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.

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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 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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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.

Height above bottom (cm)

Plant bound nutrient transport

95

90-120

60-90

30-60

Zostera Chaetomorpha

0-30

Ulva

0

2

4

6

8

kg wwt.

Figure 1. Vertical distribution of the transported plants in Venice Lagoon Malamocco site, May 1995. (Flindt et al. 1997a, reprinted with permission from Elsevier)

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, P80%) of the seagrass communities compiled lose a relatively modest percentage (0.10; microphytobenthos vs. seagrasses, P