bound N in Vellerup Vig exceeded that of Tempelkrogen by 50%, while the standing-stock of detritus-bound N in Vellerup Vig was 3.6 times higher than that in Tempelkrogen after one year. Dominance of slow-growing, perennial macrophytes did not only stimulate a greater accumulation of total and “labile” detritus (1.9 g N m-2 in Vellerup Vig versus 1.1 m-2 in Tempelkrogen), but also to a much more dramatic increase in the amount of nitrogen bound in refractory detritus (3.3 versus 0.4 g N m-2). The latter is important in relation to retention of nutrients at the ecosystem scale since refractory detritus and associated nutrients may be especially prone to permanent burial.

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Table 4. Summary of the dynamics of detritus-bound nitrogen in Vellerup Vig and Tempelkrogen, Isefjorden, Denmark (Pedersen, Banta and Nielsen, unpublished).

Vellerup Vig

Tempelkrogen

Production of detritus-bound N (g N m-2 year-1): Microalgae Ephemeral macroalgae Perennial macrophytes Total

44.0 26.0 7.6 78,8

35.8 14.5 2.0 52.3

Standing stock of detritus-bound N after one year (g N m-2): Microalgae Ephemeral macroalgae Perennial macrophytes Total

0.7 0.6 3.7 5.0

0.1 0.7 0.6 1.4

Labile detritus (g N m-2) Refractory detritus (g N m-2)

1.9 3.1

1.1 0.4

It is finally worth noting that most of the nutrients that were assimilated by plants and became tied up in detritus were in fact recycled rapidly and almost completely through decomposition over rather short time-scales. Thus, 93% and 97% of the detritus-bound nitrogen that was “produced” in Vellerup Vig and Tempelkrogen, respectively, was mineralized within the same year as it was produced. Consequently, less than 7% and 3%, respectively, of the nitrogen that was originally taken up by primary producers and incorporated in “plant” detritus could ever be subject to permanent burial, suggesting that marine plants may play a relatively minor role for the burial and the permanent retention of nutrients in coastal marine environments compared to nutrient loadings. Burial of nitrogen in estuarine and coastal sediments average about 5 g N m-2 year-1 (range: 0.2 – 14.3 g N m-2 year-1)(Nixon, Granger, & Nowicki, 1996). We do not know the rates of permanent N-burial in Vellerup Vig and Tempelkrogen but the estimated amount of nitrogen accumulated in refractory detritus was 3.1 g N m-2 year-1 in Vellerup Vig and 0.4 g N m-2 year-1 in Tempelkrogen. Long term burial of plant derived detritus along with associated nutrients could thus contribute significantly to total burial of nitrogen in Vellerup Vig but not in Tempelkrogen, assuming that total rates of burial are similar to those found elsewhere. The key to this difference is the difference in plant community composition, namely the presence of significant perennial macrophytes, especially seagrasses. Thus, while the absolute effects of permanent nutrient retention via plant detritus were relatively small in terms of nutrient loadings in both systems, the only system where it occurred at all was Vellerup Vig where seagrasses were prevalent. This is consistent with the conclusions we have drawn earlier in this chapter.

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8. CONCLUSION In this chapter we have reviewed decomposition and nutrient mineralization patterns for coastal primary producers with the goal of shedding light on the role of different plant groups for retaining assimilated nutrients in their detritus. While it is clear that there is a great deal of variation, both within and between species, there are general differences in decomposition patterns that have implications for nutrient recycling and retention. The production of refractory pools of nutrients, which particularly occurs with seagrass detritus, is especially relevant for the long-term retention of nutrients. Other plant groups (micro- macroalgae) are not expected to retain nutrients in detrital pools on a permanent basis. The extent to which particular nutrients are retained depends on a variety of factors including the availability or limitation of nutrients for the decomposer microorganisms in the environment as this affects the stoichiometry of nutrient release during decomposition. As such, this could be expected to vary greatly from system to system. More attention should be given to nutrient stoichiometry during detrital mineralization and recycling. Finally, as was evident form the case study examined, while nutrient uptake by coastal primary producers is very important for coastal nutrient dynamics, most of those nutrients are in fact recycled to the environment again via decomposition. Still, a fraction of those nutrients can be permanently retained in refractory detrital pools provided they are taken up by slow-growing perennial macrophytes (especially seagrasses). Nutrients assimilated by faster-growing primary producers will be efficiently recycled within the ecosystem via decomposition with little loss or retention. Thus changes in coastal plant communities will have some impact on coastal nutrient cycling dynamics, the importance of which will depend on the degree of external loadings in the environment of interest. 9. REFERENCES Aber, J. D., & Melillo, J. M. (2001). Terrestrial Ecosystems (2nd ed.). San Diego, USA: Academic Press. Albright, L. J., Chocair, J., Masuda, K., & Valdés, M. (1980). In situ degradation of the kekps Macrocystis integrifolia and Nerocystis luetkeana in British Columbia coastal waters. Le Naturaliste Canadien, 107(1), 3-10. Andersen, F. O. (1996). Fate of organic carbon added as diatom cells to oxic and anoxic marine sediment microcosms. Marine Ecology Progress Series, 134(1-3), 225-233. Andersen, F. O., & Kristensen, E. (1992). The importance of benthic macrofauna in decomposition of microalgae in a coastal marine sediment. Limnology and Oceanography, 37, 1392-1403. Birch, P. B., Gabrielson, J. O., & Hamel, K. S. (1983). Decomposition of Cladophora .1. Field Studies in the Peel-Harvey Estuarine System, Western Australia. Botanica Marina, 26(4), 165-171. Blackburn, T. H. (1979). Nitrogen/carbon ratios and the rates of ammonia turnover in anoxic sediments. Paper presented at the Microbial Degradation of Pollutants in Marine Environments, Gulf Breeze, FL. Borum, J., & Sand-Jensen, K. (1996). Is total primary production in shallow coastal marine waters stimulated by nitrogen loading? Oikos, 76(2), 406-410. Bourgues, S., Auby, I., deWit, R., & Labourg, P. J. (1996). Differential anaerobic decomposition of seagrass (Zostera noltii) and macroalgal (Monostroma obscurum) biomass from Arcachon Bay (France). Hydrobiologia, 329(1-3), 121-131. Brouwer, P. E. M. (1996). Decomposition in situ of the sublittoral Antarctic macroalga Desmarestia anceps Montagne. Polar Biology, 16(2), 129-137. Buchsbaum, R., Valiela, I., Swain, T., Dzierzeski, M., & Allen, S. (1991). Available and refractory nitrogen in detritus of coastal vascular plants and macroalgae. Marine Ecology Progress Series, 72(12), 131-143.

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Canfield, D. E., Kristensen, E., & Thamdrup, B. (in press). Aquatic Geomicrobiology. Cebrián, J. (1999). Patterns in the fate of production in plant communities. American Naturalist, 154(4), 449-468. Christensen, P. B., Møhlenberg, F., Krause-Jensen, D., Jensen, H. S., Rysgaard, S., Clausen, P., et al. (1994). Stoftransport og stofomsætning i Kertinge Nor/Kertiminde Fjord (No. 43). Copenhagen, Denmark: Miljøstyrelsen. Duarte, C. M. (1992). Nutrient Concentration of Aquatic Plants - Patterns across Species. Limnology and Oceanography, 37(4), 882. Duarte, C. M. (1995). Submerged Aquatic Vegetation in Relation to Different Nutrient Regimes. Ophelia, 41, 87. Duarte, C. M., & Cebrián, J. (1996). The fate of marine autotrophic production. Limnology and Oceanography, 41(8), 1758-1766. Duarte, C. M., Merino, M., Agawin, N. S. R., Uri, J., Fortes, M. D., Gallegos, M. E., et al. (1998). Root production and belowground seagrass biomass. Marine Ecology Progress Series, 171, 97-108. Enríquez, S., Duarte, C. M., & Sand-Jensen, K. (1993). Patterns in decomposition rates among photosynthetic organisms: the importance of detrital C:N:P content. Oecologia., 94, 457-471. Gabrielson, J. O., Birch, P. B., & Hamel, K. S. (1983). Decomposition of Cladophora .2. Invitro Studies of Nitrogen and Phosphorus Regeneration. Botanica Marina, 26(4), 173-179. Gallagher, J. L., Kibby, H. V., & Skirvin, K. W. (1984). Detritus Processing and Mineral Cycling in Seagrass (Zostera) Litter in an Oregon Salt-Marsh. Aquatic Botany, 20(1-2), 97-108. Garber, J. H. (1984). Laboratory study of nitrogen and phosphorus remineralization during the decomposition of coastal plankton and seston. Estuarine and Coastal Shelf Science, 18, 685-702. Godshalk, G. L., & Wetzel, R. G. (1978). Decomposition of Aquatic Angiosperms .3. Zostera-Marina L and a Conceptual-Model of Decomposition. Aquatic Botany, 5(4), 329-354. Goldman, J. C., Caron, D. A., & Dennett, M. R. (1987). Regulation of Gross Growth Efficiency and Ammonium Regeneration in Bacteria by Substrate C-N Ratio. Limnology and Oceanography, 32(6), 1239. Hemminga, M. A., & Nieuwenhuize, J. (1991). Transport, Deposition and Insitu Decay of Seagrasses in a Tropical Mudflat Area (Banc Darguin, Mauritania). Netherlands Journal of Sea Research, 27(2), 183-190. Holmer, M., & Olsen, A. B. (2002). Role of decomposition of mangrove and seagrass detritus in sediment carbon and nitrogen cycling in a tropical mangrove forest. Marine Ecology Progress Series, 230, 87101. Hunter, R. D. (1976). Changes in Carbon and Nitrogen-Content During Decomposition of 3 Macrophytes in Freshwater and Marine Environments. Hydrobiologia, 51(2), 119-128. Jewell, W. J., & McCarty, P. L. (1971). Aerobic decomposition of algae. Environmental Science and Technology, 5, 1023-1031. Josselyn, M. N., & Mathieson, A. C. (1980). Seasonal Influx and Decomposition of Autochthonous Macrophyte Litter in a North Temperature Estuary. Hydrobiologia, 71(3), 197-207. Kenworthy, W. J., & Thayer, G. W. (1984). Production and Decomposition of the Roots and Rhizomes of Seagrasses, Zostera-Marina and Thalassia-Testudinum, in Temperate and Sub-Tropical Marine Ecosystems. Bulletin of Marine Science, 35(3), 364-379. Kristensen, E. (1994). Decomposition of Macroalgae, Vascular Plants and Sediment Detritus in Seawater - Use of Stepwise Thermogravimetry. Biogeochemistry, 26(1), 1-24. Kristensen, E., Andersen, F. O., & Blackburn, T. H. (1992). Effects of benthic macrofauna and temperature on degradation of macroalgal detritus: The fate of organic carbon. Limnology and Oceanography, 37, 1404-1419. Lee, B.-G., & Fisher, N. S. (1992). Degradation and elemental release rates from phytoplankton debris and their geochemical implications. Limnology and Oceanography, 37, 1345-1360. Mateo, M. A., & Romero, J. (1997). Detritus dynamics in the seagrass Posidonia oceanica: Elements for an ecosystem carbon and nutrient budget. Marine Ecology Progress Series, 151(1-3), 43-53. Melillo, J. M., Aber, J. D., & Muratore, J. F. (1982). Nitrogen and Lignin Control of Hardwood Leaf Litter Decomposition Dynamics. Ecology, 63(3), 621-626. Newell, R. C., Lucas, M. I., & Linley, E. A. S. (1981). Rate of degradation and efficiency of conversion of phytoplankton debris by marine micro-organisms. Marine Ecology Progress Series, 6, 123-136. Newell, S. Y., Fell, J. W., Statzelltallman, A., Miller, C., & Cefalu, R. (1984). Carbon and Nitrogen Dynamics in Decomposing Leaves of 3 Coastal Marine Vascular Plants of the Subtropics. Aquatic Botany, 19(1-2), 183-192.

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Nixon, S. W., Granger, S. L., & Nowicki, B. L. (1996). An assessment of the annual mass balance of carbon, nitrogen, and phosphorus in Narragansett Bay (vol 31, pg 15, 1995). Biogeochemistry, 33(3), 217. Olson, J. S. (1963). Energy-Storage and Balance of Producers and Decomposers in Ecological-Systems. Ecology, 44(2), 322-&. Otsuki, A., & Hanya, T. (1972a). Production of Dissolved Organic-Matter from Dead Green Algal Cells .1. Aerobic Microbial Decomposition. Limnology and Oceanography, 17(2), 248-&. Otsuki, A., & Hanya, T. (1972b). Production of Dissolved Organic-Matter from Dead Green Algal Cells .2. Anaerobic Microbial Decomposition. Limnology and Oceanography, 17(2), 258-&. Peduzzi, P., & Herndl, G. J. (1991). Decomposition and Significance of Seagrass Leaf Litter (Cymodocea-Nodosa) for the Microbial Food Web in Coastal Waters (Gulf of Trieste, Northern Adriatic Sea). Marine Ecology Progress Series, 71(2), 163-174. Pellikaan, C. G. (1982). Decomposition processes of eelgrass, Zostera marina L. Hydrobiological Bulletin, 16(1), 83-92. Paalme, T., Kukk, H., Kotta, J., & Orav, H. (2002). 'In vitro' and 'in situ' decomposition of nuisance macroalgae Cladophora glomerata and Pilayella littoralis. Hydrobiologia, 475(1), 469-476. Rice, D. L., & Tenore, K. R. (1981). Dynamics of Carbon and Nitrogen During the Decomposition of Detritus Derived from Estuarine Macrophytes. Estuarine Coastal and Shelf Science, 13(6), 681-690. Rublee, P. A., & Roman, M. R. (1982). Decomposition of Turtlegrass (Thalassia-Testudinum Konig) in Flowing Sea-Water Tanks and Litterbags - Compositional Changes and Comparison with Natural Particulate Matter. Journal of Experimental Marine Biology and Ecology, 58(1), 47-58. Schmidt, C. (1980). Some aspects of marine algae decomposition. Ophelia., Suppl.1, 257-264. Staaf, H., & Berg, B. (1982). Accumulation and Release of Plant Nutrients in Decomposing Scots Pine Needle Litter - Long-Term Decomposition in a Scots Pine Forest .2. Canadian Journal of BotanyRevue Canadienne De Botanique, 60(8), 1561-1568. Swift, M. J., Heal, O. W., & Anderson, J. M. (1979). Decomposition in terrestrial systems (Vol. 5). Oxford: Blackwell. Twilley, R. R., Ejdung, G., Romare, P., & Kemp, W. M. (1986). A Comparative-Study of Decomposition, Oxygen-Consumption and Nutrient Release for Selected Aquatic Plants Occurring in an Estuarine Environment. Oikos, 47(2), 190-198. Valiela, I., Teal, J. M., Allen, S. D., Van Etten, R., Goehringer, D., & Volkmann, S. (1985). Decomposition in salt marsh ecosystems: the phases and major factors affecting disappearance of above-ground organic matter. Journal of Experimental Marine Biology and Ecology, 89, 29-54. Wahbeh, M. I., & Mahasneh, A. M. (1985). Some Aspects of Decomposition of Leaf Litter of the Seagrass Halophila-Stipulacea from the Gulf of Aqaba (Jordan). Aquatic Botany, 21(3), 237-244. Westrich, J. T., & Berner, R. A. (1984). The role of sedimentary organic matter in bacterial sulfate reduction: the G model tested. Limnology and Oceanography, 29, 236-249.

10. AFFILIATIONS G.T. Banta, M.F. Pedersen & S.L. Nielsen: Department of Life Sciences and Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark.

JACK J. MIDDELBURG, KARLINE SOETAERT, PETER M. J. HERMAN, HENRICUS T.S. BOSCHKER AND CARLO H.R. HEIP

BURIAL OF NUTRIENTS IN COASTAL SEDIMENTS: THE ROLE OF PRIMARY PRODUCERS 1. INTRODUCTION Anthropogenic nutrient inputs to aquatic systems exceed those under pristine conditions. This has resulted in numerous changes in their functioning of which eutrophication effects, the development of nuisance algal blooms, seasonal anoxia, the disappearance of seagrasses are illustrative, well-documented examples (Heip, 1995; Duarte, 1995). Concern about these changes in ecosystem functioning has initiated large amount of research on nutrient cycling with the ultimate aim to predict consequences of policy measures and management options. One of the basic tools used in the study of nutrient cycling in aquatic ecosystems is the construction of nutrient budgets. Nutrient budgets are very much alike financial bookkeeping systems with revenues and assets on one side and expenses and debts on the other side. Accountancy has developed into a mature discipline that is able to deal with internal and external losses and revenues and with disparate, not always accurate or quantifiable data. For instance, long-term depreciation costs have to be balanced against the expected enhanced profit of an investment. Biogeochemists and ecologists constructing nutrient budgets are faced with similar complexities. On the one hand there are event-like input terms such as atmospheric deposition of nutrients during rain that cover a period of a few hours at most and on the other hand there are long-term (years) loss terms such as the nutrient contained in accumulating sediments. For ecosystem budgets, it is important to distinguish between internal sinks (primary production and nutrient assimilation) and sources (remineralisation) and external (river) input and output terms (denitrification and burial). In this chapter we will address the burial term for nutrients in coastal sediment ecosystems. We will first provide a definition and then discuss the two main components (sediment accumulation and nutrient concentration). Special attention will be given to the influence of primary producers on nutrient burial via enhancement of sediment accumulation rates or via their effect on organic matter production or preservation. 2. BURIAL DEFINED The term burial is used within various subdisciplines with the consequences that it is used for different processes that operate on different time scales. Here we define burial as the longer-term removal (annual to decadal scale) of nutrients from the 217 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 217-230. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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pelagic system, by accumulation of sediments. Geoscientists working on carbon and nutrient recycling often use the term burial efficiency, which expresses the fraction of carbon or nutrient delivered to the sediment that becomes buried. Soil scientists and some plant ecologists use the term burial to distinguish between aerial, exposed degradation and decomposition within the soil or sediments, where biogeochemical conditions are different (e.g. McInerney & Bolger, 2000; Rovira & Ramón Vallejo, 2002). Burial is then assessed for instance by transferring leaf material within litterbags into a deeper layer to study decomposition of organic matter under anoxic conditions. Burial should also not be confused with storage or refractory accumulation within ecosystems sensu Duarte & Cebrian (1996) and Cebrian (2002). Detritus that is not decomposed within the study period (typically a year) accumulates temporarily on the sediment surface. However, before it is removed for a longer period of time, it is subject to further degradation within the sediment with losses of 50 to > 90 % (Heip et al., 1995). Our definition of burial derives from the geosciences (Berner, 1980; Boudreau, 1997). Sediments comprise water and particles, both of which may contain nutrients and which interact. Mineralisation and dissolution will provide a transition from solid to liquid phase, whereas primary production, adsorption and precipitation work the other way around. Two different types of transport affect the distribution of these constituents in the sediment. Accretion of sediment and compaction will remove both liquid and solid substances away from the sediment surface. Molecular diffusion and bioirrigation will further affect dissolved constituents, whilst sediment reworking by animal feeding and movement (bioturbation) or by the action of waves and tides affects particulates. A thorough discussion of these various modes can be found in Boudreau (1997). For our current interest, it suffices to note that (bio)turbation generally operates much slower than molecular diffusion and is limited to the upper layer of the sediment only. Consequently, burial of the solid phase is much more effective than burial of the liquid phase and the latter is usually ignored. Once particulate matter is transferred to below the sediment layer, which is biogeochemically and ecologically active, it is isolated from the short-term biologically controlled cycles and may either become involved in a rapid geological cycle (decades to thousands of years) or even in the long-term geological cycle (million of years). Rapid geological recycling can be related to migrating gullies that erode at one place and cause deposition at another location, or to changes in the relative rise of sea level that can cause drowning or emergence of depositional areas. Mathematically speaking, burial of particulate matter is the product of the solid fraction (1-porosity), concentration of the constituents (Ci mass/mass), the dry density of particles (ρ; mass volume-1) and sediment accretion rate (ω, length yr-1 ).

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where φ is the sediment porosity (volume of water/volume of total sediment). It is important to realize that the accumulation rate (ω) in eq. 1 refers to long-term sediment accretion at sufficient depth below the sediment surface and not to the

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deposition of material at the sediment surface. Particle deposition rates at the sediment surface can be much higher than net accumulation rates because many particles that are deposited on the sediment surface are resuspended. These deposition-resuspension cycles may occur at hourly-daily time scales in tidal systems or at seasonal time scales in many coastal, temperate sediments (Herman et al., 2001). This difference between temporary deposition and long-term accumulation should be kept in mind when constructing nutrient budgets. Dynamic nutrient budgets may include this temporal repository of sedimentary nutrient accumulation, but long-term, annual-scale nutrient budgets should include the burial term as defined in this chapter. Four terms define the removal of nutrients by burial (eq.1): porosity (φ), the dry density of particles (ρ), the deep concentration (Ci) and the sediment accumulation rate (ω). The first two terms are rather invariant and consequently do not contribute significantly to the variance in nutrient burial estimates. Porosity, the fraction of water, varies from 0.4 for coarse sands to 0.8 for mud (Boudreau, 1997). Similarly, the dry density of inorganic sedimentary particulates is usually in the range of 2.4 to 2.6 g cm-3, except for peat with an organic matter content of >45 wt% which may have bulk dry densities less than 2.0 g cm-3. The variability in porosity and dry densities can be neglected relative to those in nutrient concentration and the sediment accumulation rate. Sediment accumulation rate varies over orders of magnitude in coastal sediments and is the prime factor governing burial of carbon and nutrients (Middelburg et al., 1997a). It should be remarked that there is not necessarily a relationship between sediment accumulation rates and nutrient concentration. The mere enhancement of mineral deposition without associated organic matter and nutrients could in principle lead to a lowering of nutrient concentration by dilution but usually there is a coupling between particle and (organic) nutrient deposition and enhanced deposition rates will therefore not result in lower nutrient concentrations. 3. SEDIMENT ACCUMULATION Sediment accumulation refers to the permanent transfer of material from the seasonally influenced, biogeochemically and ecologically active layer to the deeper layer. Quantification of sediment accumulation rates in coastal ecosystems is usually based on the radionuclides such as the primordial ones (210Pb), atmospheric ones (7Be, 14C) or artificial, human-produced radionuclides such as 137Cs (Thomsen et al., 2002). Determination of sediment accumulation rates based on artificial radionuclides boils down to determining the sediment depth at which the event (e.g. Chernobyl May 1986) or maximum release (e.g. 1963 atmospheric testing of nuclear bombs treaty) can be recognized and division by time since that period (Figure 1). Natural and primordial radionuclides normally show an exponentially decaying profile with depth in the sediment due to input from the water column and subsequent radioactive decay. Fitting of an exponential regression allows then calculation of the sediment accumulation (Figure 1). Care should be taken not to apply the procedure to the surface layer which is mixed by wave or current activities

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Figure 1. Sediment accumulation rate determinations based on the radionuclides 137Cs and Pb. Activity versus depth profiles of 137Cs usually show a subsurface maximum related to maximum above-ground nuclear bombing testing in 1963. Dividing the depth of the 137Cs peak with the time period between sampling and 1963 provides an estimate for sediment accumulation (i.e. 20/(2003-1963) = 0.5 cm yr-1). Activity versus depth profiles of 210Pb typically show an exponential decline (straight line on semi-logarithmic graph) with the depth attenuation coefficient providing an estimate of mixing or sediment accumulation rate. The line shown is consistent either with a sediment accumulation rate of 0.5 cm yr-1, a bioturbation coefficient of 8 cm2 yr-1 or a sediment accumulation rate of 0.4 cm yr-1 and a mixing rate of 1.5 cm2 yr-1. Sediment accumulation rate determinations should therefore always be based on more than one natural radionuclide if sediment mixing cannot be excluded a priori, as for instance in deep, anoxic basins. 210

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and animals moving and feeding as here the rate of sediment mixing and accumulation cannot be independently derived on the basis of one radionuclide. One either has to apply additional constraints or use a second radionuclide having a different depth profile, i.e. different decay rate. Similarly, with the interest in longterm accumulation, one should use radionuclides that have comparable long halflives. Given the short half lives of 234Th (24 days) and 7Be (53 days), these radionuclides can only be used for dating in rapidly accumulating sediments subject to limited mixing. However, a detailed study of their inventories (activity per unit area) may provide useful information on seasonal accumulation of material (Canuel et al., 1990). Sediment accumulation rate may vary anywhere between zero or even negative (erosion) to more than 10 cm yr-1 (Heip et al., 1995; Figure 2). This variability is primarily due to variability in sediment supply and availability, and differences in hydrodynamic conditions. Primary producers may affect sediment accumulation rates in a number of ways that differ between macrophytes (submerged and emergent) and benthic microalgae. Several recent studies have shown that sediment stability; hence net sediment accumulation can be enhanced by biofilms and microbial mats (Paterson and Black, 1999). Stabilisation of sediments involves the production of extracellular polysaccharides by benthic organisms, in particular benthic diatoms (Holland et al., 1974; de Brouwer et al., 2000). The mud-diatom system may be of two types (Van de Koppel et al., 2001): one with low microalgal biomass and low mud content and one with high algal biomass and high mud content. Since nutrient contents and availability are related directly to mud content, it is likely that sedimentary microalgae more or less modify their environment so as to optimise nutrient availability (ecosystem engineering, Jones et al., 1997). Parallel to the seasonality in benthic microalgal production, the enhanced stabilisation of sediments due to microalgae is also a seasonal phenomena and this is likely the reason for the seasonal deposition of silty and muddy sediments on tidal flats (Fig. 3). The twoway interactions between microalgae and their hosting sediments are further complicated through intensive grazing and sediment mixing by the infauna (Middelburg et al., 2000; Herman et al., 2000). Canopies of macrophyte communities are known to promote deposition of particles and to reduce resuspension and erosion with the net result that sediment accumulation is enhanced (Scoffin, 1970; Redfield, 1972). This effect has been reported for mangroves, salt marshes, fresh-water marshes and seagrass beds and has been studied in detail for seagrasses (Zostera marina and Posidonia oceanica) and saltmarsh plants such as Spartina anglica. The presence of shoots influences the hydrodynamics of the systems: they reduce water current velocities through extraction of momentum and increase the thickness of the benthic boundary layer (Paterson and Black, 1999; Jumars et al., 2001). The effects of macrophytes on water flow and sediment deposition are reciprocal. For instance, water flow affects the performance of the seagrasses and vice versa (Fonseca et al., 1982, 1983). Moreover, canopies of coastal plants may also reduce wave energy (Fonseca and Calahan, 1992) and affect water turbulence, which either may increase or decrease depending on external conditions and canopy architecture (Jumars et al., 2001). The consequences of this on particle deposition and resuspension are not yet fully understood. Recently, Gacia and Duarte (2001) have shown that canopy-induced

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reduction in resuspension rates is more important than canopy-induced enhancement of particle deposition. Canopies may also provide a habitat for suspension feeders that cause deposition of faeces and pseudofaeces, part of which may be incorporated in the sediment (Heip et al., 1995) Canopies not only increase the total rate of particle accumulation, they also alter the composition due to preferential settling and retention of fine particles within the canopies. This preferential accumulation results in higher nutrient concentration in sediments colonized by macrophytes compared to unvegetated sediments (see below; Kenworthy et al., 1982; Benoy and Kalff, 1999). Finally, the roots and rhizomes of seagrasses, marsh plants and mangrove trees stabilize the sediments and lower erosion rates (Redfield, 1972; Woodroffe, 1992).

Figure 2. Histogram of sediment accumulation rates in unvegetated shelf sediments (based on data summarized in Middelburg et al. (1997a)).

4. NUTRIENT CONCENTRATIONS Plants may affect nutrient concentrations either directly through net accumulation of plant litter or indirectly by their impact on sediment accretion or on sediment biogeochemical conditions. Coastal ecosystems are among the most productive ecosystems: they account for about 20 % of total net primary production (Duarte and Cebrian, 1996). The majority of biomass production is degraded, grazed or exported and only a small

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amount remains available for accumulation (Cebrian, 2002). This refractory accumulation is 4 fold higher for higher plants (10-17 % of net primary production) than for algae (0.4-6 % of gross primary production; Duarte & Cebrian, 1996). This small fraction of higher plant material being preserved can sometimes contribute significantly to sediment accretion and carbon and nutrient burial (Middelburg et al., 1997b). Macrophyte production contributes especially to this sediment accretion due to litter accumulation because their intrinsic degradation rate is lower than that of algae, their degradation rate is more sensitive to redox conditions (see below) and part of their production is already allocated below-ground. Below-ground material has a higher burial potential than above-ground material because most of the latter is degraded and exported before incorporation in the sediments. However, these below-ground organs usually have rather low nutrient contents because of structural materials and the storage of energy-rich (COH-rich) compounds. As canopies promote the deposition and retention of the finer, organic-matter rich particles (see above), one often finds higher nutrient and organic matter concentrations in vegetated relative to non-vegetated sediments (Benoy and Kalff, 1999): e.g. in seagrasses (Kenworthy et al., 1982; Gacia et al., 2002), salt marshes (Middelburg et al., 1997b) and mangroves (Middelburg et al., 1996). By changing the sediment biogeochemistry, plants may also impact nutrient concentrations. Rooted macrophytes and sedimentary algae take up dissolved nutrients and may release chelating components to solubilise mineral associated nutrients, in particular phosphate. The ultimate result of these processes on nutrient burial is not clear. On the one hand plant litter enriched in nutrients is more easily degraded than nutrient poor litter (Enriquez et al., 1993; Cebrian, 1999) and nutrient uptake might thus lower nutrient burial. On the other hand, nutrient uptake stimulates production and thus enhances litter production and burial. Furthermore, plants compete with heterotrophic bacteria for dissolved nutrients (Kaye and Hart, 1997) and it depends on the relative burial efficiency of plant litter and bacterial remains whether nutrient uptake enhances or diminishes nutrient concentrations and burial. Rooted macrophytes living in anoxic sediments have a major influence on sedimentary biogeochemical conditions through release of oxygen from the roots (Sundby et al., 1997). Emergent and submerged plants have developed strategies to supply their roots and rhizome with sufficient oxygen for respiration while the environment is rich in oxygen demanding components (reduced iron and manganese, sulphide, methane, etc). Emergent wetlands may either transport oxygen by diffusion as a gas via their aerenchym or they employ an active ventilation system based on convective flow due to pressure differentials (Van der Nat et al., 1998). Part of the oxygen transported to the roots is respired and part is released into the reducing environments, where it will react with many components resulting in complex biogeochemical, coupled reactions (e.g. coupled nitrificationdenitrification, Reddy et al., 1989; chapter 10). The oxygen released by microalgae or by roots has an influence on the availability and concentrations of phosphorus via the iron cycle. Oxygen release results in the oxidation of sulphide and reduced iron with the consequences that more iron oxide is formed than in the absence of these plants (Roden and Wetzel, 1996; Van der Nat and Middelburg, 1998). In most sediment phosphate is tightly associated with solid phase iron oxides. Enhanced formation of iron oxides prevents escape of phosphate

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generated by decomposing organic matter and may ultimately result in higher phosphor concentrations (Rozan et al., 2002). Phosphor speciation, i.e. the distribution of phosphorus over various phases, is rather complex and is usually defined in the following major pools (Ruttenberg, 1992): adsorbed P, organic P, Feoxide associated P, autogenic apatite P (within sediment formed) and refractory, external apatite P. Oxygen release by benthic microalgae and roots of macrophytes may modify the form under which P becomes buried.

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The most studied and most dramatic effect of plant oxygen release on burial concerns the effect on organic matter mineralisation efficiencies. The questions whether organic matter degradation under oxic conditions is faster and more efficient than anaerobic mineralisation has been addressed in many studies (e.g. Kristensen, 2000). These studies revealed that redox conditions have little influence on the degradation rate of labile organic matter, such as algae and seston (Westrich and Berner, 1984), but that oxic conditions stimulate degradation of more refractory organic material, be it inherently more refractory higher plant material (Kristensen et al., 1995) or modified and aged marine organic matter (Hulthe et al., 1998). Particularly aromatic structures and highly polymeric compounds such as lignin are poorly degraded under anoxic conditions (Benner et al., 1984). The nature of the primary producer can thus have a major influence on the preservation of organic matter, hence burial efficiency. Recently, Dauwe et al. (2001) reported that not only the origin or quality of organic matter determines the presence of an oxygen effect on organic matter preservation. They argue that oxygen-enhanced degradation occurs at low mineralisation levels at which bacterial biomass production becomes limiting. Work by Aller and co-workers (Aller, 1994; Sun et al., 2002) have shown that degradation efficiencies of organic material depend on specific patterns of redox fluctuations. Oxygen release by benthic microalgae and macrophyte roots is governed by illumination. Changes in light availability thus cause short-term redox oscillations that favour extensive degradation of organic matter, hence lowering burial potential (Aller, 1994). 5. IMPORTANCE OF PRIMARY PRODUCERS ON ORGANIC MATTER BURIAL The relative importance of plant litter accumulation to sediment accretion varies among systems and for salt-marsh systems has led to the recognition of two endmember depositional facies: the organogenic and minerogenic modes (Allen, 1995). Organic matter accumulation determines vertical accretion in sediment-starved peaty marshes such as those found in New England, whereas mineral matter accumulation controls accretion in marshes in NW Europe (Middelburg et al., 1997b). King et al. (1992) have shown that mineral deposition may stimulate plant growth and thus stimulate organic matter accumulation indicating that some feedbacks are operational. While local plant production may contribute and sometimes even dominate organic matter and nutrient accumulation and burial, this is normally not the case. This has been shown repeatedly using stable isotopes of carbon. The carbon isotopic composition of organic matter buried in many vegetated coastal systems differs significantly from the isotopic composition of local plant production indicating that a large proportion of the carbon accumulating in these systems is external, i.e. imported from adjacent coastal systems. This will likely also apply for nutrients given the stoichiometric coupling of nutrients with carbon. A number of researchers (Ember et al., 1987; Middelburg et al., 1997b; Boschker et al., 1999) have shown that the organic matter accumulating in salt-marsh sediments is more depleted in carbon isotopes and richer in nitrogen than organic matter derived from Spartina, the dominant plant. They attributed this to trapping of allochthonous organic matter,

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input by benthic microalgae and nitrogen assimilation by heterotrophic bacteria. Other researchers have reported similar observations for seagrass systems (Gacia et al., 2002; Dauby, 1989; Boschker et al., 2000) and mangrove systems (Middelburg et al., 1996). This accumulation of non-local produced plant material can be attributed to preferential degradation of local produced material, selective preservation of allochthonous carbon or to much larger allochthonous than autochthonous inputs. The first two explanations do not apply to macrophyte systems since the algal material imported is more labile than the macrophyte material (Boschker et al., 1999), but it may apply to systems dominated by benthic microalgae. Herman et al. (2000) have shown that the organic matter accumulating in tidal flat sediments is significantly depleted in 13C compared to benthic microalgae and they attributed this to preferential use of microalgal material. 6. CONCLUSIONS While benthic primary producers often create unfavourable conditions for denitrifiers (see chapter 10) and thus limit the natural buffering mechanism of coastal ecosystems to enhanced nitrogen loadings, they do enhance nutrient removal via burial. In this chapter we have presented and discussed the factors governing burial or longterm loss of nutrients in coastal ecosystems hosting various primary producers. The presence or absence and the nature of the primary producer have major implications for the burial of nutrients. Benthic microalgae stabilise sediments via exudation of polymeric substances that glues the sediment and lowers erodability. This is a seasonal phenomenon with the result that most material being deposited during the growing is resuspended lateron (Fig. 3). The ultimate consequences for nutrient removal by burial are not well known, but probably minor. Seagrasses, marsh plants and mangroves stabilise the sediments by their roots and rhizomes and their canopies enhance deposition of particles with the result that their sediments accumulate faster and are usually richer in nutrients. Moreover, a significant fraction of the organic matter produced by these plants accumulates and may ultimately become buried. These macrophyte are therefore of major importance in regional budgets of sediment and nutrient burial. The disappearance of seagrasses due to enhanced nutrient inputs (Duarte, 1995) may consequently result in less nutrient removal by burial and thus a reduction in the resilience of coastal ecosystems towards changing nutrient inputs (Valiela and Cole, 2002).

7. REFERENCES 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. Sedimentary Geology 100: 21-45. Aller R.C. (1994). Bioturbation and remineralization of sedimentary organic matter: effects of redox oscillation. Chemical Geology 114: 331-345. Benner R., Maccubbin A.E. and R.E. Hodson (1984). Anaerobic biodegradation of the lignin and polysaccharide components of lignocellulose and synthetic lignin by sediment microflora. Applied Environmental Microbiology 47: 998-1004.

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Benoy G.A. and J. Kalff (1999). Sediment accumulation and Pb burdens in submerged macrophyte beds. Limnology & Oceanography 44: 1081-1090. Berner, R.A. (1980). Early diagenesis. A theoretical approach. Princeton University press, Oxford. Boschker, H.T.S., J.F.C. de Brouwer and T.E. Cappenberg (1999). The contribution of macrophyte derived organic matter in microbial biomass in salt marsh sediments: stable carbon-isotope analysis of microbial biomarkers. Limnology & Oceanography 44: 309-319. Boschker H.T.S., Wielemaker A., Schaub B.E.M. and Holmer M. (2000). Limited coupling of macrophyte production and bacterial carbon cycling in the sediments of Zostera meadows. Marine Ecology Progress Series 203: 181-189. Boudreau B.P. (1997). Diagenetic models and their implementation. Springer-Verlag, 414pp. Canuel E.A., C.S. Martens and L.K. Benninger (1990). Seasonal variations in Be-7 activity in the sediments of Cape Lookout Bight, North Carolina. Geochimica Cosmochimica Acta 54: 237-245 Cebrian J. (1999). Patterns in the fate of production in plant communities. American Naturalist 154: 449468. Cebrian J. (2002). Variability and control of carbon consumption, export and accumulation in marine communities. Limnology & Oceanography 47: 11-22. Dauby P. (1989). The stable isotope ratios in benthic food webs of the Gulf of Calvi, Corsica. Continental Shelf Research 9: 181-195 Dauwe, B. J.J. Middelburg and P.M.J. Herman (2001). The effect of oxygen on the degradability of organic matter in subidal and intertidal sediments of the North Sea area. Marine Ecology Progress Series 215: 13-22. de Brouwer J.F.C., Bjelic S., de Deckere E.M.G.T., Stal L.J. (2000). Interplay between biology and sedimentology in a mudflat (Biezelingse Ham, Westerschelde, The Netherlands). Continental Shelf Research 20: 1159-1177 Duarte C.M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia 41: 87-112. Duarte C.M. and J. Cebrian (1996). The fate of marine autotrophic production. Limnology & Oceanography 41: 1759-1766. Ember, L.M., Williams, D.F. and Morris, J.T. (1987). Processes that influence carbon isotopic variations in salt marsh sediments. Marine Ecology Progress Series 36: 33-42. Enriquez S., C.M. Duarte and K Sand-Jensen (1993). Patterns in decomposition rates among photosynthetic organisms: The importance of detritus C:N: P content. Oecologia 94: 457-471. Fonseca M.S., Fisher J.S., Zieman J.C. and G.W. Thayer (1982). Influence of seagrass, Zostera marina L., on current flow. Estuarine Coastal Shelf Science 15:351-364. Fonseca M.S., Zieman J.C., G.W. Thayer and J.S. Fisher (1983). The role of current velocity in structuring Eelgrass (Zostera marina L.) meadows. Estuarine Coastal Shelf Science 17:376-380. Fonseca M.S. and J.A. Calahan (1992). A preliminary evaluation of wave attenuation by four species of seagrass. Estuarine Coastal Shelf Science 35:565-576. Gacia E., C.M. Duarte and J.J. Middelburg (2002). Carbon and nutrient deposition in a Mediterranean seagrass (Posidonia oceanica) meadow. Limnology & Oceanography 47: 23-32.

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Gacia E. and C.M. Duarte (2001).Sediment retention by a Mediterranean Posidonia oceanica meadow: the balance between deposition and resuspension. Estuarine Coastal Shelf Science 52: 505-514 Heip, C.H.R. (1995). Eutrophication and zoobenthos dynamics. Ophelia 41: 113-136 Heip C.H.R., N.K. Goosen, P.M.J. Herman, J. Kromkamp, J.J. Middelburg and K. Soetaert (1995) Production and consumption of biological particles in temperate tidal estuaries. Oceanography Marine Biology Annual Reviews 33: 1-150. Herman P.M.J., J.J. Middelburg, J. Widdows, C.H. Lucas and C.H.R. Heip (2000). Stable isotopes as trophic tracers: combining field sampling and manipulative labelling of food resources for macrobenthos. Marine Ecology Progress Series 204:79-92 Herman, P.M.J., J.J. Middelburg and C.H.R. Heip (2001). Benthic community structure and sediment processes on an intertidal flat: results from the ECOFLAT project. Continental Shelf Research 21: 2055-2071 Holland, A.F., R.G. Zingmark and J.M. Dean (1974). Quantitative evidence concerning the stabilization of sediments by marine benthic diatoms. Marine Biology 27: 191-196. Hulthe G., S. Hulth and P.O.J. Hall (1998) Effect of oxygen on degradation rate of refractory and labile organic matter in continental margin sediments. Geochimica Cosmochimica Acta 62: 1319-1328. Jones C.G., J.H. Lawton and M. Shachak (1997). Positive and negative effects of organisms as physical ecosystem engineers. Ecology 78: 1946-1957. Jumars P.A., J.E. Eckman and E. Koch (2001). Macroscopic animals and plants in benthic flows. In: The benthic boundary layer, edited by Boudreau and Jørgensen, p. 320-347, Oxford University Press. Kaye J.P. and S.C. Hart (1997). Competition for nitrogen between plants and soil microorganisms. Trends Ecology Evolution 12: 139-143. Kenworthy W.J., J.C. Zieman and G.W. Thayer (1982). Evidence for the influence of seagrasses on the benthic nitrogen cycle in a coastal plain estuary near Beaufort, North Carolina (USA). Oecologia 54: 152-158. King, G.M., M.J. Klug, R.G. Weigert and A.G. Chalmers (1982). Relation of soil water movement and sulphide concentration to Spartina alterniflora production in a Georgia salt marsh. Science 218: 6163. Kristensen E. (2000). Organic matter diagenesis at the oxic/anoxic interface in coastal marine sediments, with emphasis on the role of burrowing animals. Hydrobiologia 426: 1-24 Kristensen E., S.I. Ahmed and A.H. Devol (1995). Aerobic and anaerobic decomposition of organic matter in marine sediment: which is fastest? Limnology & Oceanography 40: 1430-1437. Leonard L. A. & M.E. Luther, (1995). Flow hydrodynamics in tidal marsh canopies. Limnology & Oceanography 40: 1474-1484. McInerney M. and T. Bolger (2000). Decomposition of Quercus petraea litter: influence of burial, comminution and earthworms. Soil Biology Biochemistry 32: 1989-2000. Middelburg J.J., J. Nieuwenhuize, F.J. Slim and B. Ohowa (1996). Sediment biogeochemistry in an East African mangrove forest (Gazi Bay, Kenya). Biogeochemistry 34: 133-155. Middelburg J.J., K. Soetaert and P.M.J. Herman (1997a). Empirical relationships for use in global diagenetic models. Deep-Sea Research I 44: 327-344. Middelburg J.J., J. Nieuwenhuize, R.K. Lubberts and O. van de Plassche (1997b). Organic carbon isotope systematics of coastal marshes. Estuarine Coastal Shelf Science 45, 681-687.

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Middelburg J.J., C. Barranguet, H.T.S Boschker, P.M.J. Herman, T. Moens and C.H.R. Heip (2000). The fate of intertidal microphytobenthos carbon: an in situ 13C labelling study. Limnology & Oceanography 45: 1224-1234. Paterson D.M. and K.S. Black (1999). Water flow, sediment dynamics and benthic biology. Advances Ecological Research 29: 155-193. Reddy K.R., W.H. Patrick and C.W. Lindau (1989). Nitrification-denitrification at the plant root-sediment interface in wetlands. Limnology & Oceanography 34: 1004-1013. Redfield A.C. (1972). Development of a New England salt marsh. Ecological Monographs. 42: 201-237. Roden E.E. and R.G. Wetzel (1996). Organic carbon oxidation and suppression of methane production by microbial Fe(III) oxide reduction in vegetated and unvegetated freshwater wetland sediments. Limnology & Oceanography 41: 1733-1748 Rovira P. and V.R. Vallejo (2002). Mineralization of carbon and nitrogen from plant debris, as affected by debris size and depth of burial. Soil Biology Biochemistry 34: 327-339. Rozan T.F., M. Taillefert, R.E. Trouwborst, B.T. Glazer, S. Ma, J. Herszage, L.M. Valdes, K.S. Price and G.W. Luther III (2002). Iron-sulphur-phosphorus cycling in the sediments of a shallow coastal bay: Implications for sediment nutrient release and benthic macroalgal blooms. Limnology & Oceanography 47: 1346-1354. Ruttenberg K.C. (1992). Development of a sequential extraction method for different forms of phosphorus in marine sediments. Limnology & Oceanography 37: 1460-1482. Scoffin T.P. (1970). The trapping and binding of subtidal carbonate sediments by marine vegetations in Bimini lagoon, Bahamas. Journal Sedimentary Petrology 40: 249-273. Sun M.Y., R.C. Aller, C. Lee and S.G. Wakeham (2002). Effects of oxygen and redox oscillation on degradation of cell-associated lipids in surficial marine sediments. Geochimica Cosmochimica Acta 11: 2003-2012. Sundby, B., Vale, C., Cacador, I., Catarino, F., Madureira, M.-J. and M. Caetano (1998). Metal-rich concretions on the roots of salt-marsh plants: mechanism and rate of formation. Limnology & Oceanography 43: 245-252 Thomson J., Dyer F.M. and J.W. Croudace (2002). Records of radionuclide deposition in two salt marshes in the United Kingdom with contrasting redox and accumulation conditions. Geochimica Cosmochimica Acta 66: 1011-1023. Valiela I. and M.L. Cole (2002). Comparative evidence that salt marshes and mangroves may protect seagrass meadows from land-derived nitrogen loads. Ecosystems 5: 92-102. Van de Koppel J., P.M.J. Herman, P. Thoolen and C.H.R. Heip (2001). Do alternate stable states occur in natural ecosystems? Evidence from a tidal flat. Ecology 82: 3449-3461. van der Nat F.J.W.A. and J.J. Middelburg (1998). Effect of two common macrophytes on methane dynamics in freshwater sediments. Biogeochemistry 43: 79-104. van der Nat F.J.W.A., J.J. Middelburg, D. van Meteren and A. Wielemakers (1998). Diel methane emission patterns from Scirpus lacustris and Phragmites australis. Biogeochemistry 41: 1-22. Westrich J.T. and R.A. Berner (1984). The role of sedimentary organic matter in bacterial sulphate reduction: the G model tested. Limnology & Oceanography 29: 236-249. Woodroffe C. (1992). Mangrove sediments and geomorphology. In: Tropical mangrove ecosystems, edited by Robertson and Alongi, p. 7-41.

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J.J. Middelburg, K. Soetaert, P.M.J. Herman, H.T.S. Boschker & C.H.R. Heip: Netherlands Institute of Ecology, Centre for Estuarine and Coastal Ecology, P.O. Box 140, NL-4400 AC Yerseke, Netherlands.

KAREN J. MCGLATHERY, KRISTINA SUNDBÄCK, IRIS C. ANDERSON

THE IMPORTANCE OF PRIMARY PRODUCERS FOR BENTHIC NITROGEN AND PHOSPHORUS CYCLING

1. INTRODUCTION In shallow coastal systems where most of the seafloor lies within the photic zone, benthic photoautotrophy plays a key role in regulating nutrient cycling. In these systems, production of submerged vascular plants (seagrasses), macroalgae and benthic microalgae is high and typically exceeds that of phytoplankton (Borum and Sand Jensen 1996). Changes in the patterns of primary production as well as in habitat structure and trophic dynamics (Valiela et al. 1992; Nixon et al. 1996) have been directly related to nutrient over-enrichment; this is also recognized as one of the greatest threats to maintaining marine biodiversity in coastal regions (NRC 2000). The widely-accepted scenario following nutrient enrichment is a shift in the dominance of primary producers, from seagrasses and perennial macroalgae to fastgrowing green macroalgae and phytoplankton (Sand-Jensen and Borum 1991; Duarte 1995; Valiela et al. 1997; see Chapter 3). Interestingly, only one of these models (Sand-Jensen and Borum 1991) has included explicitly the potential importance of benthic microalgae. Ultimately, one might expect a shift to phytoplankton dominance in the most heavily eutrophied shallow estuaries (Duarte 1995; Valiela et al. 1997), even though total production might not change (Borum and Sand-Jensen 1996). Given this transition, the question then becomes, how might such a shift in primary producer dominance influence nutrient cycling in shallow coastal waters? The influence of primary producers on nutrient transformations has important consequences for the role of shallow estuaries as a buffer zone between land and sea (Fig. 1). Uptake and temporary retention of nutrients in plant biomass, burial of recalcitrant organic matter, and the direct effects of autotrophic metabolism on bacterially- and chemically-mediated processes all influence nutrient cycling rates and pathways. Variations in the rates and dominance of these processes as primary producer communities change, will ultimately determine the fate and retention of watershed nutrients on their trajectory to the open ocean.

231 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 231-261. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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Fig. 1. Role of primary producers on the fate and retention of nutrients at the land-sea interface.

In this chapter we address the influence of primary producer groups (seagrasses, benthic microalgae, macroalgae) on nutrient cycling processes. We consider the effects of plant metabolism on biogeochemical transformations within the sediments and nutrient exchanges between the sediments and water column (see Chapter 10 for discussion of denitrification). Our focus is on the retention and loss processes that influence both the responses to nutrient enrichment and the transport of nutrients through these shallow systems. We consider these processes in ecosystems that vary with respect to tidal amplitude, water residence times, seasonality and sediment type. Nutrient assimilation and turnover by benthic primary producers appear to play a dominant role in regulating the magnitude and timing of nutrient fluxes in these systems. This being the case, we pay particular attention to the competitive interactions between primary producer communities and to the fate of nutrients bound in plant tissue.

2. BENTHIC NUTRIENT CYCLING PROCESSES The role of benthic primary producer communities as sources or sinks of nutrients reflects the net influences of assimilatory nutrient uptake, nitrogen fixation, leakage by live and grazed autotrophs, oxygen release, nutrient production and losses due mineralization, and nitrification – denitrification (Fig. 2). Nutrients can be transferred between the sediment and water column as dissolved inorganic or organic forms or as gases. Exchange rates are also influenced by benthic faunal communities (e.g., Pelegri et al. 1994; Hansen and Kristensen 1997) and by physical factors such as sediment resuspension and porewater advection. Often there are diel

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patterns driven by the primary producers, with an efflux of nutrients during the dark and uptake in the light (e.g., Eyre and Ferguson 2002). The balance between nutrient losses and gains in benthic communities also is influenced by nutrient transfer via mobile foraging animals (see Chapter 6), and export or import of particulate organic matter (see Chapter 4). Nutrient cycling in benthic primary producer communities appears to be relatively conservative, with much of the nutrient demand met by recycling within the sediments. For example, Ottosen et al. (unpublished data) found by field 15N-labelling of sandy sediment that 90% of labeled 15 PON disappeared from the top 3 cm of sediment after 12 weeks. This rather long retention could be explained if remineralized N was efficiently recycled within the microbial mat (Lomstein et al. 1998), and could be responsible for sustaining high rates of productivity by benthic microalgae during summer in lownitrogen sandy sediments (Sundbäck et al. 2000). A similar closed recycling of both N and P also appears to apply to benthic microalgal communities in oligotrophic carbonate sediments of coral reefs (Miyajima et al. 2001; Suzumura et al. 2002) and for both temperate and tropical seagrass systems (Alcoverro et al. 2000; McGlathery et al. 2001b; Miyajima et al. 2001; Risgaard-Petersen and Ottosen 2001). Conservative recycling of sediment N to support autotrophic communities is also the common pattern observed in Spartina alterniflora vegetated intertidal sediments (Anderson et al. 1997). Although the nutrient demands of seagrasses appear to be met in large part by internal recycling of organic-bound nutrients in the sediment (Hemminga et al. 1991; Alcoverro et al. 2000; McGlathery et al. 2001b), seagrasses and other primary producers in the community also obtain external nutrients that are transported by water flow. Resorption of nutrients from senescing leaves reduces the demand for either recycled or external nutrients in seagrass (e.g., Pedersen and Borum 1993; Stapel and Hemminga 1997; Alcoverro et al. 2000) and in salt marsh systems (Anderson et al. 1997). 2.1 Seagrasses Seagrasses have a major impact on sediment nutrient cycling by leakage of oxygen and dissolved organic matter (DOM) from the roots. Both processes are photosynthetically-driven. Oxygen produced by photosynthesis in the leaves is transported to the roots to support aerobic respiration via a well-developed lacunal system, a series of air channels comprising up to 60% of the total plant volume. The oxygen that is not respired by the roots is released into the rhizosphere, creating oxidized zones in an otherwise reducing environment. At the same time, seagrasses release DOM from the roots, typically as simple organic carbon compounds that are the recent products of photosynthesis (Koepfler et al. 1993; Ziegler and Benner 1999). Bacterial productivity in the rhizosphere has been linked to these organic exudates. Moriarty et al. (1986) showed that 11% of recent photosynthate was released within 6 hours into the rhizosphere of the tropical seagrass Halodule wrightii. Seasonal and light-induced variation in bacterial activity and the typical sub-surface peak in bacterial processes in the sediments where root biomass is highest, also clearly indicate that root exudates stimulate bacterial activity (e.g., Welsh et al. 1996; McGlathery et al. 1998; Blaabjerg et al. 1998).

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Water

N, P

N2

DNR

N-fix

DOM

Sediment NH4

O2

+

O2 DOM

Microbial Processes

N, P

Particulate Nutrients

Burial

Fig. 2. Influence of benthic primary producers on nutrient cycling processes: assimilatory uptake and leakage, release of oxygen and dissolved organic matter (DOM) that affects microbial activity, and mineralization and/or burial of particulate organic matter. Arrows reflect the attenuation of light and the reduction of nutrient fluxes across the sediment-water interface by benthic primary producer communities. DNR=denitrification, N-fix=nitrogen fixation.

For some seagrass species, including Potamogeton perfoliatus and Zostera marina, it has been shown that oxygen released from plant roots and the creation of oxicanoxic interfaces in the rhizosphere stimulates nitrification, which produces nitrate to support enhanced denitrification (e.g., Flindt 1994; Cornwell et al. 1999). This stimulation may be linked to the seagrass growth cycle and to variations in oxygen release (Caffrey and Kemp 1992; Risgaard-Petersen et al. 1998). However, recent work using the isotope-pairing technique suggests that at least for the temperate seagrasses Z. marina and Z. noltii, enhanced coupled nitrification-denitrification in the rhizosphere is not generally observed (Rysgaard et al. 1996; Risgaard-Petersen and Ottosen 2000, see Chapter 10). The mechanisms responsible for this lack of

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stimulation are thought to be low rates of rhizosphere nitrification and competition for ammonium between nitrifying bacteria and benthic microalgae (Ottosen et al. 1999; Risgaard-Petersen and Ottosen 2000). This does not preclude the possibility that enhanced denitrification may occur in sediments vegetated by other seagrass species, in particular in tropical species such as Thalassia testudinum, which allocate more biomass to roots and rhizomes than temperate species, and which tend to be strongly P-limited in carbonate sediments so that competition between bacteria and plants for N is unlikely to occur. Organic matter concentrations and bacterial oxygen demand tend to be lower in tropical carbonates sediments, which may make oxygenation of the rhizosphere more efficient in these systems. There are relatively few measurements of denitrification in seagrass-vegetated carbonate sediments in tropical/subtropical environments, but rates tend to be high (Blackburn et al. 1994; Miyajima et al. 2001; Kemp and Cornwell 2001). The available data suggest that both denitrification rates and nitrogen fixation rates are higher in tropical/subtropical seagrass meadows than in temperate seagrass meadows (Table 1). Nitrogen fixation activity is enhanced in seagrass meadows relative to unvegetated sediments, both in temperate and tropical environments (e.g., O’Donohue et al. 1991; Welsh et al. 1996; McGlathery et al. 1998; Hansen et al. 2000). The availability of organic substrate is a major factor controlling N fixation rates in seagrass sediments. Many studies have shown increases of N fixation by addition of organic compounds, and seasonal and diel variations that are consistent with the role of photosynthetic exudates stimulating N-fixing bacteria (e.g., McGlathery et al. 1998, Hansen et al. 2000; Welsh 2000). Much of this activity (25 – 95%) has been associated with sulfate reducers (e.g., Welsh et al. 1996; McGlathery et al. 1998), which also are believed to be stimulated by root exudates (Blaabjerg et al. 1998; Hansen et al. 2000). Depth profiles of N fixation activity in Z. marina sediments showed a close association with root/rhizome biomass, with seasonal shifts in the magnitude and depth distribution of N fixation matching the shifts in belowground biomass (McGlathery et al. 1998). Overall, N fixation rates in sediments of seagrass meadows are higher in tropical/subtropical systems than in temperate systems (see references in McGlathery et al. 1998 and Welsh 2000), and this may in part be due to increased DOM release from the greater root/rhizome biomass of many tropical seagrass species. Autotrophic, heterocystous cyanobacterial epiphytes on seagrass leaves also contribute to the total N fixation rates, especially in warm tropical environments, although on an area basis the activity is lower than that of heterotrophic bacteria in the sediments (e.g., Capone and Taylor 1977; O’Donohue et al. 1991). Few studies have measured N fixation and denitrification simultaneously in seagrass beds to determine if N gains by fixation can compensate for N losses via denitrification, and the results are equivocal (Table 1). In a tropical seagrass meadow, Blackburn et al. (1994) observed that N fixation rates accounted for less than half of the N loss by denitrification, whereas in a temperate seagrass meadow, Risgaard-Petersen et al. (1998) found that N fixation more than compensated for the N loss by denitrification. Both oxygen and DOM release by seagrass roots influence the forms and availability of sediment P. Oxidation in the rhizosphere results in the formation of iron oxides that effectively bind P in the solid phase. The redox conditions in the sediment and the formation of autogenic P minerals influence the release of P to overlying waters (Rozan et al. 2002). In carbonate sediments that typically dominate tropical seagrass

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systems, P turnover rates are generally high in the porewaters relative to N, indicating that P is preferentially removed (Hines and Lyons 1982; McGlathery et al. 2001b). The mechanisms responsible for this include surface adsorption onto carbonate grains, direct precipitation of Ca-P mineral phases, and uptake by Plimited primary producers. In vegetated carbonate sediments in Bermuda, Jensen et al. (1998) found that only 2% of the total sediment P was in the loosely-adsorbed pool that could be readily released to the porewater; 15 to 20% was more strongly adsorbed to the surface of the sediment, and the remainder (80%) was bound in the mineral matrix of the sediments, probably in the form of carbonate fluorapatite and other carbonates. It was previously thought that carbonate sediments were a permanent sink for P in these tropical environments and that mineral P was not available to plants. However, recent studies suggest that seagrass metabolism facilitates the dissolution of carbonate minerals in the rhizosphere releasing bound P (Jensen et al. 1998; Burdige and Zimmerman 2003). One mechanism for this is the decrease in pH resulting from root respiration or the stimulation of bacterial respiration by root DOM release (Burdige and Zimmerman 2003). The oxidation of sulfides in the rhizosphere may also decrease pH, as has been shown for saltmarsh grasses (Giblin and Howarth 1984). Another mechanism is the release of organic acids from the roots that also acidify the rhizosphere and cause the dissolution of the carbonate minerals, as is the case for terrestrial plants (Hoffland 1992; Knight et al. 1992). Iron associated with carbonate minerals can serve as a sorption site for inorganic P (Jensen et al. 1998), and also can react with sediment sulfides to precipitate ironsulfide minerals that reduce sulfide toxicity to primary producers (Carlson et al. 1994; Erskine and Koch 2000). Chambers et al. (2001) found that iron additions increased the total sediment P pool and reduced exposure of the seagrass T. testudinum to free sulfide, but found no growth response probably because most of the sediment P was bound in unavailable forms. One might expect a more significant effect of iron additions on seagrasses in carbonate sediments where P is not as strongly limiting (i.e., where P is in more readily exchangeable forms). On an ecosystem scale, seagrass systems can be important temporary sinks for nutrients, although there is less information than for benthic micro- and macro-algal communities. The source-sink role will vary seasonally, depending on the growth requirements of seagrasses. Temperate Z. marina meadows may act as strong sinks for N from the beginning of the growing season in the spring to the time of maximum productivity in late summer; in the fall when decomposition rates are typically relatively high, but plant N demand low, the seagrass community may become a source of nutrients (Risgaard-Petersen et al. 1998; Risgaard-Petersen and Ottosen 2000). Denitrification was an important sink relative to plant uptake in these seagrass meadows only in winter when both plant demand and decomposition were low (Risgaard-Petersen and Otteson 2000). Since N fixation may only support a small percentage of seagrass N demand in these systems (Welsh 2000 and references therein), most of the nutrient influx is likely uptake of water column nutrients by seagrass leaves and associated primary producers (benthic microalgae, epiphytes, macroalgae) and accumulation of imported organic matter sedimented into the meadow.

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Table 1. Comparison of nitrogen fixation and denitrification rates in temperate and tropical/subtropical seagrass meadows.

Seagrass species Temperate Zostera marina Zostera marina Zostera marina Zostera noltii

Location

N Fixation mmol N m-2 d-1

Vaucluse Shores, Virginia Limfjord, Denmark Limfjord, Denmark

0.36 0.07 – 0.43 0.4 0.10 – 0.41

Zostera capricorni

Arcachon Bay, France Arcachon Bay and Etang du Prevost, France Edmunds Bay, Australia

Tropical/ Subtropical Thalassia testudinum Thalassia testudinum

Oyster Bay, Jamaica Florida Bay, Florida

Zostera noltii

Thalassia hemprechii Syringodium isoetifolium Enhalus acoroides Zostera capricorni

Gulf of Carpentaria, Australia Gulf of Carpentaria, Australia Gulf of Carpentaria, Australia Moreton Bay, Australia

Denitrification mmol N m-2 d-1

0.05 – 0.14 0.33

Reference

Capone (1988) McGlathery et al. (1998) RisgaardPetersen et al. (1998) Welsh et al. (2000) Rysgaard et al. (1995)

0.19

Eyre & Ferguson (2002)

1

2-4

0.7 – 4

1.6 - 3

Blackburn et al. (1994) Kemp & Cornwell (2001) Moriarty & O’Donohue (1993) Moriarty & O’Donohue (1993) Moriarty & O’Donohue (1993) O’Donohue et al. (1991)

1.1 1.1 - 3.4 1.8 0.7 – 2.9

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Seagrasses enhance nutrient concentrations in the sediments, both in particulate and dissolved forms, by an increased input of organic matter from plant biomass and imported material. Roots and rhizomes typically decay in situ, and some aboveground material may remain in the system rather than be exported by water flow. However, as with salt marsh ecosystems, it appears that much, if not most, of the organic matter buried and decomposed in seagrass sediments is imported (Pedersen et al. 1997; Boschker et al. 2000; Gacia et al. 2002; see Chapter 8). Fine, organic-rich particles settle out as the water flow is slowed by the seagrass canopy, and the roots and rhizomes also help to stabilize the sediments and prevent resuspension. As a result, organic matter and nutrient concentrations are typically higher in vegetated than in unvegetated sediments. Nutrient accumulation may occur in seagrass meadows over the time scale of years, but this source-sink role may in part depend on the trophic status of the system. Pedersen et al. (1997) found that colonization and stand development of the seagrass Cymodocea nodosa in Mediterranean carbonate sediments resulted in a net accumulation of N, but not P, over a 5 year period, with about half of the accumulated N accounted for by living and dead plant material. In C. nodosa meadows where production was nutrient limited, as is the case during stand development, Perez et al. (1997) observed a net accumulation of N and P in slowly-degrading belowground detritus, yet in eutrophic waters where belowground production was reduced and decay rates of leaves were high, the seagrass meadow becomes a source of nutrients.

2.2 Benthic microalgae The importance of microphytobenthic production in shallow coastal areas is well documented (Underwood and Kromkamp 1999), as are the effects of microalgal production on N cycling. Photosynthesis and respiration by benthic microalgae in the top few mm of the sediments cause large diel variations in oxygen concentrations and penetration depth, dissolved inorganic carbon (DIC) concentrations and pH (see Chapter 10). In addition, assimilation of porewater ammonium and nitrate influences the concentrations and depth distributions of porewater N. The net effect of these changes may be diel variations in nitrification and denitrification activity (Risgaard-Petersen et al 1994, Lorenzen et al. 1998, An and Joye 2001, see Chapter 10), with higher rates of coupled nitrificationdenitrification in light due to stimulated nitrification activity from photosynthetic oxygen release. However, this is not always the case (Risgaard-Petersen 2003 and references therein), and light stimulation may only occur when there is sufficient N to prevent N limitation (Rysgaard et al. 1995; An and Joye 2001). Recent field data from 18 European estuaries showed that net autotrophic sediments colonized by microalgae had lower rates of coupled nitrification-denitrification than net heterotrophic sediments (Risgaard-Petersen 2003). This was also the case for shallow bays in Sweden, where denitrification rates were higher (20% of remineralized N) in net heterotrophic than in net autotrophic sediments (10% of remineralized N) (Sundbäck and Miles 2002). Benthic microalgal communities in both tropical and temperate sediments are generally dominated by diatoms and cyanobacteria. Nitrogen fixation rates in

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surface sediments, and on seagrass leaves, tend to be high relative to those in the water column, and to be stimulated by light (e.g., Capone and Taylor 1977; Capone et al. 1992). Significant autotrophic N fixation occurs even in seagrass-vegetated sediments where the leaf canopy shades the sediment surface, although areaintegrated rates generally are lower than heterotrophic N fixation (e.g., O’Neil and Capone 1989). Rates are generally higher in tropical carbonate sands than in temperate siliclastic sediments (e.g., Howarth et al. 1988; Capone et al. 1992). There is currently little evidence of diurnal variations in N fixation activity in benthic microalgal communities in submersed sediments, however, Currin et al. (1996) found a night-time peak in N fixation in filamentous cyanobacterial mats and a day-time peak in coccoid cyanobacterial mats in intertidal sediments. Among the primary producers of shallow-water areas, benthic microalgae appear to be the component whose quantitative role in nutrient cycling is least understood at the ecosystem level. Integrating the effect of benthic microalgae at the ecosystem scale is difficult. Although benthic microalgae are present and active throughout the year (Rizzo et al. 1996 and references therein), their relative quantitative importance for nutrient turnover depends on the presence, seasonal pattern and area coverage of other primary producers, as well as physical characteristics of the embayment. For example, McGlathery et al. (2001a) found that the contribution of benthic microalgae to carbon production varied from 4-99% during the year in a shallow lagoon. The influence of benthic microalgae on net autotrophy and N cycling in the sediments appears to be particularly important in cool microtidal areas (Sundbäck et al. 2000), where the growth of seagrasses and ephemeral floating macroalgae is confined to the warm season (May-September). In warmer systems, macroscopic primary producers may occur during most of the year, implying more continuous shading and nutrient competition, although this has been observed primarily in eutrophic systems (e.g., Sfriso et al. 1992; Viaroli et al. 1996). Studies performed in different types of shallow-water ecosystems, ranging from cool temperate (e.g. Rizzo et al. 1992, 1996; Cerco and Seitzinger 1997, Sundbäck and Miles 2000; Thornton et al. 2002) to warm temperate (Eyre and Ferguson 2002) to subtropical and tropical areas (Miyajima et al. 2001; Suzumura et al. 2002), suggest that benthic microalgae may be an important temporary sink for nutrients. Studies from temperate, relatively N poor areas, suggest that benthic microalgae may turn sediments into sinks of N for the majority of the year, but particularly in winter and spring (e.g., Sundbäck et al. 2000). In such sandy areas, N may be limiting for benthic microalgae (Nilsson et al. 1991). Also in P-limited oligotrophic subtropical carbonate sediments, microalgal activity has been found to enhance accumulation of combined N (Miyajima et al. 2001). 2.3 Macroalgae Bloom-forming macroalgae have become dominant in nutrient-enriched waters over at least the last two decades throughout temperate and tropical regions (e.g., Cambridge & McComb 1984; Sfriso et al. 1992; Viaroli et al. 1996; Valiela et al. 1997; Pihl et al. 1999). These “nuisance” algae are typically filamentous or sheetlike forms -- mainly chlorophytes (e.g., Ulva, Cladophora, Chaetomorpha) -- that can accumulate in extensive thick mats over the sediment surface or in the water

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column. Variations in light availability within the mats, and in nutrient, oxygen, and pH conditions linked to algal photosynthesis, respiration, and decomposition, influence nutrient cycling at the sediment-water interface, and perhaps even deep into the sediments. The presence of dense macroalgal mats can move the location of the oxic-anoxic interface up from the sediments into the macroalgal mat since only the upper few cm of the mat may be within the photic zone (Krause-Jensen et al. 1999; Astill and Lavery 2001). Sediment nutrient cycling is enhanced by the presence of macroalgae, presumably due to the input of organic matter (Trimmer et al. 2000; Tyler et al. 2003). Increased mineralization rates cause a build up of ammonium deep within the mat where light does not penetrate (McGlathery et al. 1997; Astill and Lavery 2001), and may affect porewater ammonium as deep as 13 cm in the sediment (Burton et al, unpublished data). This recycling of organic-bound nitrogen may provide an important nitrogen source to sustain macroalgal production, particularly in areas where tidal exchange is limited (McGlathery et al. 1997; Trimmer et al. 2000). Increased ammonium concentrations at depth in sediments may also be partly related to local extinction of bioturbating infauna caused by anoxia (Hansen and Kristensen 1997). Where bioturbating infauna is present, increased oxygen diffusion into the sediments may stimulate nitrification rates (Rysgaard et al. 1995; Hansen and Kristensen 1997). Compared to benthic microalgae and seagrasses, there is little information on the influence of macroalgae on denitrification rates. In studies performed in harbors on the south coast of England, Trimmer et al. (2000) found that both direct denitrification, using nitrate supplied from the water column and denitrification couple to sedimentary nitrification were low in sediments underlying macroalgal mats. High free sulfide concentrations occur in organic-rich sediments underlying macroalgal accumulations (Viaroli et al. 1996), and these conditions may inhibit nitrification (Henriksen and Kemp 1998; Sloth et al. 1995) and partially account for the low denitrification rates. In a laboratory experiment, Krause-Jensen et al. (1999) showed that the net effect of dense macroalgal mats was to move the zone of denitrification from the sediments up into the mat, but not to influence the rates significantly. Denitrification rates peaked at the oxic-anoxic interface in the middle of the mat. Presumably the same regulating mechanisms would apply over the larger spatial scale in a macroalgal mat as have been determined in benthic microalgal communities in the sediment. Macroalgal photosynthesis and respiration induce diurnal variations in oxygen penetration depth within the upper productive layers of the mat, which in turn, may cause diurnal variations in nitrification activity and in coupled nitrification-denitrification. Mineralization of algal-bound nutrients in the lower layers of the mat where there is insufficient light for photosynthesis creates a diffusion gradient of ammonium up into the oxidized layers. Like benthic microalgae, it is expected that macroalgae will compete with bacteria for ammonium and nitrate. It appears that macroalgae do not stimulate denitrification, however, more work is needed to address the controls on denitrification in sediments underlying benthic macroalgal communities and within the mats themselves. At the ecosystem scale, macroalgae can store significant quantities of nutrients. In highly enriched waters, it is not unusual for macroalgal populations to attain peak biomass of over 0.5 kg dry wt m-2 and for canopy heights to exceed 0.5 m. In Waquoit Bay, Massachusetts, nitrogen stored in peak macroalgal biomass was of the

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same magnitude as the annual nitrogen load from the watershed (Valiela et al. 1997). Many studies have shown sediments overlain with macroalgae to be a net sink for inorganic nutrients (e.g., Krause-Jensen et al. 1996; Tyler et al. 2001). In Hog Island Bay, Virginia, macroalgae were a net sink for DIN throughout the year, and at the same time a net source for organic N (Tyler et al. 2003); urea and dissolved free amino acids (DFAA) were sequestered by macroalgae, but up to 22% of the total nitrogen uptake was released as dissolved combined amino acids (DCAA). On a seasonal time scale, it is common for macroalgae in eutrophic waters to switch from being a net sink early in the growing season, to a net source of nutrients in late summer when productivity declines due to self-shading within the mat and high temperatures increase respiration. High rates of macroalgal metabolism can lead to periods of oxygen supersaturation or, following senescence, oxygen depletion throughout the entire water column (e.g., Sfriso et al. 1992, Valiela et al. 1992, Boynton et al. 1996). Anoxic events in macroalgal-dominated systems tend to be episodic and to occur during the warm summer months when macroalgal production declines. Such “boom and bust’ cycles of high production and senescence are typical of dense macroalgal populations in eutrophic waters (e.g., McComb and Humphries 1992, Sfriso et al. 1992, Valiela et al. 1997). In Hog Island Bay, Virginia, Tyler et al. (2001) reported dissolved inorganic and organic N release rates following the collapse of a macroalgal bloom (primarily composed of Gracilaria tikvahiae and Ulva lactuca) that were sufficient to result in complete mineralization of the macroalgal biomass (up to 650 gdw m-2) within approximately 13 days. Release of plant-bound nutrients following these “dystrophic” events may stimulate phytoplankton and bacterial metabolism in the water column (Valiela et al. 1997; McGlathery et al. 2001a; Lunsford 2002). 3. SEDIMENT - WATER COLUMN FLUXES Nutrient assimilation by benthic primary producers can reduce the efflux of remineralized nutrients from the sediments to the overlying water, effectively decoupling nutrient turnover within the sediments from water column processes. During times of the year when primary producer populations are most active, this “filter” effect completely intercepts nutrient fluxes across the sediment-water interface and thereby reduces nutrient availability for phytoplankton and bacteria, as well as for floating macroalgal mats, in the overlying water. If the sediment nutrient sources are insufficient to meet growth demands, there also may be a downward flux of nutrients from the water column to the benthic community. It is generally assumed that in littoral zone ecosystems biogeochemical processes occurring in the benthos are closely coupled with those in the pelagic zone. For example, primary production in the water column supports heterotrophic metabolism in underlying sediments. In turn, organic matter remineralization in sediments releases nutrients that support primary production in overlying water. In situations where benthic – pelagic coupling is strong it should be possible to calculate ammonium fluxes out of sediments based on measured rates of net ecosystem metabolism (NEM; dissolved oxygen uptake or DIC release) (Hopkinson et al.

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2001). In oxic sediments where nitrification is likely to play an important role in transformation of ammonium, it may be the total dissolved inorganic nitrogen (DIN) flux that is more closely related to the stoichiometry of NEM, whereas in anoxic sediments where denitrification is important the estimated DIN flux will not equal that predicted by NEM stoichiometry. Calculation of denitrification has often been based upon this “missing” DIN. Using 15N isotope dilution techniques, it is possible to measure both gross N mineralization and consumption of ammonium in sediments amended with 15NH4+, thereby allowing determination of net mineralization rates (Anderson et al, 2003). Net mineralization may also be measured during incubation of unamended sediment/water cores. Where gross and net mineralization and sediment-water column exchanges of DIN have been measured simultaneously in autotrophic sediments, gross rates of ammonium production generally exceed net rates by a large margin and net production of ammonium in the sediments is often not reflected in fluxes to the water column (e.g., Kristensen et al. 2000, Anderson et al. 2003). Benthic microalgal assimilation is likely to be largely responsible for this (Anderson et al. 2003), although sediment bacteria may also play a role in N immobilization depending on the C/N of the sediment organic matter undergoing decomposition. In sediments with large amounts of macrophyte detritus with C/N as high as 45, and with the short turnover times typically observed for the ammonium pool (<1 d; Kristensen et al. 2000; Anderson et al. 2003), bacterial immobilization of N would be expected. In general, benthic microalgal uptake probably represents a major fate of mineralized N compared to coupled nitrification – denitrification in net autotrophic sediments. This has been observed in Hog Island Bay, Virginia, sediments (Anderson et al. unpublished), in southern England sediments dominated by macroalgae (Trimmer et al. 2000), in the Tagus Estuary, Portugal (Cabrita and Brotas 2000), and at sites along the west coast of Sweden (Sundbäck et al. 2000). The DIN fluxes observed in these littoral zone systems differ markedly from those observed in sublittoral zone systems, which typically exhibit release of ammonium or ammonium plus nitrate, closer to values predicted by NEM (e.g., Giblin et al. 1997; Burdige and Zheng 1998; Hopkinson et al. 1999). The idea that benthic microalgae may significantly modify the flux of nutrients between sediment and the overlying water column was put forward more than 20 years ago (Henriksen et al. 1980). As the density of benthic microalgae – and other microbiota– in the top few mm of sediment is orders of magnitude higher than in the water column, such a filter function should be expected, but had been previously overlooked because flux measurements were traditionally made only in the dark. More recently, several studies have shown that the presence of benthic microalgae can indeed significantly decrease the net efflux of N, P and silica, or even change the direction from efflux to influx (e.g., Reay et al.1995, Cerco and Seitzinger 1997; Sundbäck et al. 2000; Suzumura et al. 2002; Anderson et al. 2003; Tyler et al. 2003; see Table 2). This filter effect is mediated by both microalgal assimilation and photosynthetic oxygenation of the sediment surface, which in turn creates dynamic chemical and biological gradients. High-resolution microsensors have made it possible to study the influence of photoautotrophic activity on N turnover in more detail (see Chapter 10).

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The majority of studies on the regulatory effect of benthic microalgae on sedimentwater nutrient fluxes are from light/dark incubations of flooded sediment, mimicking subtidal conditions. Such incubations do not account for the effect of emersion on fluxes in intertidal sediments that occur in many shallow estuaries (e.g., Falcao and Vale 1995). However, Thornton et al. (1999) showed that illumination during emersion reduced fluxes of ammonium to the overlying water after immersion and concluded that the effect of the illumination and emersion periods should be considered when calculating nutrient budgets for intertidal, cohesive sediments. Like benthic microalgae, accumulations of benthic macroalgae can function temporarily as a filter by intercepting dissolved nutrient fluxes from the sediments to the overlying water (Thybo-Christensen et al. 1993; McGlathery et al. 1997; Tyler et al. 2001). The effectiveness of this interception depends on the biomass and productivity of the macroalgae as well as on the stability of macroalgal accumulations. Unattached macroalgal mats tend to be patchy and unstable; however, Astill and Lavery (2001) showed that nutrient and oxygen gradients developed quickly as macroalgae accumulated (within 24 hr), suggesting that this filter function occurs even in dynamic environments. Table 2. Ammonium and nitrate fluxes from sediments in two temperate locations, illustrating the effect of benthic microalgae on the magnitude and direction of nutrient fluxes. Values represent annual means (± se) from monthly – bimonthly measurements. Fluxes from Hog Island Bay sediments show the influence of net autotrophy (data from Tyler et al. 2003), while those from the Swedish coastal bays show the influence of sediment type (data from Sundbäck et al. 2000).

Hog Island Bay, Virginia Autotrophic sediments Heterotrophic sediments Swedish west coast Sandy sediments – Vallda Silty Rötången

sediments

-

Ammonium Flux µmol N m-2 d-1

Nitrate Flux µmol N m-2 d-1

-327 ± 104 -115 ± 36 418 ± 90 384 ± 126

-78 ± 29 -80 ± 27 9 ± 47 6 ± 13

108 ± 113 145 ± 72 337 ± 124

-87 ± 64 -396 ± 325 -46 ± 130

306 ± 110

-161 ± 149

Most studies have focused on macroalgal interception of dissolved inorganic nutrient fluxes (e.g., Thybo-Christesen et al. 1993; McGlathery et al. 1997); however, macroalgae also may be important in regulating the flux of dissolved organic nutrients across the sediment-water interface. Tyler et al. (in press) showed that the chlorophyte Ulva lactuca prevented the flux of urea and DFAA from the

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sediments to the overlying water. Sediments are considered to be a significant source of nutrients to supporting production of macroalgae, since nutrient concentrations are typically one to several orders of magnitude higher in sediment porewaters than in the water column, even in eutrophic waters. This has been shown in Hog Island Bay, Virginia where the efflux of DIN and urea from sediments underlying macroalgal mats was sufficient to meet 27-75% of the macroalgal N demand (Tyler et al. 2003). Likewise, Stimson and Larned (2000) determined that the N efflux from the sediments of a subtropical bay was essential to meet the growth demand of the dominant macroalgae and was at least in part responsible for the persistence of macroalgae in an otherwise oligotrophic environment. Sediment nutrient efflux is typically consumed in the bottom layers of dense macroalgal mats, and diffusion from the overlying water provides a necessary nitrogen supply for the light-saturated algae in the upper layers (McGlathery et al. 1997). Factors that reduce macroalgal productivity, such as decreased water clarity from suspended sediments or high phytoplankton biomass, decreased insolation, or selfshading within the lower layers of dense mats during summer also decrease macroalgal uptake and reduce the filter effect (Sfriso et al. 1992; D’Avanzo and Kremer 1994; McGlathery et al. 1997). Often one, or a combination of these factors, leads to the late-summer collapse of the macroalgal population, and the release of plant-bound nutrients to the water column. Macroalgal mats that are lightlimited may show a diurnal pattern of nutrient release, with nutrients diffusing from the sediments through the mat to the water column in the dark (McGlathery et al. 1997). This occurs because algae growing in low-light conditions are typically Nsaturated and lack carbon reserves; as such, they are dependent on recent photosynthate to build amino acids. This close coupling between nutrient uptake and photosynthesis is also characteristic of N-replete phytoplankton (Turpin 1991 and references therein). There have been fewer studies of mechanisms by which seagrass communities influence benthic dissolved nutrient fluxes. Like benthic micro- and macro-algae, one would expect assimilatory nutrient uptake by seagrasses to reduce the flux of nutrients across the sediment-water interface. Risgaard-Petersen & Ottosen (2000) in a study performed in Denmark showed a seasonal pattern of DIN exchange between eelgrass (Z. marina) communities and the water column similar to that which has been shown for sediments with benthic microalgal and macroalgal communities. There was an influx of N to eelgrass-vegetated sediments in the spring and summer when plant metabolism was highest, and a smaller influx in the colder winter months. In the fall when plant N demand was low, but decomposition rates were relatively high, nutrients were released to the water column. In studies performed in Virginia, where seagrass growth is high in spring and fall and decomposition rates high in July - August, seasonal patterns of nutrient exchange were slightly different from those in Denmark in that there was a large efflux of DIN (primarily ammonium) to the water column in June and August and a negligible efflux in April and October (Anderson and Moore, unpublished data). Both studies illustrate the important influence of assimilatory nutrient uptake by seagrasses on the direction of nutrient fluxes at the sediment-water interface. In tropical carbonate sediments, P fluxes are typically low (nM), in part because the P is bound in the solid phase (Jensen et al. 1998) and in part because turnover rates in sediment porewaters are high to meet seagrass demand (McGlathery et al. 2001b).

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Other primary producers associated with seagrass communities also can play a key role in regulating the magnitude and direction of the nutrient flux. Miyajima et al. (2001) found an influx of N in both bare and seagrass-vegetated carbonate sediments, and attributed this at least in part to high benthic microalgal activity. The presence of epiphytes on seagrass blades also may be important, as Eyre and Ferguson (2002) found a net influx of N in warm-temperate Australian seagrass communities during light incubations only when epiphytes were included. Likewise, it has been shown that epiphytes can remove more ammonium and nitrate from the water column than the seagrass Ruppia megacarpa even though their biomass was significantly less (Dudley et al. 2001). 4. COMPETITION BETWEEN PRIMARY PRODUCERS There is ample evidence to support the general hypothesis that phytoplankton and benthic primary producers compete for resources (nutrients, light) in shallow coastal waters (e.g., Sand-Jensen and Borum 1991; Fong et al. 1993), and that the outcome of this competition is unstable. Likewise, competition between benthic primary producers occurs, especially in systems that receive high external nutrient loads (Valiela et al. 1997 and references therein). Shifts in primary producer dominance that result from this competition have an impact on the rates and pathways of nutrient transformations within these systems. Most studies indicate that at certain times of the year the reduction, or elimination, of the internal supply of nutrients to the water column from sediment regeneration that is attributed to benthic microalgal uptake decreases nutrient availability for phytoplankton (e.g., Sundbäck et al. 2000; Suzumura et al. 2002). However, Cerco and Seitzinger (1997) have suggested that benthic microalgae actually enhance phytoplankton production on an annual time scale. By sequestering mineralized nutrients in the winter and spring, these authors suggest that benthic microalgae may function as a temporary storage pool for regenerated nutrients; nutrients are released in the summer when the benthic microalgae become light-limited by increased phytoplankton biomass in the water column. The net effect is to extend the summer phytoplankton bloom beyond the time when it would normally become nutrientlimited. It has been suggested that uptake of mineralized nutrients in the sediment by benthic microalgae also may inhibit, or at least defer, the initiation of benthic macroalgal blooms in the spring. A study in shallow sheltered embayments in Sweden suggested that benthic microalgae could successfully compete with ephemeral green algae for the sediment pool of regenerated nutrients during the period that is critical for the onset of the macroalgal growth (Sundbäck and Miles 2002). On a daily basis, the benthic microalgae decreased the efflux of inorganic N by 30 – 100%, P by 70 100% and silica by 10–95%. Dense macroalgal mats, in turn, decrease light availability at the sediment surface (90-100% for macroalgal biomass 100-500 gdw m-2; Krause-Jensen et al. 1996; Astill and Lavery 2001), and sediments underlying stable algal accumulations tend to be heterotrophic (Tyler et al. 2003). Surprisingly, there have been few direct studies of the effect of mat-forming macroalgae on benthic microalgal productivity (Sundbäck et al. 1990, 1996a). Persistent, dense macroalgal mats would be expected

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to decrease production of sediment microalgal communities by reducing light levels and by creating anoxic conditions at the sediment surface. In Hog Island Bay, Virginia, benthic microalgal production increased in importance following a midsummer macroalgal decline, suggesting that competition for light limited benthic microalgal production when macroalgal densities were high (McGlathery et al. 2001a). A more recent study in this lagoon found an inverse relationship between macroalgal biomass and both sediment chlorophyll a and sediment dissolved oxygen production over an annual cycle (Tyler et al. 2003), indicating decreased benthic microalgal activity in sediments underlying macroalgal accumulations. At lower densities, or when macroalgae are floating in the water column and thereby do not create intense shading and/or bottom water anoxia, sediment microalgal production would not be as strongly influenced (Sundbäck et al. 1996a). By sequestering and storing sediment-derived nutrients, dense macroalgal mats also may outcompete phytoplankton during the spring and early summer when macroalgal growth rates are highest (Viaroli et al. 1992; McGlathery et al. 1997; Valiela et al. 1997). This probably accounts for the high water clarity in many shallow macroalgal-dominated systems, despite high external nutrient loads (Valiela et al. 1997). Macroalgae also can sustain maximum growth rates for longer periods of time than phytoplankton because their nutrient storage capacity obviates the need for a constant nutrient supply (Fong et al. 1993; Pedersen and Borum 1996). Stored nutrients in phytoplankton can only sustain maximum growth for a day or less, whereas growth of bloom-forming macroalgae such as Ulva, Cladophora and Chaetomorpha can be sustained by stored nutrients for several days, or more, depending on growth conditions (light and temperature) (Borum 1996). However, the outcome of competition also can be reversed when phytoplankton blooms begin to shade macroalgal communities. This usually happens when macroalgal metabolism slows down in late summer and the macroalgae become a less effective filter for the effluxed sediment nutrients. As a consequence, nutrients diffuse into the water column and stimulate phytoplankton growth, which, in turn, further reduces macroalgal productivity through shading. Valiela et al. (1997) observed interannual variation in Waquoit Bay, Massachusetts, that was consistent with this model. In the most N-enriched subestuary, phytoplankton abundance and activity increased at the same time that macroalgal biomass declined. Many studies have shown that low macroalgal biomass coincides with peak phytoplankton biomass (e.g., Sfriso et al. 1992; Viaroli et al. 1992). Typically, these moderately-enriched systems are unstable, and are characterized by shifts in dominance between benthic macroalgae and phytoplankton. The same scenario has been suggested for systems in which microalgae dominate the benthic community and where shading by phytoplankton blooms can cause a temporary collapse of benthic microalgal productivity (Blanchard and Montagna 1995; Cerco and Seitzinger 1997). The competition between rooted macrophytes and benthic microalgae has not been studied specifically. One might expect low benthic microalgal biomass and production in densely-vegetated seagrass systems due to shading of the sediment by the leaf canopy, yet many studies have shown that benthic microalgae can contribute significantly to the total productivity of seagrass beds (e.g., Murray and Wetzel 1987; Moncrieff et al. 1992). The same has been observed in Spartina alterniflora vegetated intertidal sediments (Anderson et al. 1997). In some seagrass systems, benthic microalgal biomass and primary production has been found to be as high as

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in bare sediments (e.g., Moncreiff et al. 1992, Pollard and Kogure 1993, Kemp and Cornwell 2001), or even higher (Miyajima et al. 2001). The contribution of this production to nutrient retention and turnover in seagrass beds is unclear. Hansen et al. (2000) found that uptake by benthic microalgae in a temperate seagrass bed was the major sink for DIN in the spring. Likewise, Kemp and Cornwell (2001) found that benthic microalgal N uptake was as important as that by seagrasses in Florida Bay sediments in August. It is likely that the interaction between benthic microalgae and seagrass varies seasonally with changes in light and nutrient availability as the seagrass canopy develops. In nutrient-enriched systems, macroalgal blooms can outcompete seagrasses, although the outcome of this competition is reversible if macroalgal biomass decreases. Hauxwell et al. (2000) showed that dense macroalgal canopies caused a decrease in recruitment of new shoots and rates of leaf appearance in existing shoots of the eelgrass Z. marina in an enriched subestuary of Waquoit Bay, Massachusetts. Using model calculations to estimate light attenuation by the different autotrophs (phytoplankton, epiphytes, macroalgae), they confirmed that the primary cause of eelgrass loss was light reduction by the macroalgal canopy, in particular for newly recruiting shoots. One might expect that some tropical seagrasses, such as Thalassia testudinum, that have a proportionately greater allocation to belowground biomass (and carbon reserves) may be more resilient to ephemeral shading by macroalgal canopies. Increased ammonium concentrations within macroalgal mats (>25 µM) also may be toxic to seagrass (van Katwijk et al. 1997), and again, this effect is likely to be most important for newly recruiting shoots that exist entirely within the macroalgal canopy. A decrease in redox and an increase in sediment sulfide concentrations resulting from decomposition in the anoxic, organic-rich sediments and decaying macroalgal layer also may reduce seagrass photosynthesis (Goodman et al. 1995; but see Terrados et al. 1999). Decreased photosynthetic oxygen production at all light levels also decreases the potential for oxygen translocation and release to the rhizosphere, and creates a positive feedback that reduces sulfide oxidation around the roots and further elevates sediment sulfide levels, which decrease nutrient uptake and plant energy status (Pregnall et al. 1984).

5. THE ROLE OF PRIMARY PRODUCERS IN THE ESTUARINE ‘FILTER’ – FATE AND RETENTION OF ASSIMILATED NUTRIENTS Nutrient assimilation and temporary retention is probably the most important process regulating the source-sink role of benthic communities. In the short term (days-months), uptake and temporary assimilation influence the fluxes of nutrients between the sediment and water column. Carbon and nitrogen that are fixed may be rapidly re-released to the water column as DOC or DON. The DOM released by benthic autotrophs may be processed by the microbial loop, mineralized, or transferred to higher trophic levels. Abiotic processes such as humification are also operative in sediments. In the long term, while most of the biomass of benthic primary producers is decomposed, grazed, or exported from the system, some is retained for possible burial (Cebrian 2002; see Chapter 8). Seagrass-dominated systems have higher rates of permanent burial in the sediments because of the

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inherently greater refractory content of seagrass tissue and the presence of significant quantities of belowground material (roots and rhizomes) that are less likely to be exported (Buchsbaum et al. 1991; Enriquez et al. 1993; Duarte et al. 1996; Klap et al. 2000; see Chapter 8). However, decay rates of seagrass leaves can be high, and similar in magnitude as those for macroalgae (Buchsbaum et al. 1991). Risgaard-Petersen and Ottosen (2000) found for Z. marina that the pool of tissuebound N did not accumulate in the leaves on an annual time scale. Nutrients that were temporarily retained in leaf tissue were lost when the leaves were detached and transported from the systems, or if retained, were nearly completely mineralized (95-98%) within a year. Similarly, most of the fixed C and N in macroalgae is released during senescence over a period of several months (Buchsbaum et al. 1991). Benthic microalgae also generally have high growth rates and low C/N contents; thus a rapid turnover of N in benthic microalgae would be expected. Calculating assimilation is difficult because benthic primary producers fix more carbon and nutrients than are needed to meet the metabolic needs for growth and some of the fixed C and N (and probably P) is leaked from living tissue. For example, the release of up to 22% of DIN uptake by macroalgae as DON compounds, means that calculations of macroalgal N assimilation based on growth rates and tissue C:N content are severe underestimates (Tyler et al. 2003). Similarly, calculations of benthic microalgal nitrogen demand are subject to potential errors of either over- or underestimation. Calculations of N demand may be based either on measurements of net primary production (NPP) (Sundbäck and Miles 2000) or gross primary production (GPP) (Anderson et al. 2003). If NPP estimates are based on DO or DIC fluxes measured in sediment cores, sediment community respiration will result in an underestimation of microalgal NPP and, therefore, N demand. On the other hand, estimation of N demand based on GPP may result in overestimation since GPP does not take into account exudation of DOM, which may be a large component (up to 75%, see below) of the carbon fixed. Use of 13C and 15N tracers would provide more realistic measures of benthic autotroph N demand; however, a complete understanding would require both inorganic and organic tracers since benthic primary producers are known to take up DON (e.g., Nilsson and Sundbäck 1996; Tyler et al. 2003). There have been some attempts to compare the importance of benthic microalgal and macroalgal assimilation of nutrients in coastal embayments, in which the assimilation rate of N by benthic microalgae has been found to be similar to, or even exceed, that of macroalgae (e.g., McGlathery et al. 2001a; Sundbäck and Miles 2002). Carbon and nitrogen that are not used for growth may be stored in tissue, exuded as attached extracellular polymeric substances (EPS), or released to the surrounding water as colloidal organic matter (COM). For example, it has been observed that up to 75% of carbon fixed by benthic diatoms may be exuded as EPS or COM (Goto et al. 1999; Smith and Underwood 2000; Wolfstein et al. 2002). Wolfstein et al (2002) found that the percentage of fixed C excreted appeared to vary with irradiance (decreased excretion at higher irradiance) and with bacterial activity. In an in situ pulse – chase 13C labeling experiment Middelburg et al. (2000) noted rapid transfer of 13C from benthic microalgal into bacterial biomass. They suggest that exudation of EPS may be a medium for this exchange. There are also reports of DOC leakage from macroalgal fronds which may be as high as 39% of gross production (Velimirov 1986). In contrast, there appears to be more limited release of organic

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compounds from living seagrass tissues. Of the carbon that is fixed during photosynthesis, only 1-2% is released by the leaves (e.g., Velimirov 1986; Moriarty et al. 1986) and up to 11% by the roots and rhizomes (Moriarty et al. 1986). Nitrogen retention in benthic primary producer communities depends in part on the leakage of DON from live tissues, yet there is considerably less known about this process than for DOC. To our knowledge no studies have demonstrated exudation of DON by benthic microalgae, although DON release has been observed for pelagic cyanobacteria such as Synechococcus spp. (Bronk, 1999) and for the nitrogen fixing Trichodesmium spp. (Glibert and Bronk, 1994). Rates of DON release have been observed to vary with light availability and nutrient status, depth of the water column, DIN species composition and concentration, and season (Ward and Bronk, 2001). The composition of DON released includes DFAA, DCAA, dissolved primary amines, and urea (Berman and Bronk, in press). DON may be released by passive release, excretion, sloppy feeding during grazing by copepods or direct release by microzooplankton, and viral lysis. Macroalgae also may release significant quantities of DON as DCAA; Tyler et al. (in press) found benthic fluxes of DCAA that were nearly 8-fold higher in the presence of macroalgae, and were higher in the light, suggesting a photosynthetically-driven process (Tyler et al. 2003). We are unaware of any studies on DON release from living seagrasses. In phytoplankton-based systems, release of DOC or DON can stimulate water column bacteria and higher trophic levels (Azam et al. 1983). It is likely that this also occurs in benthic primary producer communities, although it is more difficult to show this link. In Hog Island Bay, Virginia, highest rates of water column metabolism of DOC and DON were observed in August, the period when the macroalgal bloom declines, suggesting that the labile DOM was derived from macroalgal detritus (Lunsford, 2002). Rates of DOM metabolism in a small Massachusetts phytoplankton-dominated estuary bounded by both fresh and saltwater marshes were only 25% of those observed in the Virginia lagoon. The DOM in the Massachusetts estuary was primarily allochthonous in origin, whereas that in the Virginia coastal lagoon was mainly autochthonous. Studies such as this suggest that bacterial activity in the water column can be an important link between sediment detrital material and higher trophic levels, and may play an important role in nutrient transformations in shallow coastal systems. The retention time of tissue-bound nutrients also depends on the transformations once assimilated, which can take three pathways in the benthic food web: 1) through the macrofaunal grazing/deposit feeding chain (Asmus and Asmus 1985; Duarte and Cebrian 1996; Herman et al. 2000), 2) through the ‘small food chain’ consisting of micro- and meiofauna (Kuipers et al. 1981; Sundbäck et al. 1996b and references therein), or 3) burial or export (Admiraal et al. 1984; Boschker et al. 1999; Middelburg et al. 2000). Since much of the organic matter is released as DOM by live and dead material, processing of detritus by grazers and bacteria only accounts for a portion of the plant organic matter. For benthic microalgae, an in situ 13 C-labelling experiment showed that assimilated C was transferred within a few days through the heterotrophic components in an order bacteria > macrofauna > meiofauna (Middelburg et al. 2000). Sundbäck et al. (1996b) found by in situ duallabeling (14C-HCO3, 3H-thymidine; Montagna 1984) of sediment, that meiofauna grazed between 2 and 12% of the microalgal biomass per day and had a significant impact on microalgal biomass in spring and autumn, but not in summer.

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The generally accepted conceptual models of eutrophication (Sand-Jensen and Borum 1991; Duarte 1995; Valiela et al. 1997) predict that with increasing nutrient loading, benthic primary production will decrease; sediments will become increasingly heterotrophic, resulting in increased efflux of nutrients from the sediment to the water column. Ultimately, this will cause a positive feedback accelerating eutrophication through internal loading of nutrients. Some empirical evidence supporting this hypothesized series of events has been found (e.g., Rizzo et al. 1992; Meyercordt and Meyer-Reil 1999; Eyre and Ferguson 2002). Initially, both seagrasses and benthic microalgae will be stimulated by increased nutrient levels, particularly in sandy sediments (Nilsson et al. 1991; Flothman and Werner 1992). However, at the same time the growth of filamentous macroalgae and epiphytes will be stimulated (e.g., Nilsson et al. 1991; Havens et al. 2001), with increased shading of the benthos. Systems dominated by fast-growing macroalgae and phytoplankton are characterized by short-lived (days-months) bloom events and rapid turnover of nutrients bound in plant biomass. Whether and when planktonic or macroalgal communities outcompete benthic microalgae and seagrasses by shading will depend on the duration and extent of bloom events. We hypothesise that, with proceeding eutrophication and light attenuation from high phytoplankton biomass, the negative impact on the microphytobenthic community will be more gradual and slower than for benthic macroscopic primary producers, including seagrasses and macroalgae. Thus, a partial beneficial “buffering” effect of benthic microalgae on shallow sediment systems may persist even in more heavily eutrophied systems, as has been shown by Krom et al. (1991) for a hypertrophic fish pond. This scenario is based on the assumption that benthic microalgal communities possess, due to high diversity and functional redundancy, a certain degree of plasticity, increasing the overall resilience of shallow-water sediment systems after pelagic bloom events. Benthic diatoms can survive periods of only a few % of incident light, or even darkness, and high sulphide levels, and can rapidly resume photosynthesis when exposed to light or after an anoxic event (Admiraal et al. 1984; Kennett and Hargraves 1985; Sundbäck and Granéli 1988; Sundbäck et al. 1990). Similarly, upward migrating benthic diatoms were found to rapidly restore the oxygenation of the sediment surface after a simulated sedimentation event (Wulff et al. 1997; Underwood and Paterson 1993). This scenario, with benthic microalgae surviving despite deteriorating conditions, may apply particularly to cool microtidal waters, where macroalgal bloom events last only a few months (e.g. Pihl et al. 1999; Dalsgaard in press; McGlathery et al. 2001a), leaving the rest of the year open to benthic microalgal primary production. In warm, eutrophic microtidal systems, benthic microalgae can be outcompeted by shading and also by dystrophic events when the macroalgal blooms eventually collapse (Viaroli et al. 1996). The efflux of N and P from the sediments will increase as eutrophication proceeds due to prolonged shading of the benthos, as suggested by experiments at varying light intensities (Rizzo et al. 1992; Sundbäck and Granéli 1988; Cerco & Setizinger 1997). This will further stimulate pelagic primary productivity, while the sediment system will become more heterotrophic. Several studies have specifically pointed

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out that the autotrophy/heterotrophy balance is a good indicator of nutrient flux from sediments (e.g., McGlathery et al. 2001a; Eyre and Fergusson 2002; Laursen and Seitzinger 2002; Sundbäck et al. 2003; Tyler et al. 2003), with net autotrophic systems functioning as sinks and net heterotrophic systems as sources. As systems become increasingly enriched with organic matter, net heterotrophy will increase and the systems will become self-generating and demonstrate hysteresis; e.g., internal nutrient loading from the sediments will continue to sustain the growth of ephemeral macroalgal mats for some period of time after nutrient loads begin to decrease (Pihl et al. 1999; Sundbäck et al. 2003). However, measurements from some Danish shallow fjords suggest that the release rate of the sediment N pool can be significantly lower only a few years after the nutrient load to the water column has been decreased (Dalsgaard et al. 1999). When nutrient loads are decreased, a reversed scenario has been predicted, where benthic primary producers are favored because of better light conditions, thus accelerating the process of recovery (Christensen et al. 1998) by reoxygenation of the sediment, interception of the nutrient flux, and by increased retention of nutrients in the sediment. Seagrass-dominated systems will have both a longer retention time of nutrients temporarily bound in plant material and a greater permanent removal of nutrients by burial.

7. CONCLUSION In this chapter we have reviewed the effects of seagrass, macroalgal, and benthic microalgal communities on specific nitrogen and phosphorus cycling processes and on the ecosystem-level role of these communities as sources or sinks of nutrients. Photosynthetic and respiratory activities of benthic primary producers alter chemical conditions in the sediments, causing spatial and diel variations in DO, DIC, pH and nutrients that influence the rates and pathways of nutrient transformations. Uptake and retention of nutrients by plant biomass and the effects of primary producers on N losses from denitrification are key regulators of sediment-water column fluxes. As such, benthic primary producers play an important role regulating phytoplankton production in the overlying water column. However, nutrients bound in plant biomass turnover rapidly and probably make little difference on annual times scales to the overall nutrient retention within the system. It also appears that although there are examples of both stimulation and suppression of denitrification rates by benthic primary producer communities, there is not a consistent and significant increase of N losses by denitrification on annual time scales caused by benthic primary producers. This being the case, the key drivers to total estuarine nutrient retention of external inputs are likely dependent on: 1) the amount of burial of refractory compounds (including particulates imported to system and trapped as well as in situ production) and 2) the water residence time (i.e., to what extent recycled nutrients support primary and secondary production within the system before being exported). Both of these processes will be dependent on the character of the primary producer community. Coupling of ecological and physical processes is central to our understanding of how shallow coastal ecosystems function as a filter of watershed nutrient inputs, yet we know relatively little about the fate and turnover of

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Krause-Jensen D., K.J. McGlathery, S. Rysgaard and P.B. Christensen. (1996). Production within dense mats of the filamentous macroalga Chaetomorpha linum in relation to light and nutrient availability. Marine Ecology Progress Series 134: 207-216 Krause-Jensen D., P.B. Christensen and S. Rysgaard. (1999). Oxygen and nutrient dynamics within mats of the filamentous macroalga Chaeotomorpha linum. Estuaries 22: 31-38. Kristensen E., F.O. Andersen, N. Holmboe, M. Holmer and N. Thongtham. (2000). Carbon and nitrogen mineralization in sediments of the Bangrong mangrove area, Phuket, Thailand. Aquatic Microbial Ecology 22: 199-213 Krom M.D. (1991). Importance of benthic productivity in controlling the flux of dissolved inorganic nitrogen through the sediment-water interface in a hypertrophic marine ecosystem. Marine Ecology Progress Series 78: 163-172 Kuipers B.R., P.A.W.J. de Wilde and F. Creutzberg. (1981). Energy flow in a tidal flat ecosystem. Marine Ecology Progress Series 5: 215-221 Laursen A.E. and S.P. Seitzinger. (2002). Measurement of denitrification in rivers: an integrated, whole reach approach. Hydrobiologia 485: 67-81. Lomstein B.A., A.-G.U. Jensen, J.W. Hansen, I.B. Andreasen, L.S. Hansen, J. Berntsen and H. Kunzendorf. (1998). Budgets of sediment nitrogen and carbon cycling in the shallow water of Knebel Vig, Denmark. Aquatic Microbial Ecology 14: 69-80 Lorenzen J., L.H. Larsen, T. Kjaer and N.P. Revsbech. (1998). Biosensor determination of the microscale distribution of nitrate, nitrate assimilation, nitrification, and denitrification in a diatom-inhabited freshwater sediment. Applied and Environmental Microbiology 64: 3264-3269 Lunsford T. (2002). A comparison of the fate of dissolved organic matter in two coastal systems: Hog Island Bay, VA (USA) and Plum Island Sound, MA (USA). Ph.D. Thesis, Virginia Institute of Marine Science, College of William and Mary. McComb A.J. and R. Humphries. (1992). Loss of nutrients from catchments and their ecological impacts on the Peel-harvey estuarine system, Western Australia. Estuaries 15: 529-537. McGlathery K.J., D. Krause-Jensen, S. Rysgaard and P.B. Christensen. (1997). Patterns of ammonium uptake within dense mats of the filamentous macroalga Chaetomorpha linum. Aquatic Botany 59: 99-115 McGlathery K.J., N. Risgaard-Petersen and P.B. Christensen. (1998). Temporal and spatial variation in nitrogen fixation activity in the eelgrass (Zostera marina L.) rhizosphere. Marine Ecology Progress Series 168: 245-258 McGlathery K.J., I.C. Anderson and A.C. Tyler. (2001a). Magnitude and variability of benthic and pelagic metabolism in a temperate coastal lagoon. Marine Ecology Progress Series 216: 1-15 McGlathery K.J., P. Berg and R Marino. (2001b). Using porewater profiles to assess nutrient availability in seagrass-vegetated carbonate sediments. Biogeochemistry 56: 239-263 Meyercordt J. and L.A. Meyer-Reil. (1999). Primary production of benthic microalgae to two shallow coastal lagoons of different trophic status in the southern Baltic Sea. Marine Ecology Progress Series 178: 179-191 Middelburg J.J., C. Barranguet, H.T.S. Boschker, P.M.J. Herman, T. Moens and C.H.R. Heip. (2000). The fate of intertidal microphytobenthos carbon: an in situ 13C-labelling study. Limnology and Oceanography 45: 1224-1234

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Miyajima T., M. Suzumura, Y. Umezawa and I. Koike. (2001). Microbial nitrogen transformation in carbonate sediments of a coral-reef lagoon and associated seagrass beds. Marine Ecology Progress Series 217: 273-286 Moncreiff C.A., M.J. Sullivan, and A.E. Daehnick. (1992). Primary production in seagrass beds of Mississippi Sound: the contributions of seagrass, epiphytic algae, sand microflora, and phytoplankton. Marine Ecology Progress Series 87: 161-171 Montagna P.A.. (1984). In situ measurement of meiobenthic grazing rates on sediment bacteria and edaphic diatoms. Marine Ecology Progress Series 18: 119-130 Moriarty D.W., J.R. Iverson and P.C. Pollard. (1986). Exudation of organic carbon by the seagrass Halodule wrightii and its effect on bacterial growth in the sediment. Journal of Experimental Biology and Ecology 96: 115-126 Moriarty D.W. and M.J. O’Donohue. (1993). Nitrogen fixation in seagrass communities during summer in the Gulf of Carpentaria, Australia. Australian Journal of Marine and Freshwater Research 44: 117-125 Murray L. and R.L. Wetzel. (1987). Oxygen production and consumption associated with the major autotrophic components in two temperate seagrass communities. Marine Ecology Progress Series 38: 231-239 Nielsen L.P.. (1992). Denitrification in sediment determined from nitrogen isotope pairing. FEMS Microbiology Ecology. 86: 357-362 Nilsson C. and K. Sundbäck. (1996). Amino acid uptake in natural microphytobenthic assemblages studied by microautoradiography. Hydrobiologia 332: 119-129. Nilsson P., B. Jönsson, Lindström, I. Swanberg and K. Sundbäck. (1991). Response of a marine shallowwater sediment system to an increased load of inorganic nutrients. Marine Ecology Progress Series 71: 275-290 Nixon S.W. et al. (1996). The fate of nitrogen and phosphorus at the land-sea margin of the North Atlantic Ocean. Biogeochemistry 35: 141-180. National Research Council. (2000). Clean coastal waters: understanding and reducing the effects of nutrient pollution. National Academy Press, Washington, DC. O’Donohue M.J., D.J. Moriarty and I.C. MacRae. (1991). Nitrogen fixation in sediments and the rhizosphere of the seagrass Zostera capricorni. Microbial Ecology 22: 53-64 O’Neil J.M. and D.G. Capone. (1989). Nitrogenase activity in tropical carbonate marine sediments. Marine Ecology Progress Series 56: 145-156 Ottosen, L.D.M., N. Risgaard-Petersen and L.P. Nielsen. (1999). Direct and indirect measurements of nitrification and denitrification in the rhizosphere of aquatic macrophytes. Aquatic Microbial Ecology 19:81-91 Pedersen M.F. and J. Borum. (1996). Nutrient control of algal growth in estuarine waters. Nutrient limitation and the importance of nitrogen requirements and nitrogen storage among phytoplankton and species of macroalgae. Marine Ecology Progress Series 142: 261-272 Pedersen M.F., C.M. Duarte and J. Cebrian. (1997). Rates of changes in organic matter and nutrient stocks during seagrass Cymodocea nodosa colonization and stand development. Marine Ecology Progress Series 159: 29-36. Pelegri S.P., L.P. Nielsen and T.H. Blackburn. (1994). Denitrification in estuarine sediment stimulated by the irrigation activity of the amphipod Corophium volutator. Marine Ecology Progress Series 105:285-290.

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Perez M., M.A. Mateo, T. Alcoverro and J. Romero. (1997). Variability in detritus stocks in beds of the seagrass Cymodocea nodosa. Botanica Marina 44: 523-531. Pihl L., A. Svenson, P.O. Moksnes and H. Wennhage. (1999). Distribution of green algal mats throughout shallow soft bottoms of the Swedish Skagerrak archipelago in relation to nutrient sources and wave exposure. Journal of Sea Research 41: 281-294 Pollard P.C. and K. Kogure. (1993). The role of epiphytic and epibenthic algal productivity in a tropical seagrass, Syringodium isoetifolium (Aschers) Dandy, community. Australian Journal of Marine and Freshwater Research 44: 141-154 Pregnall A.M., R.D. Smith, T.A. Kursar and R.S. Alberte. (1984). Metabolic adaptations of Zostera marina (eelgrass) to diurnal periods of root anoxia. Marine Biology 83: 141-147 Reay W.G., D.L. Gallagher and G.M.J. Simmons. (1995). Sediment-water column oxygen and nutrient fluxes in nearshore environments of the lower Delmarva Peninsula, USA. Marine Ecology Progress Series 118: 215-227 Risgaard-Petersen N. (2003). Coupled nitrification-denitrification in autotrophic and heterotrophic estuarine sediments: On the influence of benthic microalgae. Limnology and Oceanography 48: 93105 Risgaard-Petersen N., S. Rysgaard, L.P. Nielsen and N.P. Revsbech. (1994). Diurnal variation of denitrification and nitrification in sediments colonized by benthic microphytes. Limnology and Oceanography 39: 573-579. Risgaard-Petersen N., T. Dalsgaard, S. Rysgaard, P.B. Christensen, J. Borum, K.J. McGlathery and L.P. Nielsen. (1998). Nitrogen balance of a temperate eelgrass Zostera marina bed. Marine Ecology Progress Series 174: 281-291 Risgaard-Petersen N. and L.D.M. Ottosen. (2000). Nitrogen cycling in two temperate Zostera marina beds: seasonal variation. Marine Ecology-Progress Series 198:93-107 Rizzo W.M., S.K. Dailey, G.J. Lackey, R.R. Christian, B.B. Berry, and R.L. Wetzel. (1996). A metabolism-based trophic index for comparing the ecological values of shallow-water sediment habitats. Estuaries. 19: 247-256 Rizzo W.M., G.L. Lackey, and R.R. Christian. (1992). Significance of euphotic, subtidal sediments to oxygen and nutrient cycling in a temperate estuary. Marine Ecology Progress Series 86: 51-61

Rozan T.F., M. Taillefert, R.E. Trouwborst, B.T. Glazer, S. Ma, J. Herszage, L.M. Valdes, K.S. Price and G.W. Luther III (2002). Iron-sulfur-phosphorus cycling in the sediments of a shallow coastal bay: Implications for sediment nutrient release and benthic macroalgal blooms. Limnology and Oceanography 47: 1346-1354. Rysgaard S., P.B. Christensen and L.P. Nielsen. (1995). Seasonal variation in nitrification and denitrification in estuarine sediment colonized by benthic microalgae and bioturbating infauna. Marine Ecology Progress Series 126: 111-121 Rysgaard S., N. Risgaard-Petersen and N.P. Sloth. (1996). Nitrification, denitrification, and nitrate ammonification in sediments of two coastal lagoons in Southern France. Hydrobiologia 329:133-141 Sand-Jensen, K. and Borum J. (1991). Interactions among phytoplankton, periphyton, and macrophytes in temperate freshwaters and estuaries. Aquatic Botany 41: 137-175. Sfriso A., B. Pavoni, A. Marcomini and A.A. Orio. (1992). Macroalgae, nutrient cycles, and pollutants in the lagoon of Venice. Estuaries 15: 517-528

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Sloth N.P., H. Blackburn, L.S. Hansen, N. Risgaard-Petersen and B.A. Lomstein. (1995). Nitrogen cycling in sediments with different organic loading. Marine Ecology Progress Series 116:163-170 Smith D.J. and G.J.C. Underwood. (2000). Exopolymer production by intertidal epipelic diatoms. Limnology and Oceanography 43:1578-1591. Stapel J. and M.A. Hemminga. (1997). Nutrient resorption from seagrass leaves. Marine Biology 128: 197-206 Stimson J. and S.T. Larned. (2000). Nitrogen efflux from the sediments of a subtropical bay and the potential contribution to macroalgal nutrient requirements. Journal of Experimental Marine Biology and Ecology 252: 159-180 Sundbäck K. and W. Granéli. (1988). Influence of microphytobenthos on the nutrient flux between sediment and water: a laboratory study. Marine Ecology Progress Series 43: 63-69 Sundbäck K., B. Jönsson, P. Nilsson and I. Lindström. (1990). Impact of accumulating drifting macroalgae on a shallow-water sediment system: an experimental study. Marine Ecology Progress Series 58: 261-274 Sundbäck K., L. Carlson, C. Nilsson, B. Jönsson, A. Wulff and S. Odmark. (1996a). Response of benthic microbial mats to drifting green algal mats. Aquatic Microbial Ecology 10: 195-208 Sundbäck K., P. Nilsson, C. Nilsson and B. Jönsson (1996b). Balance between autotrophic and heterotrophic components and processes in the sandy sediments: a field study. Estuarine Coastal and Shelf Science 43: 689-706 Sundbäck K. and A. Miles. (2000). Balance between denitrification and microalgal incorporation of nitrogen in microbial sediments, NE Kattegat. Aquaict Microbial Ecology 22: 291-300 Sundbäck K. and A. Miles. (2002). Role of microphytobenthos and denitrification for nutrient turnover in embayments with floating macroalgal mats: a spring situation. Aquat Mar Ecol. 30: 91-101 Sundbäck K., A. Miles and E. Göransson. (2000). Nitrogen fluxes, denitrification and the role of microphytobenthos in microtidal shallow-water sediments: an annual study. Marine Ecology Progress Series 200: 59-76 Sundbäck K., A. Miles, A. Hulth, L. Pihl, P. Engström, E. Selander and A. Svenson. (2003). Importance of benthic nutrient regeneration during initiation of macroalgal blooms in shallow bays. Marine Ecology Progress Series 246: 115-126 Suzumura M., T. Miyajima, H. Hata, Y. Umezawa, H. Kayanne and I. Koike. (2002). Cycling of phosphorus maintains the production of microphytobenthic communities in carbonate sediments of a coral reef. Limnology and Oceanography 47: 771-781 Terrados J., C.M. Duarte, L. Kamp-Nielsen, N.S.R. Agawin, E. Gacia, D. Lancap, M.D. Fortes, J. Borum, M. Lubanski and T. Greve. (1999). Are seagrass growth and survival constrained by reducing conditions of the sediment? Aquatic Botany 65: 175-107 Thornton D.C.O., L.F. Dong, G.J.C. Underwood and D.B. Nedwell. (2002). Factors affecting microphytobenthic biomass, species composition and production in the Colne Estuary (UK). Aquatic Microbial Ecology 27: 285-300 Thornton D.C.O., G.J.C. Underwood and D.B. Nedwell. (1999). Effect of illumination and emersion period on the exchange of ammonium across the estuarine sediment-water interface. Marine Ecology Progress Series 28: 11-20 Thybo-Christesen M., M.B. Rasmussen and T.H. Blackburn. (1993). Nutrient fluxes and growth of Cladophora sericea in a shallow Danish bay. Marine Ecology Progress Series 100: 273-81

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Trimmer M., D.B. Nedwell, D.B. Sivyer and S.J. Malcolm. (2000). Seasonal benthic organic matter mineralisation measured by oxygen uptake and denitrification along a transect of the inner and outer River Thames estuary, UK. Marine Ecology Progress Series 197:103-119 Turpin D.H.. (1991). Effects of inorganic N availability on algal photosynthesis and carbon metabolism. Journal of Phycology 27: 14-20 Tyler A.C., K.J. McGlathery and I.C. Anderson. (2001). Macroalgal mediation of dissolved organic nitrogen fluxes in a temperate coastal lagoon. Estuarine Coastal and Shelf Science 53: 155-168 Tyler A.C., K.J. McGlathery and I.C. Anderson. (2003). Benthic algae control sediment-water column fluxes of nitrogen in a temperate lagoon. Limnology and Oceanography In press. Underwood G.J.C. and J. Kromkamp. (1999). Primary production by phytoplankton and microphytobenthos in estuaries. Advances in Ecological Research 29: 93-153 Underwood G.J.C. and D.M. Paterson. (1993). Recovery of intertidal benthic diatoms after biocide treatment and associated sediment dynamics. Journal of the Marine Biology Association of the United Kingdom 73: 25-45 Valiela I., K. Foreman, M. LaMontagne, D. Hersh, J. Costa, P. Peckol, B. DeMeo-Andreson, C. D’Avanzo, M. Babione, S. Sham, J. Brawley and K. Lajtha. (1992). Couplings of watersheds and coastal waters: sources and consequences of nutrient enrichment in Waquoit Bay, Massachusetts. Estuaries 15: 443-457 Valiela I., J. McClelland, J. Hauxwell, P.J. Behr, D. Hersh and K. Foreman. (1997). Macroalgal blooms in shallow estuaries: Controls and ecophysiological and ecosystem consequences. Limnology and Oceanography 42: 1105-1118 Velimerov B. (1986). DOC dynamics in a Mediterranean seagrass system. Marine Ecology Progress Series 28: 21-41. van Katwijk M.M., L.H.T. Vergeer, G.H.W. Schmitz and J.G.M. Roelofs. (1997). Ammonium toxicity in eelgrass Zostera marina. Marine Ecology Progress Series 157: 159-173 Viaroli P. et al. (1992). Ulva rigida growth and decomposition processes and related effects on nitrogen and phosphorus cycles in a coastal lagoon (Sacca di Goro, Po River Delta), In: G. Colombo, et al. (ed.), Marine eutrophication and population dynamics, 77-84. Olsen and Olsen, Fredensborg. Viaroli P., M. Bartoli, C. Bondavalli and R.R. Christian. (1996). Macrophyte communities and their impact on benthic fluxes of oxygen, sulphide and nutrients in shallow eutrophic environments. Hydrobiologia. 329: 105-119 Ward B.B. and D.A. Bronk. (2001). Net nitrogen uptake and DON release in surface waters: importance of trophic interactions implied from size fractionation experiments. Marine Ecology Progress Series 219:11-24 Welsh D.T., S. Bourgues, R. de Wit and R.A. Herbert. (1996). Seasonal variations in nitrogen-fixation (acetylene reduction) and sulphate-reduction rates in the rhizosphere of Zostera noltii: nitrogen fixation by sulfate-reducing bacteria. Marine Biology 125: 619-628 Welsh D.T. (2000). Nitrogen fixation in seagrass meadows: Regulation, plant-bacteria interactions and significance to primary productivity. Ecology Letters 3:58-71 Welsh D.T., M. Bartoli, D. Nizzoli, G. Castaldelli, S.A. Riou and P. Viaroli. (2000). Denitrification, nitrogen fixation, community productivity and inorganic-N and oxygen fluxes in an intertidal Zostera noltii meadow. Marine Ecology Progress Series 208: 65-77.

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Wolfstein K., J.F.C. de Brouwer and L.J. Stal. (2002). Biochemical partitioning of photosynthetically fixed carbon by benthhic diatoms during short-term incubations at different irradiances. Marine Ecology Progress Series 245:21-31. Wulff A., K. Sundbäck, C. Nilsson, L. Carlson and B. Jonsson. (1997). Effect of sediment load on the microbenthic community of a shallow-water sandy sediment. Estuaries 20: 547-558 Ziegler S. and R. Benner. (1999). Dissolved organic carbon cycling in a subtropical seagrass-dominated lagoon. Marine Ecology Progress Series 180: 149-160

9. AFFILIATIONS K.J. McGlathery: Department of Environmental Sciences, P.O. Box 400123, University of Virginia, Charlottesville, VI 22904, USA. K. Sundbäck: Gothenburg University, Department of Marine Ecology, P.O. Box 461, S-40530 Gothenburg, Sweden. I.C. Anderson: School of Marine Science, Virginia Institute of Marine Science, College of William and Mary, Gloucester Point, VI 23602, USA.

NILS RISGAARD-PETERSEN

DENITRIFICATION

1. INTRODUCTION As nitrogen passes through the estuarine environment it undergoes a complex series of transformations. Inorganic N is incorporated into pelagic primary producers, and a large fraction reaches the seafloor either through sedimentation of algal cells or via grazing followed by deposition of faecal pellets and dead organisms. At the seafloor the organic bound nitrogen is either buried permanently or undergoes decomposition resulting in release of NH4+ or urea. Alternatively, inorganic N is taken up directly by the sediment system and incorporated into benthic primary producers or used as substrate in microbial metabolic processes. Remineralized nitrogen may be liberated from the sediment to the water column, often through a complex series of microbiologically mediated metabolic processes that determine the form in which N is liberated and consequently the degree to which remineralized N is made available for new primary production. Sediment nitrification and denitrification are central microbiological processes in this context. Nitrification is the aerobic conversion of NH4+ to NO3-, the most oxidized form of nitrogen. The process is mediated by bacteria of the family Nitrobacteraceae that grow chemolitho-autotrophically, meaning that they derive energy from the oxidation of inorganic compounds (N) and use CO2 as sole carbon source. The oxidation of NH4+ to NO3- proceeds in two steps and is carried out by two types of organisms that make use of different metabolic pathways: The oxidation of NH4+ to NO3- is performed by ammonium oxidizing bacteria while the oxidation of NO2- to NO3- is performed by nitrite oxidizing bacteria. Denitrification is the anaerobic conversion of oxidized nitrogen (NO3- or NO2-) to reduced gaseous compounds (N2O or N2). The process is mediated by facultative aerobes capable of utilizing oxidized N as terminal electron acceptor in a dissimilatory process when O2 becomes limiting. The ability to perform this process is found in a variety of taxonomic groups including litotrophic and phototrophic bacteria. Most denitrifiers grow organotrophically, however (Zumft 1992), meaning that they are able to derive energy from the oxidation of reduced organic carbon. Denitrification is an important factor determining the N-availability in estuarine environments, since only a limited number of marine organisms can utilize the end product N2 as an N-source (Howarth, et al. 1988).

263 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 263-280. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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Table 1. Recent measurements of denitrification rates in European estuaries.

Site Rörtungen , Sweden

Denitrification rate (mmol N m-2 d-1) 0.02-0.1

Randers Fjord, Denmark

0.1-8

Targus Estuary, Portugal

0-6

Løgstør Bredning, Denmark 0.08-0.26 River Thames Estuary, UK

0.02-273

Bassin d’Arcachon, France

0.05-0.14

Source Sundback et al. 2000 Nielsen et al. 2001 Cabrita and Brotas 2000 Risgaard-Petersen et al. 1998 Trimmer et al. 2000 Welsh et al. 2000

Rates of nitrification and denitrification have been measured intensively in European estuaries during the last decades. Sediment denitrification and nitrification are regulated by a complex matrix of biological, physical and chemical factors, and actual in situ activity may vary orders of magnitude (Table 1) depending on season, nutrient load and the composition of the ecosystem. In the following section major factors regulating the processes will be discussed with special reference to responses to organic loads, nutrient concentrations and benthic primary production.

2. NITRIFICATION AND DENITRIFICATION IN SEDIMENTS Basically the processes of nitrification and denitrification are located in the upper few mm of the sediment. Figure 1 shows examples of porewater profiles of NO3-, NH4+ and O2 in unvegetated sediment as measured with microsensors. On the basis of these profiles the distribution of nitrification and denitrification can be estimated (Berg et al. 1998). Being a strictly aerobic process, nitrification takes place in the oxic zone, as indicated in the figure, and the substrate NH4+ is supplied mainly by mineralization processes in the anoxic sediment strata while O2 is supplied by the water column. Denitrification takes place just below the oxic/anoxic interface, and apart from ensuring a supply of NO3- from the water column, the close proximity of the nitrification and denitrification processes facilitates an efficient coupling between the two processes through transport of nitrification products (NO2- /NO3-) across the oxic/anoxic interface. This coupling implies a loss of remineralized N from the system, and nitrification therefore plays a key role in determining the fate of organic bound nitrogen that is deposited on the seafloor.

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In order to understand the regulation of denitrification in estuarine sediments it is useful to look at the dependency on the NO3- source: The dependency of denitrification on bottom-water NO3- and on NO3- produced by sediment nitrification, respectively. Both of these dependencies are measurable with the isotope pairing technique (Nielsen 1992) and can also be deduced from interpretation of NO3porewater profiles (Meyer et al. 2001).

O2 (µM)

Depth(mm)

0 0 1 2 3 4 5 6 7

100

200

300

-

NO3

+

NH4 O2

0 20 40 60 80 100 -0.2 -0.1 0.0 0.1 0.2 -

+

NO3 /NH4 (µM)

-2

-1

Production (nmol cm s )

Figure 1. Porewater profiles of NO3-, NH4+ and O2 as measured with ion selective sensors (Ref) (left) and volume specific NO3- production (right). Positive production indicates nitrification, while negative production indicates denitrification. Data from Berg et al. (1998).

3. DENITRIFICATION OF NITRATE FROM THE WATER COLUMN A prerequisite for denitrification of bottom-water NO3- is the transport of NO3- from the water column to the anoxic sediment strata. Fich’s laws of diffusion imply that the diffusive transport is dependent on both the thickness of the oxic zone and the concentration of NO3- in the water column. The rate of denitrification of bottomwater NO3- is therefore ideally inversely proportional to the O2 penetration depth and directly proportional to the concentration of NO3- in the water column. This ideal relationship was demonstrated experimentally by Rysgaard et al. (1994) via manipulations of bottom-water O2 and NO3- concentrations. In estuarine systems oxygen penetration depths and bottom-water NO3concentrations may vary independently. The O2 penetration depth is determined by the bottom-water O2 concentration and in sediments without benthic photosynthesis

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by the rate at which O2 is consumed in the sediment (Revsbech et al. 1980). Thus, an increase in O2 consumption will lead to a decrease in O2 penetration depth. On seasonal scale maximum sediment O2 consumption rates and minimum O2 penetration depths are found mainly during the late summer months, while the NO3load is largest in the winter and early spring months. Therefore, the correlation between bottom-water NO3- and denitrification observed in seasonal studies can be poor (Dong et al. 2000 a. o.). Christensen et al. (1990) developed a simple model to predict the rate of denitrification of bottom-water NO3- from the bottom-water O2 and NO3concentrations and the sediment O2 consumption:

⎡ ⎤ D( NO3−) C NO3 1 ) − 1⎥ (9.1) Dw = FO2 α ⎢ (1 + D (O 2 ) C O α ⎢⎣ ⎥⎦ 2 Where Dw is the rate of denitrification of NO3- from the water column and FO2 the sediment O2 consumption rate. D(NO3-) and D(O2) represent the diffusion coefficients of NO3- and O2, respectively, while α is the ratio between depth-specific denitrification and O2 consumption activity (app. 0.8 (Christensen et al. 1990)). The model assumes that the depth-specific activities of denitrification and O2 consumption are constant and that diffusion is the major transport process and, furthermore, that denitrification is the sole NO3- consuming process. This implies that the model is not useful in sediments with significant algal activity or in sediments where bacteria capable of reducing NO3- to NH4+ dominate NO3- reduction. In such sediments the model typically overestimates denitrification. In sediments where the model assumptions are fulfilled, however, the model predicts the rate of denitrification well. Data from station 75, Randers Fjord (Nielsen et al. 2001), is a good example of application of the model (Fig. 2). Rysgaard et al. (1995) demonstrated an equally good agreement between modelled and measured data in the Danish estuary Kertinge Nor. The model obviously predicts that an increase of the NO3- load will result in an increase in denitrification activity. The model also predicts that increasing the organic load to sediments (for instance through sedimentation of phytoplankton cells following an algal bloom) and a resultant increase in the sediment O2 consumption rate (Sloth et al. 1995) will lead to elevated removal of bottom-water NO3- via denitrification. This relationship has been demonstrated through in situ measurements as well as experimentally. Jensen (1988) thus observed increasing denitrification activity following the spring phytoplankton bloom in the Bay of Aarhus. Caffrey et al. (1993) and Sloth et al. (1995) likewise demonstrated experimentally that increasing the organic load to sediments would result in elevated denitrification of bottom-water NO3-.

Denitrification

267

300

-2

-1

Measured Dw (µmol N m h )

350

250 200 150 100 50 0 0

50

100

150

200

250

300 -2

350

-1

Model estimate (µmol N m h ) Figure 2. Measured vs. modelled rates of denitrification of NO3- supplied from the water column (DW). The model accurately predicts the measured values (r2 =0.9289, p<0.0001). Data from Nielsen et al. (2001).

4. DENITRIFICATION BASED ON NO3- PRODUCED BY NITRIFICATION The critical step in the removal of remineralized N from sediments through coupled nitrification-denitrification is generally the nitrification process and its vertical distribution in the oxic zone. The distribution of the processes shown in Fig. 1 is typical of a stable sediment environment where bacterial populations have established themselves in zones with an optimal substrate supply, and in sediments such as this a very efficient coupling between nitrification and denitrification can be achieved. In more unstable environments, e.g. sediments exposed to frequent resuspension events, bioturbation or benthic primary production, a more even distribution of nitrification in the oxic zone can be observed (Meyer et al. 2001). In this type of sediment the coupling between nitrification is less efficient and a higher proportion of the NO3produced through sediment nitrification will escape denitrification and thus become available for new primary production. Being an aerobic litotrophic process nitrification is dependent on the supply of both NH4+ and O2. Therefore, O2 deficit in the water column will obviously cause nitrification to cease and with it the coupled nitrification-denitrification (Rysgaard et al. 1994).

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In estuarine sediments NH4+ produced through mineralization of organic bound nitrogen is the major NH4+ source for nitrification since NH4+ concentrations in the water column are generally low. However, mineralization of organic bound nitrogen results in consumption of O2 either directly or indirectly via the oxidation of reduced compounds such as H2S, Mn+ and Fe++ produced via anoxic mineralization. Thus, the biogeochemical processes responsible for the production of substrate for nitrification at the same time compete with the nitrifying bacteria for O2. Blackburn (1990) demonstrated this antagonistic dependency of nitrification and coupled nitrification-denitrification on the sediment mineralization processes through model simulations. Based on these simulations it was concluded that increasing the organic load to the sediment and thereby promoting mineralization does not result in a parallel stimulation of the removal of nitrogen via coupled nitrificationdenitrification. Beyond a certain point nitrification becomes O2 limited due to oxygen consumption resulting from increased carbon availability, and this reduces the loss of nitrogen via coupled nitrification-denitrification. Experimental evidence for this idea was provided by Sloth et al. (1995) who measured nitrification and denitrification in sediment cores loaded with different amounts of organic material (Fig. 3). Nitrification and denitrification rates were highest in the moderately loaded sediments because the availability of NH4+ was greater in these cores than in the unloaded control cores.

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Figure 3. Coupled nitrification-denitrification in sediment cores loaded experimentally with yeast. 30 g DW and 100 g DW designate mixing of 30 g DW and 100 g DW of organic material into the sediment, respectively, while 100 surf designates the addition of 100 g of organic material to the sediment surface. Redrawn from Sloth et al. (1996).

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In cores heavily loaded with organic material nitrification and consequently coupled nitrification-denitrification was significantly reduced or absent because of O2 limitation and probably also due to irreversible inhibition of nitrification by H2S (Henriksen and Kemp 1988; Joye and Hollibaugh 1995) during the initial phases of the experiment. A case story from Kertinge Nor, Denmark, reported by Dalsgaard et al. (1999) is a good example of how organic loading to the estuaries may influence N loss from sediments via coupled nitrification-denitrification. The external nutrient load to the estuary was reduced from the year 1990, but two years later, in 1992, the estuary was still hypereutrophic and exhibited a high sediment O2 demand (2705 ton O2 y-1) due to deposition of planktonic alga cells fuelled by a large pool of nutrients that had accumulated in the sediment. In 1995 the sediment nutrient pool was markedly reduced, and this resulted in reduced pelagic primary production and a decrease in the organic load reaching the sediment. Sediment O2 consumption was reduced to 1578 t O2 y-1 and coupled nitrification-denitrification increased from 6.1 t N y-1 in 1992 to 17.1 t N y-1 in 1995.

5. ALTERNATIVE NO3- REDUCTION PATHWAYS: DNRA Dissimilatory NO3- reduction to NH4+ (DNRA) represents an alternative NO3reduction pathway which may compete with denitrification. Bacteria capable of reducing NO3- to NH4+ may dominate anaerobic NO3- reduction in carbon rich environments low in NO3- (Tiedje 1982, Tiedje 1988). Gilbert et al. (1997) thus reported a capacity for DNRA in sediments exposed to high organic loading from mussel farms in a Mediterranean lagoon. In a study in the Danish estuary Horsens Fjord (Christensen et al. 2000) it was observed that the relative importance of DNRA decreased with increasing distance from fish cages deployed in the fjord. Just below the cage approximately 95% of NO3- reduction proceeded via the ammonification pathway, while denitrification was the dominant NO3- reducing process 100 m from the cage. The dataset suggested a strong positive correlation between C-mineralization (measured as sediment O2 consumption) and the relative importance of DNRA (Fig. 4). Thus, with increasing mineralization activity DNRA becomes relatively more important than denitrification. Despite the fact that DNRA has been known to take place in decades there is still a lack of understanding of the role of the process in estuaries, and systematic measurements of the distribution of the process relative to denitrification are very much needed.

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Figure 4. Proportion of denitrification and DNRA to total NO3- reduction as function of the sediment O2 consumption rates. Data from Christensen et al. (2000).

6. DENITRIFICATION IN SEDIMENTS COLONIZED BY BENTHIC MICROALGAE Benthic microalgae are concentrated at the sediment/water interface, and their activity may significantly alter the chemical microenvironment in the surface sediment. Microphytobenthic photosynthesis and respiration may result in large diurnal oscillations in both O2 concentrations and the concentrations of inorganic carbon as well as in pH (Revsbech 1983). Likewise, through assimilatory nutrient uptake, algae may have a significant impact on the porewater concentrations of NO3- and NH4+ in the surface sediment. Examples of diurnal oscillations in O2 and pH in estuarine sediment colonized by microalgae are given in Figure 5. During illumination (200 µmol photons m-2 s-1) photosynthesis resulted in an O2 concentration of 500 µM in the most active layer approximately 0.2 mm below the surface. The O2 penetration depth was 2.5 mm. At the depth of maximum activity pH was 9.5 due to intense microalgal carbon assimilation. In darkness respiration resulted in a decrease in both sediment O2 concentrations and pH. The O2 penetration depth in darkness was only 25% of the depth measured in light.

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Examples of porewater profiles of NO3-, NH4+ and O2 in freshwater sediment colonized by benthic microalgae are given in Figure 6. During illumination NH4+ was not detectable in the oxic zone and a significant reduction in NO3- concentration was observed in the 0.5-mm depth interval. The zone within the oxic layer exhibiting the lowest NO3- concentration corresponded to the zone in which the O2 concentration peaked, and this suggests that microphytobenthic NO3- uptake was responsible for the observed decrease in NO3-. In darkness no indications of microphytobenthic N uptake were seen. Ammonium penetrated the oxic zone and was released to the water column, while NO3- was consumed exclusively in the anoxic zone, a fact that can be attributed to denitrification.

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Figure 5. Porewater profiles of O2 and pH measured with microsensors in a microalgaecolonized sediment. Original (data).

The dramatic changes in porewater chemistry of the surface sediment induced by benthic microalgae must be expected to have a significant impact on the nitrification and denitrification processes. The presence of benthic microalgae may lead to diurnal variations in nitrification and denitrification activity and to a reduction of the overall capacity of the sediment to remove nitrogen via denitrification. Nielsen et al. (1990) investigated the effect of photosynthesis and assimilation on denitrification of bottom-water NO3- in stream sediment exposed to high NO3concentrations (300 µM). The authors found that the overall rate of denitrification was reduced by 70% in light. This reduction was ascribed to the O2 penetration depth being greater in light than in the dark, leading to an increase in diffusion distance for NO3- from the water column to the anoxic sediment strata. Calculated porewater NO3profiles suggested that microalgal NO3- assimilation did not influence denitrification. Risgaard-Petersen et al. (1994) demonstrated a similar mechanism in a study on nitrification and denitrification in estuarine sediments. Using NO3- biosensors (Larsen

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et al. 1997) Meyer et al. (2001) showed that at low NO3- concentrations (4 µM) microalgae are efficient sinks for water-column NO3-, however, and are able to significantly impede the flux of NO3- toward anoxic sediment layers through nutrient uptake.

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Figure 6. Porewater profiles of NO3-, NH4+ and O2 measured with ionselective sensors (NO3/NH4+) in an alga-colonized sediment. Original data.

The diurnal variations in O2 penetration depth may also induce diurnal variations in nitrification activity and consequently in coupled nitrification-denitrification activity. Risgaard-Petersen et al. (1994), Lorenzen et al. (1998) and An and Joye (2001) observed higher rates of coupled nitrification-denitrification in light than in darkness. This stimulation was attributed to a photosynthetically mediated increase in O2 penetration depth leading to increased nitrification activity (Risgaard-Petersen et al. 1994). However, several field studies have indicated that light-induced stimulation of nitrification is not a general phenomenon (Rysgaard et al. 1995; Cabrita and Brotas 2000; Sundback and Miles 2000). It has been suggested that stimulation of coupled nitrification-denitrification in light only occurs when competition between algae and nitrifiers and denitrifiers for DIN is not limiting the bacterial processes (Rysgaard et al. 1995; Dong et al. 2000). Although attractive this hypothesis is unsupported by experimental evidence at present. In a review on nitrification Henriksen and Kemp (1988) presented experimental data showing that the potential for nitrification in the uppermost sediment strata was reduced in sediments exposed to light/dark cycles for 6 weeks as compared to sediments incubated in the dark. It was suggested that light and a combination of factors linked to algal activity (i.e. high O2 concentrations, high pH and induction of NH4+ and CO2 limitation) were responsible for the observed inhibition.

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Figure 7. Coupled nitrification-denitrification rates in bare sediments with and without microphytobenthic photosynthesis. Sediments are classified as heterotrophic if O2 fluxes in light ≤ O2 fluxes in the dark and <0 (n=200) Sediment is classified as autotrophic if O2 fluxes in light > O2 fluxes in the dark and >0 (n=944). Data from Dalsgaard et al. (2000).

Based on their observations the authors proposed that this reduced ability of nitrifying bacteria to survive in alga-colonized sediment would lead to a reduced loss of nitrogen via coupled nitrification-denitrification from the system as compared to sediment without benthic microalgae. Results from the EU-funded NICE project (Nitrogen Cycling in Estuaries (Dalsgaard et al. 2000)) have shown that mean coupled nitrification-denitrification rates in sediments colonized by benthic microalgae are significantly lower (ca. 50%, Fig. 7) than rates measured in heterotrophic sediments. Experimental studies indicate that a major factor responsible for the reduction in coupled nitrification-denitrification in alga-colonized sediments is the generation of unfavourable growth conditions for nitrifying bacteria in these systems. Furthermore, induction of nitrogen limitation of nitrifying bacteria seems to be a major controlling mechanism in alga-colonized sediments, while factors associated with photosynthesis (e.g. high pH and O2 concentrations) are of minor importance (Risgaard-Petersen, 2003). Benthic microalgae may enhance the flux of inorganic N from the water column toward the sediment through nutrient assimilation and are likewise able to reduce the efflux of remineralized nitrogen from the sediment. The suppression of nitrification and denitrification in these sediments suggests a highly conservative environment for nitrogen characterized by efficient coupling between sediment N regeneration and N assimilation.

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Rooted aquatic angiosperms may alter the sediment environment and thereby the niches occupied by nitrifying and denitrifying bacteria. The plants may release O2 to the rhizosphere via radial O2 loss from the roots if the respiratory demand of the roots is lower than the O2 supply or if O2 consuming processes in the rhizosphere take place closer to the O2 rich cortex tissue compared to O2 consuming processes in the root tissue (Armstrong 1994). Consequently, free O2 may be present in the otherwise anoxic rhizosphere if the reductant formation through chemical and microbial processes is sufficiently slow (Caffrey and Kemp 1992; Pedersen et al. 1995 a.o.). This introduction of O2 into the rhizosphere creates oxic/anoxic interfaces in an otherwise reduced environment. Aquatic plants also assimilate and incorporate inorganic nitrogen into organic matter and a significant part of this nitrogen may be supplied from the sediment (Short and McRoy 1984; Caffery and Kemp 1992; Pedersen and Borum 1992). Thus, the plants compete with nitrifying and denitrifying bacteria for nitrogen, and recent results have shown that angiosperm roots may be better competitors than nitrifiers for sediment NH4+ (Verhagen et al. 1995). Several studies on freshwater systems have demonstrated the ability of aquatic plants to stimulate nitrogen loss through O2 release in the rhizosphere (Reddy et al. 1989; Christensen et al. 1986; RisgaardPetersen and Jensen 1997; Arth et al. 1998). In the case of marine species the dataset is more limited. Direct evidence for rhizosphere-associated coupled nitrificationdenitrification has only been presented for a few species and the activity and the ability of the plants to stimulate coupled nitrification-denitrification via O2 release is apparently strongly linked to the growth cycle of the plants and therefore displays a distinct seasonal pattern. Caffrey and Kemp (1992) observed that denitrification activity in Potamogeton perfolitiatus-colonized sediments was higher than the activity measured in bare sediments in July when the plant biomass was high. On the other hand, Risgaard-Petersen et al. (1998) observed that in Zostera marina surfaceand rhizosphere-associated coupled nitrification-denitrification was only measurable in the spring. In the summer months a high sediment O2 demand probably prevented the formation of oxic/anoxic interfaces in the rhizosphere and with it root-associated nitrification-denitrification. It has been suggested that denitrification activity is enhanced in seagrass-colonized sediments compared to the activity in bare sediments due to the ability of the plants to stimulate coupled nitrification-denitrification via root O2 release (Flindt 1994). However, direct measurements using 15N isotopes have shown that the ability of for instance Zostera marina and Z. noltii to stimulate coupled nitrification-denitrification via root O2 release is generally low (Rysgaard et al. 1996; Risgaard-Petersen et al. 1998; Ottosen et al. 1999; Welsh et al. 2000). A comparison between Danish sediments with and without benthic macrophytes presented by Risgaard-Petersen and Ottosen (2000) showed that coupled nitrification-denitrification on average corresponded to 24% of the activity reported from various unvegetated sediments in Danish fjords

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Figure 8. Depth profiles of coupled nitrification-denitrification measured in light and in the dark in Z. marina-colonized sediment. In April maximum activity was measured in the upper sediment strata in light and denitrification was measurable until a depth of 8 cm. Coupled nitrification-denitrification activity in the deeper sediment strata was caused by O2 release from the roots. In darkness the activity was below detection limit. In August coupled nitrification-denitrification activity was undetectable. Redrawn from Risgaard-Petersen et al. (1998).

This suggests that eelgrass-colonized sediment is an unfavorable environment for nitrifying and denitrifying bacteria, possibly due to the effect of competition with the plants for nitrogen as shown in for instance Rumex species (Engelaar et al. 1991). Alternatively, the unfavorable conditions may be a result of enhanced organic loading to the sediment induced by plants (Hemminga et al. 1991) or enhanced sediment SO42- reduction activity (Blaaberg et al. 1998; Holmer and Nielsen 1997). These conditions may inhibit nitrification (Henriksen and Kemp 1988; Sloth et al. 1995) and improve conditions for bacteria capable of dissimilatory NO3- reduction to NH4+ (Tiedje 1988). The latter process has recently been shown to play a major role in seagrass-colonized sediments (Rysgaard et al. 1996). Just like microphytobenthos colonized sediments the seagrass meadows may thus represent a highly conservative environment for nitrogen, where the N-cycle is dominated by the primary productivity of the plant community and the associated assimilatory demand for fixed N to support this productivity.

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Fig 9. Coupled nitrification-denitrification rates in Z. marina-colonized sediments (n=23) and bare sediments from Danish estuaries (n=71) exposed to N-concentrations below 12 µM. Data from Risgaard-Petersen and Ottosen (2000).

8. CONCLUSIONS In this chapter the process of denitrification has been discussed in relation to responses to organic loads, nutrient concentrations and benthic primary production. On the basis of this discussion I propose the following scenarios regarding the process in relation to changes in nutrient load to the estuarine environment: When nutrient loading to estuaries is increased and with it the concentration of nitrogen in the water column, denitrification will increase in response to the elevated NO3- concentrations. Given that increased N loading results in elevated pelagic primary production (Borum and Sand-Jensen 1996) and a concurrent increase in deposition of organic matter at the seafloor, sediment mineralization will increase and the NH4+ availability for nitrifying bacteria improve. This will, to a certain point, lead to enhanced nitrogen removal via coupled nitrification-denitrification. However, as mineralization increases, nitrification will become O2 limited leading to a reduced loss of nitrogen via coupled nitrification-denitrification. If water-column O2 concentrations decline to anoxia as often happens below the pycnocline when the organic load is high, coupled nitrification-denitrification will cease and remineralized nitrogen will escape from the sediment as NH4+. At high nutrient loads denitrifying bacteria may eventually become out-competed by bacteria capable of

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reducing NO3- to NH4+. This may further contribute to the establishment of a positive feedback loop between pelagic primary production and sediment NH4+ generation. When nutrient loading to estuaries is reduced the amount of NO3- available for denitrification decreases. Furthermore, as light conditions at the seafloor improve, a shift from pelagic to benthic primary production is to be expected (Borum and SandJensen 1996). A shift such as this will significantly alter the benthic N cycle. Benthic primary producers will assimilate a greater amount of water-column NO3- before it becomes available for denitrification, and may likewise reduce the transport of NO3from the water column to the anoxic sediment strata via a photosynthetically mediated increase in O2 penetration depth. Moreover, benthic primary producers seem to create unfavorable growth conditions for nitrifying bacteria resulting in inhibition of coupled nitrification-denitrification. Thus, reduced nutrient loading will lead to reduced loss of N via denitrification and a temporary accumulation of N in plant biomass.

9. REFERENCES An, S.M. and S.B. Joye (2001). Enhancement of coupled nitrification-denitrification by benthic photosynthesis in shallow estuarine sediments. Limnology and Oceanography 46:62-74 Arth, I., P. Frentzel and R. Conrad (1998). Denitrification coupled to nitrification in the rhizosphere of rice. Soil. Biology and Biochemistry 40:509-515 Berg, P., N. Risgaard-Petersen and S. Rysgaard (1998). Interpretation of measured concentration profiles in sediment pore water. Limnology and Oceanography 43:1500-1510 Blackburn, T.H. (1990). Denitrification model for marine sediments 323-337. In: Revsbech, N. P. and J. S. Sørensen [eds.], Denitrification in soils and sediments.Plenum Press Blaaberg, V., K. Mouritsen and K. Finster (1998). Diel cycles of sulphate reduction rates in sediments of a Zostera marina bed, Denmark. Aquatic . Microbial . Ecology l 15:97-102 Borum, J. and K. Sand-Jensen (1996). Is total primary production in shallow coastal matine waters stimulated by nitrogen loading? Oikos 76:406-410 Cabrita, M.T. and V. Brotas (2000). Seasonal variation in denitrification and dissolved nitrogen fluxes in intertidal sediments of the Tagus estuary, Portugal. Marine Ecology Progress Series 202:51-65 Caffrey, J.M. and W.M. Kemp (1992). Influence of the submersed plant Potamogeton Perfuliatus, on nitrogen cycling in estuarine sediments. Limnology and Oceanography 37:1483-1495 Caffrey, J.M., N.P. Sloth, H.F. Kaspar, and T.H. Blackburn (1993). Effect of organic loading on nitrification and denitrification in a marine sediment microcosm. FEMS Microbial Ecology. 12:159167 Christensen, P.B., L.P. Nielsen, J. Sørensen and N.P. Revsbech (1990). Denitrification in nitrate-rich streams: Diurnal and seasonal variation related to benthic oxygen metabolism. Limnology and Oceanography 35:640-651 Christensen, P.B., S. Rysgaard, N.P. Sloth, T. Dalsgaard and S. Schwaerter (2000). Sediment mineralization, nutrient fluxes, denitrification and dissimilatory nitrate reduction to ammonium in an estuarine fjord with sea cage trout farms. Aquatic Microbial Ecology 21:73-84

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Dalsgaard, T., P.B. Christensen, S. Rysgaard and N. Risgaard-Petersen (1999). Kvælstoffjernelse i danske kystnære farvande: betydning og regulering 102-118. In: Aa., L. B. [eds.], Havmiljøet ved årtusindskiftet.Olsen &Olsen Dalsgaard, T., L.P. Nielsen, V. Brotas, P. Viaroli, G.J.C. Underwood, D.B. Nedwell, K. Sundbäck, S. Rysgaard, A. Miles, M. Bartoli, L.F. Dong, D.C.O. Thornton, O.L.D.M, G. Castaldelli and N. Risgaard-Petersen (2000). Final Scientific report for NICE Nitrogen Cycling in Estuaries Commision of the European Communities DC XII, Brussels Dong, L.F., D.C.O. Thornton, D.B. Nedwell and G.J.C. Underwood (2000). Denitrification in sediments of the River Colne estuary, England. Marine Ecology Progress Series 203:109-122 Flindt, M.R (1994). Measurements of nutrient fluxes and mass balances by on-line in situ dialysis in a Zostera marine bed culture. Verhandlungen Internationale Vereinigung Limnologie 25:2259-2264 Gilbert, F., P. Souchu, M. Bianchi and P. Bonin (1997). Influence of shellfish farming activities on nitrification, nitrate reduction to ammonium and denitrification at the water-sediment interface of the Thau Lagoon, France. Marine Ecology Progress Series 151:143-153 Hemminga, M.A., P.G. Harrison and F. van Lent (1991). The balance of nutrient losses and gains in seagrass meadows. Marine Ecology Progress Series 71:85-96 Henriksen, K. and M. Kemp (1988). Nitrification in estuarine and coastal marine sediments 201-255. In: Blackburn, T. H. and J. Sørensen [eds.], Nitrogen cycling in coastal marine environments, SCOPE. John Wiley & Sons Ltd Holmer, M. and S.L. Nielsen (1997). Sediment sulfur dynamics related to biomass-density in Zostera marina (eelgrass) beds. Marine Ecology Progress Series. 146:163-171 Howarth, R.W., R. Marino, J. Lane and J.J. Cole (1988). Nitrogen-Fixation in Fresh-Water, Estuarine, and Marine Ecosystems .1. Rates and Importance. Limnology and Oceanography 33:669-687 Jensen, M.H., T.K. Andersen, and J. Sørensen (1988). Denitrification in coastal bay sediment: regional and seasonal variation in Aarhus Bight, Denmark. Marine Ecology Progress Series. 48:155-162 Larsen, L.H., T. Kjær and N.P. Revsbech (1997). A microscale NO3- biosensor for environmental applications. Analytcal Chemistry. 69:3527-3531 Lorenzen, J., L.H. Larsen, T. Kjaer and N.P. Revsbech (1998). Biosensor determination of the microscale distribution of nitrate, nitrate assimilation, nitrification, and denitrification in a diatom-inhabited freshwater sediment. Applied and Environmental Microbiology 64:3264-3269 Meyer, R.L., T. Kjaer and N.P. Revsbech (2001). Use of NOx- microsensors to estimate the activity of sediment nitrification and NOx- consumption along an estuarine salinity, nitrate, and light gradient. Aquatic Microbial Ecology 26:181-193 Nielsen, K., N. Risgaard-Petersen, B. Somod, S. Rysgaard and T. Bergo (2001). Nitrogen and phosphorus retention estimated independently by flux measurements and dynamic modelling in the estuary, Randers Fjord, Denmark. Marine Ecology Progress Series 219:25-40 Nielsen, L.P. (1992). Denitrification in sediments determined from nitrogen isotope pairing. FEMS Microbial. Ecology. 86:357-362 Nielsen, L.P., P.B. Christensen, N.P. Revsbech and J. Sørensen (1990). Denitrification and photosynthesis in stream sediment studied with microsensor and whole-core techniques. Limnology and Oceanography 35:1135-1144

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Ottosen, L.D.M., N. Risgaard-Petersen and L.P. Nielsen (1999). Direct and indirect measurements of nitrification and denitrification in the rhizosphere of aquatic macrophytes. Aquatic Microbial Ecology 19:81-91 Risgaard-Petersen, N 2003. Coupled nitrification-denitrification in autotrophic and heterotrophic estuarine sediments: On the influence of benthic microalgae. Limnology and Oceanography 48:93-105. Risgaard-Petersen, N. and L.D.M. Ottosen (2000). Nitrogen cycling in two temperate Zostera marina beds: seasonal variation. Marine Ecology Progress Series 198:93-107 Risgaard-Petersen, N., S. Rysgaard, L.P. Nielsen and N.P. Revsbech (1994). Diurnal variation of denitrification and nitrification in sediments colonized by benthic microphytes. Limnology and Oceanography 39:573-579 Risgaard-Petersen, N., T. Dalsgaard, S. Rysgaard, P.B. Christensen, J. Borum, K. McGlathery and L.P. Nielsen (1998). Nitrogen balance of a temperate eelgrass Zostera marina bed. Marine Ecology Progress Series 174:281-291 Rysgaard, S., P.B. Christensen and L.P. Nielsen (1995). Seasonal variation in nitrification and denitrification in estaurine sediment colonized by benthic microalgae and bioturbating infauna. Marine Ecology Progress Series. 126:111-121 Rysgaard, S., N. Risgaard-Petersen and N.P. Sloth (1996). Nitrification, denitrification, and nitrate ammonification in sediments of two coastal lagoons in Southern France. Hydrobiologia 329:133-141 Rysgaard, S., N. Risgaard-Petersen, N.P. Sloth, K. Jensen and L.P. Nielsen (1994). Oxygen Regulation of Nitrification and Denitrification in Sediments. Limnology and Oceanography 39:1643-1652 Sloth, N.P., H. Blackburn, L.S. Hansen, N. Risgaard-Petersen and B. Lomstein, Aa. (1995). Nitrogen cycling in sediments with different organic loading. Marine Ecology Progress Series. 116:163-170 Sundback, K. and A. Miles (2000). Balance between denitrification and microalgal incorporation of nitrogen in microtidal sediments, NE Kattegat. Aquatic Microbial Ecology 22:291-300 Sundback, K., A. Miles and E. Goransson (2000). Nitrogen fluxes, denitrification and the role of microphytobenthos in microtidal shallow-water sediments: an annual study. Marine Ecology Progress Series 200:59-76 Tiedje, J.M., Sexstone,A..J.,Myrold,D.D. and Robinson,J.A. (1982). Denitrification: ecological niches, competition and survival. Antonie van Leeuwenhoek 48:569-586 Tiedje, J.M. (1988). Ecology of denitrification and dissimilatory nitrate reduction to ammonium 179-244. In: Zender, A. J. B. [eds.], Ecology of anaerobic microorganisms. John Wiley and Sons Trimmer, M., D.B. Nedwell, D.B. Sivyer and S.J. Malcolm (2000). Seasonal benthic organic matter mineralisation measured by oxygen uptake and denitrification along a transect of the inner and outer River Thames estuary, UK. Marine Ecology Progress Series 197:103-119 Welsh, D.T., M. Bartoli, D. Nizzoli, G. Castaldelli, S.A. Riou and P. Viaroli (2000). Denitrification, nitrogen fixation, community primary productivity and inorganic-N and oxygen fluxes in an intertidal Zostera noltii meadow. Marine Ecology Progress Series 208:65-77 Zumft, W. G. (1992). The denitrifying prokaryotes. 544-582 in: A. Balows,The prokaryotes : A handbook on the biology of bacteria: Ecophysiology, isolation, identification, applications. Springer

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N. Risgaard-Petersen: Department of Marine Ecology, National Environmental Research Institute, Vejlsøvej 25, DK-8600 Silkeborg, Denmark.

SØREN LAURENTIUS NIELSEN, MORTEN FOLDAGER PEDERSEN & GARY T. BANTA

ATTEMPTING A SYNTHESIS – PLANT/NUTRIENT INTERACTIONS 1. PRIMARY EFFECTS OF EUTROPHICATION The overall purpose of this book has been to summarize and evaluate the mutual interactions between nutrients and vegetation in shallow coastal ecosystems. Several books and scientific papers have focused on how nutrient richness modify the structure and dominance patterns of marine plant communities, while less attention has been given to the evaluation of how different types of marine plants may affect nutrient dynamics within the systems they live. The major topic of this book is therefore to review how plants affect the overall nutrient dynamics (i.e. transport, recycling, temporary or permanent retention) in shallow coastal ecosystems and to judge whether or not such effects may vary systematically with dominance of different plant types – “plants” in this context includes algae. Shallow coastal ecosystems are, in contrast to the ocean where phytoplankton is the only important primary producer, inhabited by taxonomically and functionally diverse plant communities. The plant communities in shallow coastal ecosystems are typically made up by different plant types including rooted plants (e.g. seagrasses, salt marsh plants etc.), slow-growing, perennial macroalgae, fast-growing, ephemeral macroalgae and benthic, epiphytic and pelagic microalgae that each may contribute significantly to total plant biomass and production (Chapter 1). Comparative studies show that these plant types differ systematically from each other in a number of morphologically, physiological, functional and ecological properties as discussed by Sand-Jensen and Nielsen in Chapter 2. These variations led us to expect that a shift in dominance from seagrasses and/or slow-growing macroalgae to more or less complete dominance by phytoplankton would increase the importance of within system remineralization and recycling of nutrients and reduce the importance of nutrient retention and removal through permanent burial and other processes. But is this so? 1.1 Eutrophication and plant responses A large number of abiotic and biotic factors contribute to determine the exact structure of plant communities in shallow coastal ecosystems, but the role of nutrients has received considerable attention in the last couple of decades due to the increasing problems related to coastal eutrophication. It is now clear that increasing nutrient richness of estuaries and other coastal marine areas – known as cultural eutrophication – leads to structural shifts in the plant assemblage, although it does not 281 S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling: The IInfluence of Primary Producers, 281-292. © 2004 Kluwer Academic Publishers. Printed in the Netherlands.

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necessarily lead to an increase in total primary production (Borum 1995; Borum & Sand-Jensen 1996; Schramm and Nienhuis 1996). The composition of coastal plant communities and the structural changes that follow eutrophication is dealt with in detail by Hauxwell and Valiela in Chapter 3. Seagrasses and large, slow-growing, perennial macroalgae (mainly brown algae) tend to dominate plant communities in shallow coastal ecosystems under pristine conditions. Although the total plant biomass in such systems may be completely dominated by slow-growing, perennial macrophytes, total primary production is often much more “diverse”, receiving significant contributions from other plant types such as ephemeral macroalgae, benthic microalgae and phytoplankton. This divergence emerges as a result of the inherent variation in physiological and ecological rates brought about by the different plant types - seagrasses and large, perennial macroalgae have lower specific growth rates and thus a lower productivity than ephemeral macroalgae and most types of microalgae (Borum & Sand-Jensen 1996; Nielsen et al., 1996). One of the primary effects of eutrophication in coastal marine areas is an increase in the contribution of ephemeral macroalgae, microphytobenthos and phytoplankton to total plant biomass and production, whereas the importance of slow-growing, perennial macrophytes decrease (Sand-Jensen & Borum, 1991). The mechanisms behind these structural changes are now relatively well understood - seagrasses and slow-growing, perennial macroalgae are superior competitors for nutrients due to the low nutrient requirements and a high capacity to store nutrients and these plants therefore dominate the plant community under nutrient poor conditions (Pedersen & Borum, 1996). Fast-growing macroalgae and microalgae have, in contrast, higher nutrient demands per unit biomass and time and are therefore stimulated by high nutrient supply. These plant types are superior competitors for light due to their thin thalli (Agustí et al., 1994) and because of their epiphytic, free-floating or pelagic nature, and these attributes allow them to overgrow benthic and slower growing primary producers in situations where nutrients are no longer limiting. Light limitation caused by shading from drifting macroalgae, epiphytes and/or phytoplankton seem to be the most important cause for the observed decline of seagrasses and other benthic macrophytes during eutrophication, but direct nitrogen toxicity (Burkholder et al., 1992; Touchette & Burkholder, 2002) and unfavorable changes in the biogeochemical environment of the seagrasses (e.g. Greve et al., 2003; Holmer & Bondgaard, 2001) have lately been brought forward as additional potential causes. Whether eutrophication of a shallow coastal ecosystem leads to final dominance of ephemeral macroalgae or by phytoplankton may depend on the resulting nutrient regime but also on the water residence time of the system in question since low residence time seems to favor dominance by ephemeral macroalgae whereas longer residence time is claimed to favor dominance by phytoplankton. Although the overall relationship between eutrophication and changes in plant composition is well documented through a large number of studies it should be remembered that these relationships are not as simple and predictable as often suggested in the literature. Several processes may either affect the composition directly or indirectly through feedback effects on the availability of nutrients. Substantial losses of seagrasses and other benthic macrophytes may, for example, increase the frequency and importance of resuspension events (Pedersen et al., 1995),

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which may in turn enhance the release of nutrients from the sediment and thus stimulate the productivity of fast-growing plant types even further (Duarte, 1995). Other processes, such as grazing, can delay or even prevent the accumulation of fastgrowing ephemeral plants and thus act as a buffer against one of the primary responses to nutrient enrichment. Selective grazing on specific plant types may, however, at the same time increase rates of remineralization and thus contributes to the stimulation of growth of other plant types that are not grazed to the same extent. Intense grazing on phytoplankton by benthic filter feeders, for example, may stimulate the productivity of ephemeral macroalgae and thus contribute to the regulation of the plant assemblage. The primary response to eutrophication may finally be influenced by the hydrography of the system in question since low water residence time and/or high physical exposure to waves may stimulate the export of certain plant types. The exact nature of these processes, and the extent to which they contribute to modify the structure of marine plant communities, needs further investigation. 2. EUTROPICATION THE OTHER WAY AROUND – EFFECTS OF PLANTS ON NUTRIENT DYNAMICS Marine plants and algae are often highly abundant in shallow coastal ecosystems and these systems rank among the most productive biomes in the world (Chapter 1). The preliminary analysis carried out in Chapter 1 showed that most of the nutrients received by coastal ecosystems from terrestrial and atmospheric sources are assimilated by the primary producers regardless of the structural composition of the plant community. Total assimilation seems therefore to depend mostly on total nutrient availability (i.e., loadings). The further fate of these nutrients must therefore be closely coupled to the fate of the primary production. Since the fate of the produced plant material depends to a large extent on the plant type in question, it can be expected that dominance of specific plant types influence whether nutrients are “lost” from the ecosystem through export and burial or whether they are remineralized and recycled to support further primary production within the system. 2.1 Mass transport and export While the massive assimilation of nutrients by plants may delay the horizontal transport and subsequent export of nutrients from the ecosystem, losses of plantbound nutrients through mass transport of plants themselves may prevent plant material and associated nutrients from accumulating within the system. The export of nutrients bound in plants and algae seems to be an important, but usually overlooked process. Flindt et al. (Chapter 4) reports a series of studies undertaken to elucidate the importance of nutrient losses from coastal ecosystems through mass transport of plant matter. These studies show that total export of plant-bound nutrients is highly variable among different systems but also that it can contribute substantially to the export of particle bound nutrients. Export of plant-bound nutrients thus represented 63% of the total N-export from Venice Lagoon, Italy, 21% of the total N-export from Roskilde Fjord, Denmark and less than 5% of the total N-export from Mondego estuary in Portugal. The substantial export of plant-bound nutrients observed in some systems suggests that this export may represent a significant percentage of the

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nutrients originally received by the system from external sources or assimilated by the primary producers. Unfortunately, very few studies allow such comparisons (see, however, Salomonsen et al., 1999 and Bergamasco et al., 2003 for examples). The absolute size of the export depends to a great extend on the hydrographical conditions of the system in question but may also depend upon the dominance patterns of the plant assemblages. Strong river influence and near unidirectional , flow of water seaward will, other things being equal, increase the importance of mass transport, although net transport will also depend upon the morphometry of the recipient. Plant material that is exported from the system will invariably be lost if the system opens on a deeper water body where the plant matter sink to greater depths, while some of the material may return when the tide turns or if the wind changes when the system opens on a shallow water body, thus reducing the net export. Different plant types show different patterns of transport. Attached macrophytes with roots or strong hold-fasts (i.e. seagrasses and most perennial, slow-growing macroalgae) are less susceptible to mass transport than most ephemeral macroalgae, which are easily detached from their substrate, and phytoplankton, which is suspended in the water column. Even strongly attached plants and algae may, however, occasionally become detached by storms, grazing and other types of physical disturbances. Furthermore, seagrass leaves are lost continuously as old leaves are shed as part of the growth process. When detached, most macroalgae moves as bedload along the bottom and different types of macroalgae appear to have more or less the same threshold for transport, measured as water velocity needed to keep them in suspension. The transport of macroalgae is therefore especially dependent upon current speed and direction. Sloughed seagrass leaves move, in contrast, primarily at the water surface, probably because of aerenchymatic air and the transport is therefore affected by both currents and wind speed and direction. All these observations taken together suggest that the absolute loss of nutrients due to mass transport should increase with eutrophication since this should lead to a decline of benthic, attached vegetation and an increase in the amount of free-floating, ephemeral macroalgae and suspended microalgae that are more prone to mass transport. The few studies available do, however, not allow a test of the hypothesis that fast-growing plant types lose a larger proportion of their net primary production than slow-growing, perennial plant types, since parallel estimates of production and export losses are rare. Comparative studies that compare the proportions of assimilated nutrients lost from coastal ecosystems dominated by different plant types should therefore be conducted in the future. 2.2 Grazing losses – consequences for nutrient mineralization The shift from dominance by seagrasses and other slow-growing perennial macroalgae to dominance by ephemeral macroalgae and phytoplankton during eutrophication is expected to affect the flux of nutrients to higher trophic levels since these changes also represent a shift from production of low quality food to production of high quality food (i.e. higher nutrient concentrations and lower content of structural compounds and defense chemicals). Since ecosystem processes such as grazing and decomposition, rely partly on these characteristics (e.g. Cebrián, 1999), it is expected that remineralization of nutrients through grazing and decomposition will change concomitantly with changes in plant composition.

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Petersen and Cebrián deal extensively with the potential role of grazing upon phytoplankton and benthic plants in Chapters 5 and 6, respectively. Benthic suspension feeders – by far the most important grazers on phytoplankton in shallow coastal ecosystems - can exert a strong top-down control on phytoplankton and essentially strip the water column for suspended primary producers. For example, the mean residence time for algal cells in the water above a dense populations of benthic suspension feeders can be as low as 1 hour (ranging from 1 hour to 2 days) in shallow areas. Whether or not this potential grazing impact is reached is partly under hydrographical control since benthic filter feeders will be able to control the phytoplankton biomass if their clearance time is lower than the water residence time, whereas they are unable to control the phytoplankton when the system is stratified or if the water residence time is lower than the clearance time. The grazing pressure on benthic primary producers is less intensive than on pelagic primary producers in most coastal ecosystems. Relative grazing intensity (i.e. percent of biomass or production grazed) is partly controlled by the nutrient content of the primary producers and grazing intensity is therefore lower for nutrient poor, perennial macrophytes than for ephemeral macroalgae and microalgae, which have more nutrient rich tissues. This pattern is similar to that found when grazing intensity is compared across a wider range of primary producers including both terrestrial vegetation as well as phytoplankton (Cebrián, 1999). The variations in grazing intensity indicate, at first glance, that more plant biomass is channeled to higher trophic levels when fast-growing, nutrient rich plants dominate the plant community, but Cebrián shows elegantly that this is not the case - the absolute amount of biomass (or carbon) lost to grazers per unit area and time is approximately the same for different plant groups. The explanation is that absolute grazing losses are positively correlated to absolute production and that slow-growing perennial macrophytes tend to accumulate larger biomass and produce more per unit area and time than faster growing plant types. Thus, the absolute amount of biomass (or carbon) that is grazed per unit area and time remains therefore more or less the same when compared across plant communities dominated by different plant groups. Eutrophication leads to a replacement of nutrient poor, slow-growing, perennial macroalgae by nutrient rich, ephemeral macroalgae and microalgae and grazing intensity therefore increases with eutrophication, whereas the total flux of biomass (carbon) remains approximately the same. Some field observations have, however, shown that that the flux of carbon to herbivores may decrease under severe eutrophication. This phenomenon may be due to some of the side effects of eutrophication, i.e., the replacement of seagrasses and large, perennial macroalgae by ephemeral macroalgae and microalgae which may represent a destruction of habitats to many benthic herbivores and the increasing production of easy decomposable organic matter which may increase the frequency of anoxia events. Both of these changes can have detrimental effects on herbivore density and activity. The absolute flux of nutrients that are channeled to higher trophic levels may, however, increase with eutrophication, although this aspect is not addressed specifically by Cebrián in Chapter 6. The N and P content of phytoplankton and benthic microalgae are typically 2-5 fold higher than in slow-growing macrophytes (Duarte, 1992; Cebrián, Chapter 6), so the flux of these nutrients to higher trophic levels and, hence, the absolute amount of remineralized nutrients, may increase during eutrophication, even if the total flux of carbon remain constant or even

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decreases slightly. Whether or not nutrient fluxes from primary producers to herbivores vary systematically across systems dominated by different plant groups warrants further investigation in the future. It is also unclear whether higher nutrient fluxes to herbivores will lead to higher rates of mineralization - grazing may either lead to mineralization and thus increase nutrient availability that can support primary production or herbivores may assimilate and store nutrients as biomass. Storage of nutrients in biomass is of course temporary, unless the herbivores in question are economically important and therefore removed from the system by harvest (e.g. mussels, oysters, prawns) or migrate themselves out of the ecosystem. Little is known about the balance between mineralization and assimilation but a larger proportion of nutrients should, other things being equal, be mineralized when nutrient-rich food items (i.e. ephemeral macroalgae and microalgae) are consumed. More studies that specifically address these questions are needed before firm conclusions regarding the quantitative role of grazing on different plant types on overall nutrient cycling can be drawn. 2.3 Decomposition and permanent burial Decomposition of plant litter represents yet another process that may affect the balance between accumulation and recycling of nutrients within shallow coastal ecosystems. Banta et al. show in Chapter 7 that patterns of decomposition differ markedly between different groups of primary producers although decomposition patterns do vary substantially within each group. Slow-growing macrophytes (especially seagrasses) decompose more slowly and less completely when compared to most macro- and microalgae. Especially important seems to be the proportion of the detritus that is resistant to fast decomposition (i.e. the refractory component) – this fraction is typically much larger in seagrasses, and especially so in their roots and rhizomes, than in marine macro- and microalgae. Surprisingly, detritus from phytoplankton also contain a significant refractory portion, although this observation may be in part an artifact due to methodological differences (i.e., litter-bag studies not being well-suited for the study of phytoplankton decomposition). The mineralization rates and patterns of N and P (relative to carbon decomposition) appear to be controlled by the initial nutrient status of the litter with nutrients that are limiting for decomposer organisms (usually N) being preferentially retained relative to carbon and non-limiting nutrients. Systematic variations in decomposition and mineralization rates and in the amount of refractory detritus among different plant groups may have important consequences for the retention of detritus and associated nutrients at the ecosystem level. Low rates of decomposition lead to long residence times for litter in the detritus pool, which suggests that detritus and associated nutrients from slow-growing macrophytes are more prone to permanent burial than detritus from fast-growing plants. Permanent burial is defined as the long-term (annual to decades) removal of nutrients from the pelagic component by accumulation of sediments (Middelburg et al., Chapter 8). Comparative studies based on carbon flows have shown that 10-17% of the net primary production from rooted, slow-growing macrophytes are buried whereas much less (about 1-5%) of the net primary production of algae are buried permanently. Although this difference may seem significant, it is worth remembering that the difference may be much smaller when expressed in units of nutrients (N or P)

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because slow-growing macrophytes typically contain fewer nutrients per unit biomass than fast-growing algae. Banta et al. present such a comparison of nutrient budgets for estuaries dominated by different primary producers at the end of Chapter 7. Their model simulations based on data from two shallow estuaries show that accumulation of detritus-bound nutrients is markedly higher in systems dominated by slow-growing macrophytes than in systems dominated by fast-growing macro- and microalgae. This is because total primary production is larger in systems dominated by the former plant group and because a larger proportion of that production ends up as detritus. Furthermore, the turnover rate of that detritus is much slower and less complete than for fast-growing plant types resulting in greater nutrient retention. Even when seagrasses dominate, however, permanent nutrient retention in detrital pools is only a minor fraction (about 10%) of the nutrients entering such systems. This retained fraction falls towards zero when faster growing plants dominate the community. One would immediately think that the main contribution of local primary producers to permanent burial in an ecosystem would be their own biomass, or rather, necromass. This is not necessarily the case, however. Studies using stable carbon isotopes have shown that a large proportion of the carbon that accumulates in many shallow coastal ecosystems is in fact of allochthonous (i.e., external) origin. The main contribution of local primary producers to permanent burial of plant detritus and associated nutrients may therefore rather be to reduce resuspension, dampen water movement and, thus, stimulate net sedimentation of allochthonous particulate matter. This process may have little effect on nutrient cycling in coastal ecosystems as these nutrients were already assimilated in detrital pools before they entered the system and were therefore unavailable for plant uptake. Benthic microalgae stabilize sediments through exudation of polymeric substances that may lower erodability and help keep detritus particles bound to the sediment surface. This effect is, however, seasonal and organic matter “trapped” by these polymeric substances may become resuspended later. The role of benthic microalgae for burial of organic matter and associated nutrients is therefore most probably of minor importance over annual time scales. Perennial, rooted macrophytes, on the other hand, have a stronger and longer-lasting stabilizing effect on the sediment. In addition, rooted macrophytes increase sedimentation of suspended particulate matter. These effects of reducing resuspension and increasing deposition lead to enhanced accumulation of sediments including organic matter and associated nutrients within plant meadows. Structural changes in the plant community that lead to the loss of rooted macrophytes may therefore have important consequences for the long-term retention of major nutrients through burial and preservation. 2.4 Benthic nutrient cycling Risgaard-Petersen and McGlathery et al. provide excellent analyses of the potential effects of different plant groups on benthic nutrient cycles in Chapters 9 and 10. Benthic processes are important because they affect the concentration of porewater nutrients and, thus, the benthic-pelagic flux of nutrients. The flux of nutrients to the water phase may, in the long run, determine whether nutrients released through decomposition remain “trapped” in the sediment and thus become prone to permanent burial or, whether they are released again to the overlying water phase and become

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available for assimilation and/or export once again. Benthic processes may finally influence whether nitrogen becomes lost permanently from the ecosystem through the process of denitrification. Four processes seem especially important for regulating the availability of nutrients in sediments: 1) mineralization, where dissolved inorganic nutrients are “produced” from organic detritus, 2) coupled nitrification and denitrification where ammonia is transformed via nitrate to free N2 that is lost to the atmosphere, 3) N-fixation, by which gaseous nitrogen is incorporated into organic matter and thereby made biologically available and 4) chemical binding of P to iron ferrihydroxides and in other complexes that may reduce porewater concentrations of dissolved P and fluxes to the water phase and may stimulate the permanent burial of P. Risgaard-Petersen and McGlathery et al. (Chapters 9 and 10) show through a large number of examples that most benthic, marine plant groups can affect these processes, but the question is whether some plant types assert stronger effects than others and, what the net effects are. Mineralization is typically larger when slow-growing, perennial plants dominate because absolute production, deposition (from allocthonous sources) and decomposition of detritus is larger per unit area and time in systems dominated by these plant types, despite the fact that mineralization rates per unit organic matter are lower (Banta et al. Chapter 7). Rooted macrophytes may further stimulate decomposition of slowly decomposable detritus through leakage of oxygen in the rhizosphere and rates of mineralization are therefore often higher in sediments covered by seagrasses when compared to bare sediments. This does not necessarily mean that the flux of nutrients to the water is higher in seagrass meadows, however, because mineralization is often closely coupled to assimilation by plant roots in productive seagrass beds. A close coupling between mineralization and assimilation also exists in dense populations of benthic microalgae and within dense mats of floating macroalgae although rates are typically lower. Nitrification and denitrification affect the concentrations of N species and may lead to permanent losses of nitrogen from the ecosystem. Both nitrification and denitrification are profoundly affected by the presence of plants. Several studies have shown that rooted plants and benthic microalgae both can stimulate nitrification through exudation of oxygen to anoxic sediments. It has been suggested that the presence of seagrasses should stimulate coupled nitrification-denitrification, but more recent studies have shown that this is rarely the case. Rates of coupled nitrificationdenitrification are in fact typically lower in areas vegetated by benthic microalgae and seagrasses due to plant uptake. Both benthic microalgae and seagrasses seem to be better competitors for ammonium than nitrifying bacteria and coupled nitrificationdenitrification may thus be reduced by 50-75% in the presence of benthic microalgae and seagrasses. A positive effect of benthic microalgae and seagrasses on coupled nitrification-denitrification may, however, occur under N-replete conditions where plants and nitrifiers do not compete for ammonium. Little is known about the effect of macroalgae on nitrification-denitrification but the presence of dense mats of floating macroalgae often leads to anoxic conditions in the sediment and the zone of nitrification/denitrification may move up into the algal mat temporarily. In such cases, nitrification-denitrification is usually reduced again due to competition for nutrient uptake with the macroalgae There is little consensus about the potential effects of different primary producers on N-fixation. Some studies have shown that N-fixation rates can be substantially higher

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in sediments vegetated by seagrasses while other studies have not been able to demonstrate such an effect. The relatively few reported rates on N-fixation in seagrass beds show that rates are approximately within the same range as rates of denitrification. Rates of N-fixation and denitrification are, however, relatively low when compared to other N-transport rates in shallow coastal ecosystems, so the effects of seagrasses on these processes have relatively little ecological significance. The release of oxygen to the sediment by benthic microalgae and rooted plants, or the depletion of oxygen by dense mats of floating macroalgae or benthic microalgae (in the dark), may dramatically affect redox conditions and thus benthic P-dynamics. Stimulation of oxidized conditions in the uppermost sediment favors the binding of P to oxidized iron compounds and thus lowers both the concentrations of dissolved P in the porewater and reduces the flux of P to the water column. In contrast, lower redox conditions during oxygen depletion have the opposite effect and may stimulate the release of P to the water phase. The strongest effect that plants may exert on benthic nutrient dynamics may be through their direct interference with nutrient fluxes across the sediment-water interface. Seagrasses and benthic microalgae stabilize the sediment and thus reduce the release of sediment nutrients to the water via resuspension events. Both these plant groups are further able to intercept the flux of dissolved nutrients from the sediment to the water through their own assimilation. Especially benthic microalgae seem potentially important in this respect, since dense mats of these algae can prevent benthic-pelagic fluxes of nutrients completely. The efficiency of benthic microalgae to assimilate nutrients may vary substantially, however, over diurnal and longer time scales since the community must be net autotrophic in order to be able to assimilate enough nutrients to prevent diffusive transport of nutrients to the water. There are many temporarily variable factors that affect benthic algal production and thus their ability to assimilate nutrients and function as a filter at the sediment-water interface. The temporal and spatial variations in the importance of these algae for nutrient release require further studies. Seagrasses may also reduce the flux to the water and lower porewater nutrient concentrations through assimilative uptake of nutrients via the roots, but again, no studies allow a direct comparison between the effect of seagrasses and other plant types. Finally, dense mats of floating macroalgae may have the same nutrient filtering effect as benthic macroalgae. Experimental studies have shown that dense mats of macroalgae are able to adsorb nutrients released from sediments completely, but such mats are often highly “ephemeral” in time and space, and therefore most probably not very efficient filters when evaluated over larger temporal and spatial scales. The different marine plant groups each have the potential to influence benthic nutrient cycling though a number of mechanisms, but quantitative studies that allow direct comparisons of the net effects of theses different plant types on an ecosystem basis are sorely lacking.

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3. CONCLUSION When we first embarked on the work behind this book (see Chapter 1), we were rather convinced that the observed changes in dominance by different plant types, observed as a consequence of coastal eutrophication, would lead to predictable changes in the cycling and ultimately the fate of nutrients in these ecosystems. We were also convinced that these relationships could be used to predict changes in cycling and fate of nutrients caused by changes in the composition of coastal marine primary producers brought about by other factors than eutrophication. After our own studies and with the contributions of our colleagues, we realize that the world may be slightly more complicated. We had hoped to be able to give typical rates for the various main processes (export, grazing, decomposition etc.) under dominance by seagrasses, slow and fast growing macroalgae and phytoplankton, but this is hardly possible, due to a lack of data. Instead, we can attempt to assess the constancy and importance of these processes. Mass transport of plant-bound nutrients out of the estuarine systems (export) is highly variable between different systems, but can undoubtedly be of huge importance in systems where export of plant-bound nutrients constitute a large percentage of the total nutrient export. Unfortunately, comparative studies where export is measured together with primary production are largely absent and it is therefore difficult, if not impossible, to assess the significance of export of plant-bound nutrients relative to the primary production. It is equally difficult to conclude on how export may change as a consequence of changes in the plant community. Grazing has been shown to be of a constant magnitude, when measured in biomass or carbon units. As ephemeral macroalgae and phytoplankton have higher contents of nutrients in their tissue than seagrasses and slow-growing macroalgae, however, this should mean that the amount of nutrients going through the grazingfood chain will increase if ephemeral macroalgae or phytoplankton replace seagrasses or fucoid macroalgae. Unfortunately, data to confirm this are largely lacking. Probably most surprising to us were the results relating to the decomposition of various plant types. That below-ground parts of seagrasses decomposed slowly and with a large refractory component was to be expected, but that phytoplankton also had a large refractory component was more surprising, given their chemical composition, although this may be a experimental artifact and needs further consideration. Large differences between different types of macroalgae existed, but their overall contribution to nutrient retention via refractory pools were minor. Refractory detrital pools can ultimately give rise to a loss of nutrients from the system through permanent burial, and indeed data show that the percentage of the primary production that is lost to permanent burial is significantly larger when the system is dominated by slow growing primary producers. However, the percentage lost to permanent burial is always a minor fraction of the primary production and as such of minor significance in overall nutrient budgets. Another loss process is denitrification. Data indicate that denitrification varies with several orders of magnitude between different estuarine systems, and that primary producers have a tendency to lower the rates of denitrification through competition with the bacteria for substrate (i.e., NO3-). Unfortunately data do not allow comparisons between systems dominated by different types of primary producers to allow us to asses the importance of this loss term in the context of now marine plants alter nutrient cycling

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At the onset of this work, we were convinced that the composition of the plant community was paramount in determining cycling and fate of nutrients in the coastal marine ecosystems. This work has shown, however, that other factors are of equal importance. Most important of these factors are undoubtedly the hydrography and the nutrient loading of the system in question. Hydrography (morphometry, water residence time) can be decisive in determining which plant type will dominate, the degree of export of plant-bound nutrients and the importance of various grazers. Nutrient loading in itself is probably of equal significance as the plants in determining the various fluxes as well as the ultimate fate of the nutrients. However, the plant component of coastal ecosystems is still very important as virtually all nutrients pass through the primary producers at least once on their way from the terrestrial to the oceanic ecosystems. On the other hand, the role of the estuary as a filter is questionable. Our data show that from 1 to 17 % of the primary production becomes permanently buried, depending on the primary producers that dominate. These numbers, however, assume steady-state conditions, and the question is to which extent steady-state conditions prevail in costal marine ecosystems, that could be seen as going through cycles of eutrophication or other detrimental anthropogenic influence, followed by restoration initiatives. As it can be inferred from the above, many data are still lacking, but we are now in a better situation than before, as we are now able to describe and model the individual processes involved in coastal nutrient cycling (the arrows in our conceptual model, Fig. 4 in Chapter 1) mechanistically correct. Models of these individual processes should now be synthesized in dynamic ecosystem models and validated by empirical studies in comparable ecosystems, where all individual processes are compared and related to total primary production and the contributions from the different plant types. In these studies nutrient loading and hydrography should be included as important controlling parameters.

4. REFERENCES Agustí, S., Enríquez, S., Frost-Christensen, H., Sand-Jensen, K., & Duarte, C. M. (1994). Light harvesting among photosynthetic organisms. Functional Ecology, 8, 273-279. Bergamasco, A., De Nat, L., Flindt, M.R. & Amos, C.L. (2003). Interactions and feedbacks among phytobenthos, hydrodynamics, nutrient cycling and sediment transport in estuarine ecosystems. Continental Shelf Research 23, 1715-1741. Burkholder, J. M., Mason, K. M., & Glasgow, H. B. (1992). Water-column nitrate enrichment promotes decline of eelgrass Zostera marina: evidence from seasonal mesocosm experiments. Marine Ecology Progress Series, 81, 163-178. Cebrián, J., & Duarte, C. M. (1994). The dependence of herbivory on growth rate in natural plant communities. Functional Ecology, 8, 518-525. Cebrián, J., & Duarte, C. M. (1995). Plant growth-rate dependence of detrital carbon storage in ecosystems. Science, 268, 1606-1608. Cebrián, J., & Duarte, C. M. (1998). Patterns in leaf herbivory on seagrasses. Aquatic Botany, 60, 67-82. Cloern, J. E. (2001). Our evolving conceptual model of the coastal eutrophication problem. Marine Ecology-Progress Series, 210, 223-253. Duarte, C. M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia, 41, 87-112. Duarte, C. M., Benavent, E., & Sanchez, M. d. C. (1999). The microcosm of particles within seagrass Posidonia oceanica canopies. Marine Ecology Progress Series, 181, 289-295. Enríquez, S., Duarte, C. M., & Sand-Jensen, K. (1993). Patterns in decomposition rates among photosynthetic organisms: The importance of detritus C:N:P content. Oecologia, 94, 457-471.

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Greve, T. M., Borum, J., & Pedersen, O. (2003). Meristematic oxygen variability in eelgrass (Zostera marina). Limnology and Oceanography, 48(1), 210-216. Hauxwell, J., Cebrían, J., & Valiela, I. (2003). Eelgrass Zostera marina loss in temperate estuaries: relationship to land-derived nitrogen loads and effect of light limitation imposed by algae. Marine Ecology Progress Series, 247, 59-73. Holmer, M., & Bondgaard, E. J. (2001). Photosynthetic and growth response of eelgrass to low oxygen and high sulfide concentrations during hypoxic events. Aquatic Botany, 70, 29-38. Kennedy, V. S. (Ed.). (1984). The estuary as a filter. Orlando: Academic Press. Krause-Jensen, D., McGlathery, K., Rysgaard, S., & Christensen, P. B. (1996). Production within dense mats of the filamentous macroalga Chaetomorpha linum in relation to light and nutrient availability. Marine Ecology Progress Series, 134, 207-216. Littler, M. M., & Littler, D. S. (1980). The evolution of thallus form and survival strategies in benthic marine macroalgae: Field and laboratory tests of a functional form model. American Nauralistt, 116(1), 25-44. Nielsen, S. L., Enríquez, S., Duarte, C. M., & Sand-Jensen, K. (1996). Scaling maximum growth rates across photosynthetic organisms. Functional Ecology, 10, 167-175. Nielsen, S. L., & Sand-Jensen, K. (1990). Allometric scaling of maximal photosynthetic growth rate to surface/volume ratio. Limnology and Oceanography, 35, 177-181. Pedersen, M. F., & Borum, J. (1996). Nutrient control of algal growth in estuarine waters. Nutrient limitation and the importance of nitrogen requirements and nitrogen storage among phytoplankton and species of macroalgae. Marine Ecology Progress Series, 142, 261-272. Pedersen, O., Borum, J., Duarte, C. M., & Fortes, M. D. (1998). Oxygen dynamics in the rhizosphere of Cymodocea rotundata. Marine Ecology Progress Series, 169, 283-288. Pedersen, O. B., Christiansen, C., & Laursen, M. B. (1995). Wind-induced long term increase and short term fluctuations of shallow water suspended matter and nutrient concentrations, Ringkøbing Fjord, Denmark. Ophelia, 41, 273-287. Salomonsen, J., Flindt, M.R., Geertz-Hansen, O. & Johansen, C. (1999). Modelling advective transport of Ulva lactuca (L) in the sheltered bay, Møllekrogen, Roskilde Fjord, Denmark. Hydrobiologia 397, 241-252. Sand-Jensen, K., & Borum, J. (1991). Interactions among phytoplankton, periphyton, and macrophytes in temperate freshwaters and estuaries. Aquatic Botany, 41(1-3), 137-175. Schramm, W., & Nienhuis, P. H. (1996). Introduction. In W. Schramm & P. H. Nienhuis (Eds.), Marine benthic vegetation. Recent changes and the effects of eutrophication. (pp. 1-6). Berlis, Heidelberg, New York: Springer. Terrados, J., Duarte, C., Fortes, M., Borum, J., Agawin, N., Bach, S., et al. (1998). Changes in community structure and biomass of seagrass communities along gradients of siltation in SE Asia. Estuarine, Coastal and Shelf Science, 46, 757-768. Touchette, B. W., & Burkholder, J. M. (2002). Seasonal variations in carbon and nitrogen constituents in eelgrass (Zostera marina L.) as influenced by increased temperature and water-column nitrate. Botanica Marina, 45, 23-34. Worm, B., L., L. H., Boström, C., Engkvist, R., Labanauskas, V., & Sommer, U. (1999). Marine diversity shift linked to interactions among grazers, nutrients and propagule banks. Marine Ecology Progress Series, 185, 309-314.

5. AFFILIATIONS S.L. Nielsen, M.F. Pedersen & G.T. Banta: Department of Life Sciences and Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark

SUBJECT INDEX

A/V-ratio;28;29;34;35;38;39;40;44;45 Advection;2;8;13;17;31;112;146;159;345 Advective transport;2;8;13;17;31;112;146;159;345 Aerenchym;110;120;143;261;335 Aerobic;11;56;94;96;106;231;273;308;310;315 Algae Brown;23;24;28;44;332 Green;23;24;52;288 Red;24;113;114;121 Unicellular;1;23;37;43;44;59 Algal blooms;77;83;103;154;190;193;196;197;207;253 Algal mats;90;93;94;195;197;200;208;209;264;302;304;340 Allochthonous;10;263;292;338 Amino acids;69;282;286 Ammonification;11;303;317;329 Amphipods;183;199 Anaerobic;9;13;94;96;168;231;232;234;240;250;262;267;308;317;329 Angiosperm;1;23;25;26;60;63;159;250;300;322;323 Animals;9;22;31;32;62;98;149;150;155;156;157;169;175;199;258;266;267;272 Anoxia;13;17;53;94;96;98;101;105;112;166;180;192;193;196;197;199;201;202;233; 249;250;253;254;257;261;263;267;274;281;282;283;288;290;294;298;303;311; 312;315;319;320;323;326;336;340 Anthropogenic influence;76;344 Ascidians;149;150;159;174 Autochonous;10 Autotrophic;vii;1;2;4;7;14;15;60;101;204;205;214;250;265;270;272;275;279;280;284 ;285;294;298;302;303;304;322;328;341 Autotrophy;280;285;294 Bacteria;11;30;31;47;48;56;69;73;149;208;236;240;260;263;274;275;282;283;284; 292;293;301;306;308;312;315;322;323;324;325;326;329;340;344 Bathymetry;3;132;155;163 Bedload;108;110;114;115;135;142;335 Benthic;v;vii;1;9;10;11;12;16;17;21;22;25;33;47;48;51;52;53;54;56;57;58;59;61;62; 66;68;73;85;88;101;102;107;108;126;132;135;147;149;150;151;152;153;154;155; 156;157;158;159;160;162;163;164;165;166;167;168;169;171;172;173;174;175; 176;177;178;179;180;181;182;183;184;185;186;187;190;195;196;197;199;200; 202;203;204;205;208;209;210;241;243;249;251;258;261;263;264;265;266;267; 270;271;272;274;276;279;280;281;283;284;285;286;287;288;289;290;291;292; 293;294;295;296;297;298;299;300;301;303;304;305;306;308;310;312;314;318; 319;320;322;323;325;326;327;328;329;331;332;333;335;336;337;339;340;341; 345;346 Benthic diatoms;62;258;266;291;294;305 293

294

Benthos;52;54;103;170;173;175;209;283;293;294 Biofilm;258 Biogeochemistry;6;18;60;71;86;93;94;96;98;99;254;260;261;267;271;283;297;315; 333 Bioirrigation;254 Biomass;2;3;4;7;8;12;13;15;16;17;21;22;31;33;34;35;38;41;42;43;46;47;50;51;53;55; 56;57;58;61;64;73;88;89;90;91;98;99;100;101;102;104;117;118;119;121;122;124; 125;126;127;129;130;135;146;152;153;154;156;159;160;161;162;163;164;165; 167;169;170;171;172;173;174;176;177;178;179;180;181;182;183;185;187;193; 199;200;201;202;203;204;206;207;209;214;216;217;219;227;231;232;236;240; 242;243;244;245;250;258;260;263;265;270;273;274;275;279;282;283;285;286; 287;288;289;290;292;293;295;305;323;326;328;331;332;336;337;338;343;346 Biome;2;214;333 Bioturbation;75;254;257;314 Bivalves;21;105;149;150;152;155;156;161;168;169;170;171;172;173;174 Bottom-up control;153 Boundary layer;29;30;31;32;33;40;54;56;58;59;61;62;89;155;156;157;158;159;160;171;173; 175;258;266 Brown tide;78;88;104;106;296 Buoyancy;110;119 Burial;v;6;10;11;12;13;54;99;234;238;242;247;248;253;254;255;260;262;263;264; 267;270;274;290;293;295;331;334;337;338;339;343 C/N/P-ratio;7;236 C/N-ratio;237 C/P-ratio;237;238 Carbonates;15;53;56;62;69;103;267;272;274;275;276;279;280;281;287;297;300;301; 302;304 Carotenoids;53 Carrying capacity;70;151;160;161;163;171;172 Catchment area;5 Cellulose;216;229 Chlorophyll;5;36;38;49;50;51;52;54;55;56;57;69;83;88;100;133;141;153;154;155; 157;158;159;164;165;288 Clearance rate;149;151;165 Clearance time;162;163;164;336 Colonial algae;37 Competition;3;47;61;85;98;101;168;206;207;211;274;280;287;288;289;321;324;329; 340;344 Copepods;168;175;182;197;204;205;206;292 Corals;68;76;103 Decay;70;90;215;216;231;233;235;256;258;279;290 Decomposition;v;2;6;7;8;9;10;13;14;22;26;50;53;69;99;145;147;208;214;215;216; 217;219;221;227;228;229;230;231;232;233;234;235;236;237;238;239;240;241; 242;243;244;246;247;249;250;251;252;254;265;267;276;281;284;286;290;298; 305;308;335;337;338;339;340;343;345 Denitrification;v;vii;6;11;13;17;53;168;242;253;261;267;271;272;274;275;276;277; 279;281;283;295;296;297;298;300;301;302;303;304;305;306;308;310;311;312;

295

313;314;315;316;317;318;319;320;321;322;323;324;325;326;327;328;329;339;3 40;341;343 Detritus;7;8;9;10;13;14;17;47;59;108;145;146;159;163;166;176;192;214;215;216; 217;219;227;230;231;232;234;235;236;237;238;239;240;241;242;243;244;245; 246;247;248;249;250;251;254;265;279;284;292;293;297;298;302;337;338;339; 340;343;345 DIC;279;283;291;295 Diffusion;30;31;32;40;89;94;173;254;261;281;282;286;312;320 DIN;133;134;141;142;282;283;286;289;291;292;296;321 Dissolved Inorganic Carbon;279;283;291;295 Dissolved Inorganic Nitrogen;133;134;141;142;282;283;286;289;291;292;296;321 Dissolved Organic Carbon;290;292;305 Dissolved Organic Matter;273;274;275;276;290;291;292;293 Dissolved Organic Nitrogen;138;290;291;292;306 DNRA;317;318 DOC;290;292;305 DOM;273;274;275;276;290;291;292;293 Dominance;vii;9;10;12;13;168;215;241;242;244;270;287;289;331;333;334;335;343 DON;138;290;291;292;306 Drift;21;23;25;26;47;76;82;98;101;103;109;110;112;115;193;196;197;198;207;208; 304;333 Dry weight;5;201;215;217;227;240 Ecosystem;v;vii;1;2;3;4;5;6;7;8;9;10;12;13;14;15;17;20;21;25;54;58;59;61;62;64;65; 66;68;73;75;76;77;85;93;99;101;102;103;104;106;109;142;144;150;151;154;155; 158;161;162;164;165;166;167;171;172;173;174;178;190;192;196;202;203;204; 205;206;211;214;215;218;231;234;240;241;242;247;249;251;252;253;254;256; 258;260;264;266;268;271;276;279;280;282;283;295;296;297;299;300;305;310; 328;331;332;333;334;335;336;337;338;339;340;341;342;343;344;345 Epiphytes;26;64;65;73;82;85;87;89;92;93;100;101;102;103;105;106;111;181;183; 190;207;275;276;287;290;293;333 Erosion;73;110;113;115;116;117;118;119;121;124;125;135;258;259 Estuaries;v;15;16;17;21;41;42;61;63;64;65;66;68;71;74;75;77;83;85;87;88;90;91;92; 93;94;96;97;100;101;102;103;105;106;108;109;110;112;113;115;116;119;135;141 ;142;143;144;145;146;149;152;156;159;161;162;163;164;165;168;169;170;171; 172;173;174;175;201;203;204;205;206;207;209;210;211;218;243;266;270;279; 284;292;295;298;300;303;305;310;313;317;325;326;327;328;329;332;334;338; 344;345 Eutrophication;vii;1;4;14;16;17;47;59;63;66;68;76;83;101;103;104;105;106;144;145; 146;153;154;155;159;160;164;166;169;170;171;173;177;179;180;186;187;189; 190;191;195;198;199;200;201;202;203;204;205;206;207;208;209;210;241;242; 253;266;293;294;296;297;298;305;331;332;333;335;336;337;343;344;345;346 Evolution;16;24;262;345 Excretion;9;153;160;166;167;176;291;292 Export;6;7;8;12;13;99;108;109;112;113;116;122;125;127;131;133;134;136;138;139; 141;142;143;144;147;204;265;272;293;297;333;334;335;339;343;344 Exudation;176;264;291;292;299;339;340 Faecal pellets;308 Filter effect;284;286

296

Filter feeders;149;172;173;174;333;336 Filtration;149;150;151;152;153;155;156;157;159;167;172;341 Fish;64;68;73;78;101;105;169;172;182;190;191;199;204;205;206;207;208;209;294; 317 Food limitation;151;160 Fucoid;8;9;44;76;224;343 Gastropods;153;193;206 Gene flow;24 Genetic diversity;25 Geology;5;77;85 Geometry;22;192 Ghost tissue;119;123 GPP;291 Grazing;v;2;6;7;8;9;12;13;15;21;22;31;34;48;50;51;52;54;55;60;61;81;98;103;104;10 5;117;122;123;124;125;126;127;129;130;131;145;149;150;152;153;154;155;156;1 60;164;165;166;168;169;171;173;174;175;176;178;186;203;204;205;206;207;208; 209;210;211;212;214;243;258;292;293;296;301;308;333;335;336;337;343 Green tides;76;101 Growth rate;2;14;31;35;38;41;42;43;44;45;46;50;52;55;59;63;64;65;88;94;119;121;123; 124;125;126;127;129;140;142;152;153;157;165;170;171;183;204;208;210;288; 291;332;345 Habitat;20;22;24;25;27;45;49;52;53;54;56;66;68;75;76;88;101;126;135;150;153;155; 173;180;190;191;195;204;205;206;208;209;210;212;259;270;303;336 Herbivore;9;12;15;104;176;178;179;180;181;182;183;184;185;187;189;197;199;200; 202;207;208;210;211;336;337 Herbivory;3;9;14;15;16;59;75;99;100;176;177;178;179;180;181;182;183;184;185; 186;187;189;190;195;196;197;198;199;200;201;202;204;205;206;207;210;211; 212;345 Hold-fast;8;334 Hydrodynamic forces;124 Hypoxia;94;102;174;192;193;197;199;205;210;345 Infauna;150;152;157;190;192;196;208;209;258;281;299;303;329 Isometry;28;40;44 Isopods;124;183;199 Kelp;8;9;61;62;66;159;178;181;204;206;208;211 Labile matter;215 Land use;5;83;177 Leaching;215;228;229 Light;2;3;15;16;20;21;22;25;28;31;34;35;36;38;42;46;47;48;49;50;51;52;53;54;55;56 ;57;58;59;60;61;62;63;64;65;73;78;85;87;88;89;90;91;92;93;94;96;100;101;102; 103;104;108;121;122;129;132;135;140;152;164;165;168;174;177;184;186;192; 195;196;199;206;211;218;249;263;272;273;274;279;280;281;282;284;286;287; 288;289;291;292;294;295;299;300;319;320;321;322;324;326;328;332;333;344; 345 Light attenuation;47;49;50;51;56;58;63;88;89;90;92;100;177;289;294 Lignin;13;216;229;263;265;300 Linear dimension;27;28;30;40

297

Litter;214;215;216;217;219;227;231;243;260;263;267;337;338 Macroalgae;vii;1;3;4;7;8;10;11;12;15;16;17;21;25;26;27;28;29;33;34;36;37;38;39;40 ;41;42;43;44;45;46;47;48;49;50;51;55;56;60;61;63;64;66;68;73;74;76;82;83;85;87 ;89;90;91;92;93;94;96;97;100;102;103;104;105;106;108;109;110;112;113;114; 115;116;117;118;119;120;121;122;124;127;129;130;131;133;135;140;142;144; 145;146;147;176;177;178;179;180;181;182;184;185;186;187;189;190;191;192; 193;195;196;197;198;199;200;201;202;203;204;205;206;207;208;209;210;211; 212;214;219;223;224;227;228;229;230;231;232;233;234;235;236;239;241;243; 244;245;246;247;249;250;251;267;270;271;276;280;281;282;283;284;285;286; 288;289;290;291;292;293;294;295;296;298;299;300;301;302;303;304;305;331; 332;333;335;336;337;340;341;343;345 Macrofauna;196;249;251;293;298 Macrophyte;vii;1;2;3;4;7;8;9;10;11;12;15;17;18;21;27;30;33;42;45;46;48;49;51;52; 54;55;56;57;58;64;65;88;100;101;102;103;105;106;108;109;110;112;113;116;119 ;121;122;126;131;139;141;142;147;154;204;210;214;218;219;227;229;230;231; 233;234;235;238;240;241;242;243;244;245;246;247;248;249;250;251;258;259; 260;261;263;264;265;268;284;289;296;302;303;306;323;328;332;333;334;336; 337;338;339;340;345 Mangroves;98;250;259;264;267;268;300 Meiofauna;293 Mesocosms;75;86;87;100;105;106;153;155;164;167;168;171;174;187;206;208;210; 211;345 Microalgae;vii;1;3;4;7;10;11;12;17;21;22;25;26;27;28;29;30;31;33;36;38;41;44;45;4 6;47;48;51;52;53;54;55;56;57;65;68;73;74;88;176;177;178;179;183;188;190;203; 214;219;222;227;228;229;230;231;235;236;241;243;244;245;246;247;249;258; 261;262;263;264;270;271;272;274;276;279;280;282;284;285;286;287;288;289; 291;292;293;295;296;298;301;303;304;318;319;320;322;328;329;331;332;335; 336;337;338;339;340;341 Microenvironment;31;93;318 Microphytobenthos;10;11;25;33;47;48;52;53;54;56;68;73;88;176;177;178;179;180; 181;182;183;184;185;187;188;189;191;192;200;202;203;241;243;258;261;263; 267;270;271;272;274;276;279;280;281;284;285;287;288;289;291;292;293;298; 301;303;304;305;318;319;320;322;324;328;329;332;337;339;340;341 Model;7;8;16;21;44;49;51;54;58;59;60;66;91;101;106;109;113;119;125;127;129;131 ;145;147;150;151;152;157;163;164;165;166;167;168;170;171;172;173;174;204; 211;215;216;217;232;233;235;250;252;264;265;267;268;270;289;293;296;312; 313;314;315;326;328;338;344;345 Morphological;1;2;9;25;114;119;231 Morphology;25;63;106;121 Mortality;61;94;96;193;199;210;297 Multi-G models;215;216 Mussels;48;64;98;149;150;151;153;154;155;157;158;159;161;164;165;166;167;168; 169;170;171;172;173;174;175;317;337 N/P-ratio;236;238 Net Primary Production;291 N-fixation;5;11;339;340 Nitrate;96;308;310;311;312;313;314;317;318;319;320;321;324;325;326;328;344 Nitrification;vii;11;13;17;53;261;272;274;279;281;283;296;301;302;303;308;310;

298

311;314;315;316;317;320;321;322;323;324;325;326;327;328;329;339;340 Nitrogen;v;1;3;4;5;7;11;12;14;15;16;17;20;21;33;39;40;41;42;43;50;53;59;60;61;62; 63;64;65;66;68;69;70;71;72;74;75;76;77;79;82;83;85;86;87;88;89;90;92;93;94;95; 96;97;98;100;101;102;103;104;105;106;107;108;109;131;132;133;134;135;136; 138;139;140;141;142;144;145;146;147;163;166;167;168;170;171;172;174;175; 177;183;184;185;186;189;191;198;199;201;203;204;205;206;207;208;209;210; 211;212;214;215;218;221;223;224;225;228;229;230;231;233;234;235;236;237; 238;239;240;243;244;246;247;248;250;251;263;264;265;266;267;268;270;272; 274;275;276;277;279;280;281;282;283;284;285;286;287;288;289;290;291;292; 294;295;296;297;298;299;300;301;302;303;304;305;306;308;310;311;312;314; 315;317;319;320;322;323;324;325;326;327;328;329;330;333;334;337;338;339; 340;345;346 Nitrogen fixation;69 Nutrient assimilation;6;7;215;253;271;283;290;322 Nutrient cycling;vii;5;6;12;13;71;92;145;168;174;202;218;241;242;249;253;270;271;272; 273;274;280;281;299;300;303;337;339;341;344 Nutrient retention;10;144;218;219;230;234;240;241;242;248;289;295;331;338;343 Nutrient supply;20;26;33;40;76;153;161;288;332 Nutrients;v;vii;1;2;3;4;5;6;7;8;9;10;11;12;13;14;15;16;17;18;20;21;22;25;26;28;29; 30;31;33;34;36;38;40;41;42;43;44;45;46;47;48;52;53;54;55;57;58;59;60;61;62;63; 64;66;68;69;71;73;75;76;77;79;81;83;85;87;89;92;93;96;97;98;99;100;101;102; 103;104;105;106;108;109;116;127;129;131;132;133;134;135;136;139;140;141; 142;143;144;145;146;147;152;153;160;161;164;165;166;167;168;169;171;172; 174;176;177;179;180;183;184;186;187;189;190;195;198;199;200;201;202;204; 205;206;207;208;209;210;211;212;214;215;216;217;219;221;227;230;234;235; 236;238;239;240;241;242;243;247;248;249;250;251;252;253;254;255;258;259; 260;263;264;265;266;267;270;271;272;273;274;276;279;280;281;282;283;284; 285;286;287;288;289;290;291;292;293;294;295;296;298;299;300;301;302;303; 304;305;306;310;317;319;320;322;325;326;327;331;332;333;334;335;336;337; 338;339;340;341;343;344;345;346 Organic acids;11;56;276 Organic matter;1;6;10;16;26;35;47;49;53;58;70;131;150;157;166;168;169;171;174;214; 216;218;227;228;229;230;231;234;241;252;254;255;260;261;262;263;264;265; 266;267;268;270;272;273;274;276;279;281;283;291;293;294;301;302;305;323; 325;329;336;339;340 Overshadowing;25;26 Oxygen;11;13;17;20;26;31;33;49;52;53;54;56;57;58;59;61;65;89;94;96;97;100;101; 102;115;116;119;144;147;160;161;192;193;195;197;206;261;262;265;266;268; 272;273;274;275;279;281;282;283;285;288;290;298;303;305;306;308;310;311; 312;313;315;317;318;319;320;321;322;323;324;326;327;329;340;341;345 Particle load;149 Pelagic;v;1;9;10;12;63;135;146;149;154;161;166;167;170;171;172;173;175;254;283; 292;294;301;308;317;325;326;331;332;336;338;339;341 Perennial macroalgae;215;234;240;270;331;332;335;336 pH;11;20;31;54;56;57;58;276;279;281;295;319;320;321;322 Phenol;9

299

Phosphorus;v;1;4;5;7;11;12;14;15;16;17;18;21;31;43;50;51;54;56;58;59;60;61;62;63; 64;65;66;69;71;88;100;101;102;103;104;105;106;107;108;111;131;132;133;134; 135;138;139;140;141;142;144;145;146;147;170;171;172;173;174;175;179;180; 183;184;185;189;191;203;204;205;206;207;208;209;210;211;214;215;218;221; 223;224;225;228;229;230;231;233;234;235;236;237;238;239;240;249;250;251; 252;261;265;266;267;268;269;272;274;275;276;279;281;284;287;288;291;294; 295;296;297;298;299;300;301;302;303;304;305;306;326;327;328;329;337;338; 339;341;345;346 Photic zone;31;48;49;50;51;54;57;270;281 Photoacclimatization;48;52 Photoinhibition;38;48;53 Photosynthesis;1;2;11;12;14;21;22;23;24;27;28;29;34;35;36;38;41;43;44;45;46;48;49 ;50;53;54;55;56;57;58;59;60;61;62;63;64;65;89;92;94;101;102;103;105;106;119; 145;146;192;208;250;265;273;275;279;281;282;284;286;290;292;294;295;296; 298;299;305;312;319;320;322;326;328;344;345 Physical mixing;156;166;170 Physiology;1;9;25;34;40;49;57;63;64;73;93;94;96;102;112;119;149;150;160;169;170 ;212;297;331;332 Phytoplankton;8;15;16;17;21;25;31;33;36;37;38;40;42;46;47;48;49;50;51;52;53;54; 55;56;57;58;59;61;63;64;65;66;73;83;85;87;88;89;92;93;97;98;104;105;106;108; 133;135;136;138;139;141;142;143;152;153;154;155;157;159;160;161;162;163; 164;165;166;168;169;170;171;173;174;175;177;187;190;192;202;210;214;221; 231;241;242;243;251;270;283;286;287;288;290;292;293;295;298;299;301;302; 303;305;313;331;332;333;335;336;337;343;345 Pigments;34;36;38;43;48;53;55;58;60 Plant communities;vii;1;4;8;9;12;14;20;21;22;23;26;46;47;48;49;59;63;65;103;106;140; 146;204;211;214;218;241;243;246;248;249;250;265;324;331;332;333;334;336; 339;343;344;345 Porewater;11;272;275;279;281;301;310;311;319;320;339;341 Porosity;255 Primary producers;v;vii;1;10;12;16;18;20;21;22;46;47;58;62;68;69;73;75;76;83;87;97;98; 99;103;129;135;149;176;177;204;214;215;217;219;221;223;224;225;227;228;229; 233;240;241;242;243;247;249;253;258;263;264;270;271;272;274;275;276;280; 283;287;290;291;292;294;295;297;308;326;331;332;334;336;337;338;340;343; 344 Primary production;1;2;4;6;7;9;12;14;15;17;47;59;60;68;69;73;85;87;99;101;102;104;132; 135;142;145;150;151;153;160;161;162;163;164;165;166;167;176;178;179;180; 182;184;185;187;189;200;202;203;209;210;214;243;244;245;250;253;254;260; 270;283;289;291;293;294;296;299;301;305;308;310;314;317;325;326;327;332; 334;335;337;338;343;344 Pycnocline;158;326 Recycling;v;9;176;179;180;196;200;202;205;214;215;240;241;247;249;254;255;272; 281;295;331;334;337 Redox;9;10;11;93;94;260;262;265;268;275;290;341 Refractory pool of detritus

300

13;14;73;99;216;217;219;221;227;228;229;230;231;232;233;234;235;238;240;241 ;242;244;246;248;249;250;254;260;261;263;266;290;295;297;337;338;343 Refuge;190;191;195;197;198 Relative surface area;1 Remineralization;8;9;13;33;166;167;168;169;214;215;216;217;227;229;230;231;235; 236;237;238;240;241;242;244;246;247;249;272;274;281;283;287;288;290;300; 311;315;317;325;327;335;337;338;339;340 Respiration;15;21;32;35;43;44;51;55;56;62;93;94;96;261;273;276;279;281;282;291; 295;319;322 Resuspension;10;11;88;112;134;145;157;159;175;255;258;266;272;279;314;333;339 ;341 Reynolds number;30 Rhizosphere;11;12;56;105;273;274;275;276;290;301;302;306;322;323;326;328;340; 345 Root;8;11;13;15;17;27;33;56;73;86;93;94;96;105;106;109;111;226;232;233;234;235; 237;238;239;240;259;261;263;264;267;268;273;274;275;276;279;290;292;303; 322;323;324;334;337;340;341 Rooted;vii;1;7;10;11;23;27;47;55;56;64;93;110;112;113;116;120;121;131;144;242; 289;331;338;339;340;341 Rooted macrophytes;vii;1;7;10;11;110;112;131;242;289;339 Salinity;3;25;78;94;103;135;140;146;147;203;211;232;235;300;328 Salt marshes;68;106;135;142;143;144;146;258;260;268;298 Sampling strategy;108 Scaling;2;35;39;40;41;43;44;46;60;64;66;208;297;345 Sea urchins;98;103;105;178;181;183;206;207;210;211 Seagrass;v;vii;8;9;10;11;13;15;16;17;23;25;27;28;32;33;42;46;47;50;55;56;59;60;61; 62;63;64;65;68;73;74;75;76;77;78;80;81;82;83;85;87;88;89;90;91;92;93;94;96;97; 98;99;100;101;102;103;104;105;106;107;109;111;121;133;143;159;160;163;171; 172;175;176;177;178;179;180;181;182;183;184;185;186;187;189;190;191;195; 197;198;199;200;202;204;205;206;207;208;209;210;211;212;214;219;224;225; 226;227;228;229;230;231;232;233;234;235;236;239;240;241;242;243;248;249; 250;251;252;253;258;259;260;264;265;266;268;270;271;272;273;274;275;276; 277;279;280;281;286;287;289;290;292;293;295;296;297;298;299;300;301;302; 303;304;305;306;322;323;324;327;328;331;332;333;334;335;336;337;338;340; 341;343;345;346 Seagrass decline;75;76;77;83;85;98;177;187;189;195;297 secondary production;150;160;174;195;202;204;295 Sediment;v;6;9;10;11;13;14;15;16;17;18;20;21;25;26;27;31;32;33;53;54;55;56;58;59 ;60;62;63;65;69;73;75;88;93;94;95;96;100;101;102;105;113;115;116;120;123;134 ;135;145;146;168;171;176;177;182;188;190;192;195;196;197;202;203;205;208; 209;231;241;248;249;250;253;254;255;256;257;258;259;260;261;263;264;265; 266;267;268;271;272;273;274;275;276;279;280;281;282;283;284;285;286;287; 288;289;290;291;292;293;294;295;296;297;298;299;300;301;302;303;304;305; 306;308;310;311;312;313;314;315;316;317;318;319;320;321;322;323;324;325; 326;327;328;329;333;338;339;340;341;345 Sedimentation;2;10;13;33;48;52;124;125;159;160;294;308;313;339 Sediment-water interface;10;59;274;281;283;286;300;305;341

301

Self-shading;35;48;192;282;286 Senescence;26;34;48;51;55;111;120;282;291 Seston;155;156;157;159;171;173;221;250;262 Settling;53;109;110;113;119;120;121;122;143;259 Shading;21;35;48;85;87;92;93;98;129;177;187;190;192;203;280;282;286;288;289; 290;293;294;333 Shape;1;10;20;21;22;28;29;34;35;38;43;52;55;57;120;121 Shelter;73;110;180;190;196 Sink;12;43;73;106;134;135;152;160;168;227;253;272;275;276;279;280;282;289;290; 294;295;298;320;334 Sinking rate;2;8 Size;1;5;10;20;22;23;25;26;27;28;29;30;31;34;35;38;39;40;41;43;44;45;46;48;52;55; 57;58;60;61;62;63;66;109;111;114;116;120;121;123;149;150;156;160;161;174; 190;196;200;205;217;219;227;228;231;232;233;235;236;237;238;267;306;334 Sloughing;112;117;118;119;140;142;143 Sporulation;117;122;124;125;126;127;141 Stability;20;21;26;28;57;209;258;285 Stratification;153;154;156;158;159;161;165;175;193;199 Substratum;21;25;26;27;30;31;32;33;47;52;58;59;100;155;192;196 Sugars;215 Sulphide;20;94;96;147;193;195;202;206;261;266;294;306 Suspension feeders;v;21;149;150;151;152;153;154;155;156;159;160;162;163;164;165;166; 168;169;171;172;175;259;335 Taxonomy;23;27;308 Temperature;3;20;24;32;58;70;77;83;94;103;106;120;126;149;150;151;175;186;231; 232;235;251;282;288;346 Terrestrial plants;24;43;44;215;276 Thallus;16;28;29;34;37;38;44;64;119;120;121;122;123;208;332;345 Thickness;1;28;29;30;32;33;35;37;38;40;43;55;56;64;94;258;312 Threshold velocity;109;114;116 Tide;3;8;76;78;88;101;104;106;110;112;113;116;122;128;129;143;159;203;205;254; 296;334 Top-down control;164;178;179;336 Toxicity;85;86;87;94;100;107;276;305;333 Transects;124;262;305;329 Transport;v;2;5;6;7;8;12;13;14;15;21;22;30;31;32;33;35;38;41;54;55;56;57;58;59;75; 94;108;109;110;111;112;113;114;115;116;118;119;122;126;127;129;130;131;132; 133;134;135;136;138;139;140;141;142;143;144;145;152;166;173;250;254;261; 271;311;312;326;331;334;335;341;343;345 Trophic levels;172;290;292;335;336;337 Turbidity;21;76;78;79;104;140;143;160;163;193 Turbulence;30;32;33;65;156;157;159;160;165;169;174;212;259 Unicellular algae;1;23;37;43;44;59 Upwelling;68;161;170 Viscosity;30;31;32;33 Water motion;22;26;30;61 Water residence time;6;75;77;83;85;87;97;161;162;163;164;271;295;333;336;344

302

Watershed;17;71;75;77;83;103;106;177;198;199;211;271;282;295;305 Wave;3;32;34;62;73;112;159;186;254;256;259;266;302;333 Wetlands;98;145;146;261;267 Zooplankton;47;48;152;154;164;168;172

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AQUATIC ECOLOGY BOOK SERIES http://www.springeronline.com/sgw/cda/frontpage/0,11855,4-40109-69-33114010-0,00.html

Series Editor: Prof. Robert G. Wetzel, Department of Environmental Sciences and Engineering, The University of North Carolina, Chapel Hill, North Carolina 27599-7431, USA. E-mail: [email protected] Current volumes in this series

Volume 1

Ecology of a Glacial Flood Plain J.V. Ward, U. Uehlinger Hardbound, ISBN 1-4020-1792-8, December 2003, 320 pp.

Volume 2

The Influence of Primary Producers on Estuarine Nutrient Cycling S.L. Nielsen, G.T. Banta, M.F. Pedersen Hardbound, ISBN 1-4020-2638-2, 2004

Forthcoming volumes:

Harmful Cyanobacteria J. Huisman, H. C.P. Matthijs, P. M. Visser Ecological History of the Lowland River Basins of Rhine and Meuse P.H. Nienhuis Ecology of Russian Rivers A.V. Zhulidov, V.V. Khlobystov, R.D. Robarts, A.A. Ischenko Dr. J. Headley Biology of Polluted Rivers T.E.L. Langford

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